Acceptable Phosphorus Concentrations in Soils and Impact ...

Acceptable Phosphorus Concentrations in Soils and Impact on the Risk of Phosphorus Transfer from Manure Amended Soils to Surface Waters

A Review of Literature for the Manitoba Livestock Manure Management Initiative Phase 1 of MLMMI Project #02-HERS-01

May 1, 2003

Don Flaten1, Ken Snelgrove1, Ian Halket1, Kathy Buckley2, Grant Penn2, Wole Akinremi1, Brian Wiebe1, Ed Tyrchniewicz1

1University of Manitoba and 2AAFC Brandon Research Centre Authors are listed in order of presentation within the report

Table of Contents Executive Summary ....................................................................................................................... i

Background .................................................................................................................................. i Amounts of P Discharged to Environment from Agriculture and Other Sources ................ i Methods of Reducing P Discharge from Livestock..................................................................ii Impact of Soil Type, P Management, Landscape and Climate on P Retention and Release

by Soil .............................................................................................................................iii Legislation and Regulation Regarding P Management in Other Jurisdictions ...................vi Summary....................................................................................................................................vii

Foreword.....................................................................................................................................viii Introduction .................................................................................................................................. ix

Concerns for Phosphorus in the Environment........................................................................ ix Implementing Measures to Reduce the Risk of P Contamination of Water .........................x Factors Affecting the Risk of Phosphorus Transfer from Agricultural Land: An

Overview..........................................................................................................................x References.................................................................................................................................xiii

1 Relative Magnitude of the Phosphorus Discharged ............................................................1 1.1 Chapter Summary...............................................................................................................1 1.2 Introduction .........................................................................................................................2 1.3 Phosphorus in the Environment – Why P is Important ..................................................3 1.4 Forms of P in Water............................................................................................................ 3 1.5 P in Manitoba’s Waters ......................................................................................................5 1.6 Sources of P in Manitoba..................................................................................................12

1.6.1 Atmospheric P Sources..................................................................................................12 1.6.2 Terrestrial Sources .........................................................................................................14

1.6.2.1 Urban P sources .......................................................................................................14 1.6.2.2 Natural P sources .....................................................................................................16 1.6.2.3 Agricultural P sources .............................................................................................17

1.6.3 In-Stream Processes.......................................................................................................27 1.7 Temporal Distribution of P in Manitoba’s Waters........................................................28 1.8 Conclusions ........................................................................................................................32 1.9 Abbreviations and Definitions .........................................................................................34 1.10 References ..........................................................................................................................35

2 Reducing Phosphorus Contribution from Animal Agriculture in Manitoba .................41 2.1 Chapter Summary.............................................................................................................41 2.2 Introduction .......................................................................................................................42 2.3 Phosphorus Excretion by Livestock and Poultry...........................................................43 2.4 Phosphorus Compounds in Manure – Terminology and Analysis...............................45 2.5 Feed Treatments, Additives and Novel Ingredients.......................................................46

2.5.1 Addition of Phytase to Diets..........................................................................................46 2.5.2 Pre-Treatment of Feed ...................................................................................................51 2.5.3 Low Phytate Feed Ingredients .......................................................................................52 2.5.4 Augmentation of Endogenous Phytase in Livestock.....................................................54 2.5.5 Enhancement of Phytase Activity in Feed.....................................................................55

2.6 Feeding and Management of Poultry and Livestock to Reduce P Excretion ..............56

2.6.1 Feeding and Management of Poultry.............................................................................56 2.6.1.1 Phosphorus requirements.........................................................................................56 2.6.1.2 Feeding and management strategies ........................................................................57 2.6.1.3 Summary of strategies for P decreases ....................................................................59

2.6.2 Feeding and Management of Swine ..............................................................................59 2.6.2.1 Phosphorus requirements.........................................................................................59 2.6.2.2 Feeding and management strategies ........................................................................61 2.6.2.3 Summary of strategies for P decreases ....................................................................64

2.6.3 Feeding and Management of Ruminants .......................................................................65 2.6.3.1 Phosphorus requirements.........................................................................................65 2.6.3.2 Feeding and management strategies for dairy cattle ...............................................68 2.6.3.3 Feeding and management strategies for beef cattle.................................................69 2.6.3.4 Feeding and management strategies for small ruminants........................................70 2.6.3.5 Summary of strategies for P reduction and management for ruminants .................70

2.7 Manure Treatment and Utilization Technologies to Manage Excreted Phosphorus .70 2.7.1 Manure Treatment Alternatives.....................................................................................71

2.7.1.1 Solids removal .........................................................................................................71 2.7.1.2 Natural sedimentation..............................................................................................71 2.7.1.3 Mechanical separation .............................................................................................72 2.7.1.4 Nutrient inactivation ................................................................................................73 2.7.1.5 Combined treatments ...............................................................................................79 2.7.1.6 Biological phosphorus removal ...............................................................................80 2.7.1.7 Aeration systems for phosphorus removal ..............................................................81 2.7.1.8 Natural systems of phosphorus removal..................................................................81 2.7.1.9 Solid manure treatments ..........................................................................................83

2.7.2 Strategies to Recover and Use Excess Phosphorus .......................................................84 2.7.2.1 Manure as a feed ingredient.....................................................................................84 2.7.2.2 Biological tissues as hyperaccumulators of P..........................................................84

2.7.3 Potential Problems Associated with Adoption of Treatment and Utilization Strategies 85

2.8 Conclusions ........................................................................................................................86 2.9 Abbreviations and Definitions .........................................................................................88 2.10 References ..........................................................................................................................89

3 Impact of Soil Type, P Management, Landscape and Climate on P Retention and Release by Soil ...........................................................................................................................103

3.1 Chapter Summary...........................................................................................................103 3.2 Introduction .....................................................................................................................105 3.3 Reactions of phosphorus in soil......................................................................................106

3.3.1 Adsorption and Desorption Reactions.........................................................................106 3.3.1.1 The Freundlich isotherm........................................................................................109 3.3.1.2 The Langmuir isotherm .........................................................................................110 3.3.1.3 Gunary isotherm ....................................................................................................111 3.3.1.4 Kinetics of adsorption............................................................................................111 3.3.1.5 Effect of organic matter on sorption and desorption .............................................113

3.3.2 Precipitation and Dissolution Processes ......................................................................114 3.3.3 Mineralization and Immobilization .............................................................................116

3.3.4 Plant Uptake ................................................................................................................117 3.3.5 Soil Factors Affecting Phosphorus Retention and Release .........................................117

3.3.5.1 Soil texture.............................................................................................................117 3.3.5.2 Carbonate content ..................................................................................................118 3.3.5.3 Soil pH...................................................................................................................119 3.3.5.4 Fe and Al oxides ....................................................................................................120 3.3.5.5 Redox potential (effect of flooding) ......................................................................121 3.3.5.6 Soil organic matter.................................................................................................121

3.3.6 Inorganic and Organic P Amendment Reactions in Soil .............................................122 3.3.7 Management Factors Affecting Phosphorus Retention and Release...........................126

3.3.7.1 Application method, placement and timing...........................................................126 3.3.7.2 Application rate .....................................................................................................127 3.3.7.3 Crop choice............................................................................................................ 127 3.3.7.4 Conservation tillage ...............................................................................................128 3.3.7.5 Vegetative buffer strips and riparian areas ............................................................128 3.3.7.6 Constructed wetlands.............................................................................................128 3.3.7.7 Grazing intensity and streambank protection ........................................................129 3.3.7.8 Nutrient content of manure....................................................................................129

3.4 Mode of Phosphorus Transport to Surface Waters .....................................................130 3.4.1 Surface Pathways of P Loss.........................................................................................134 3.4.2 Subsurface Pathways of P Loss ...................................................................................136 3.4.3 Snowmelt Runoff and P Loss ......................................................................................137

3.5 Soils of Manitoba.............................................................................................................140 3.5.1 Central District ............................................................................................................141 3.5.2 Eastern District ............................................................................................................143 3.5.3 Interlake District ..........................................................................................................144 3.5.4 Mid-Western District ...................................................................................................145 3.5.5 Parkland District ..........................................................................................................147 3.5.6 Western District ...........................................................................................................149

3.6 Methods Of Assessing Risk Of Phosphorus Transfer From Soil To Water ..............150 3.6.1 Soil Test Phosphorus ...................................................................................................151 3.6.2 Balance of P Removal and Addition ...........................................................................155 3.6.3 Degree of Phosphorus Saturation ................................................................................155 3.6.4 P Index .........................................................................................................................157 3.6.5 Phosphorus Transport Models .....................................................................................165

3.6.5.1 FHANTM ..............................................................................................................165 3.6.5.2 NLM ......................................................................................................................168 3.6.5.3 EFPEM ..................................................................................................................168

3.7 Conclusions ......................................................................................................................169 3.8 Abbreviations and Definitions .......................................................................................171 3.9 References ........................................................................................................................172

4 Legislation and Regulations Regarding Phosphorus Management in Other Jurisdictions...............................................................................................................................185

4.1 Chapter Summary...........................................................................................................185 4.2 Introduction .....................................................................................................................187 4.3 Background to Environmental Regulation in Agriculture..........................................187

4.4 Review of Regulations relating to Phosphorus Management in Agriculture ............189 4.4.1 Environmental Regulation Affecting Intensive Livestock Operations in Canada ......190

4.4.1.1 Alberta ...................................................................................................................190 4.4.1.2 Saskatchewan.........................................................................................................191 4.4.1.3 Ontario ...................................................................................................................192 4.4.1.4 Quebec ...................................................................................................................195

4.4.2 US Jurisdictions...........................................................................................................197 4.4.2.1 North Carolina .......................................................................................................198 4.4.2.2 Iowa .......................................................................................................................199 4.4.2.3 Minnesota ..............................................................................................................200 4.4.2.4 Maryland................................................................................................................202

4.4.3 European Jurisdictions.................................................................................................202 4.4.3.1 The Netherlands.....................................................................................................203 4.4.3.2 Denmark ................................................................................................................205 4.4.3.3 England ..................................................................................................................206 4.4.3.4 Ireland ....................................................................................................................207

4.5 Conclusions ......................................................................................................................208 4.5.1 Summary......................................................................................................................208

4.5.1.1 National level regulations of intensive livestock operations .................................208 4.5.1.2 Phosphorus-based regulations ...............................................................................208

4.5.2 Assessment of Regulatory Approaches .......................................................................210 4.5.2.1 Assessment criteria ................................................................................................210 4.5.2.2 Linking technical phosphorus considerations and regulations ..............................211 4.5.2.3 Concluding observations .......................................................................................212

4.6 Abbreviations...................................................................................................................213 4.7 References ........................................................................................................................214

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Executive Summary

Background This review was initiated by a request for proposals from the Manitoba Livestock Manure Management Initiative (MLMMI) in the fall of 2001. According to recent studies completed by Manitoba Conservation, there are increasing concentrations of phosphorus (P) in the rivers draining major watersheds in Southern Manitoba. Agriculture has been suggested as one contributor to the P loads that reach water bodies, particularly from diffuse sources. Manure P has been identified as a significant source of soil P enrichment in areas of high density of confined livestock operations. Some of this phosphorus may reach important water bodies such as Lake Winnipeg. As a result of these concerns for water quality, the Government of Manitoba has recently announced a “Water Strategy“ (April 21, 2003), in which the management of manure nutrients receives considerable attention.

Existing information regarding the behaviour of phosphorus in agricultural production systems is an important tool for developing agricultural management practices that will minimize the risk of P transfer to water bodies. It is also important to review all pertinent information to clearly identify the knowledge gaps that should be addressed in future research efforts. Future research should aim at describing the scope of the problem, the timeframe involved, the information necessary to establish sound codes of practice and the best management practices to reduce the potential impact of a sustainable and expanding animal based agriculture in Manitoba.

Objectives The objective of this project was to review and adapt the bodies of existing knowledge on the role and fate of P in livestock and crop production systems specifically relevant to Manitoba, to identify the gaps in knowledge and briefly describe what should be done, including:

1. Review and tabulate information on the amounts of phosphorus discharged to the environment from agriculture (including livestock and crop production) and other users of land and water.

2. Review the knowledge on the methods of reducing phosphorus contribution from animal agriculture in Manitoba, including reduction of excreted P and manure treatment.

3. Review the literature on the impact of soil type, P management, landscape and climate on P retention and release from soil to water.

4. Review legislation and regulation regarding P management in other jurisdictions.

Amounts of P Discharged to Environment from Agriculture and Other Sources Accumulations of excessively high concentrations of P in soil create an increase in the risk of P transfer into surface water bodies and subsequently, the risk of declining water quality due to eutrophication (e.g., increased algal growth, oxygen depletion, fish kills and release of algal toxins). The sources of P in surface water are diverse, ranging from natural sources such as vegetation, erosion, wildlife, snow, rain and phosphine gas to human sources such as agriculture and direct discharge from industrial, municipal and septic waste systems. However, according to recent studies by Manitoba Conservation, the concentration of P is increasing in many of Southern Manitoba’s rivers and streams, in regions where human activities, including

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agricultural production, are most intensive. Furthermore, these same studies indicate that agricultural land is a very large source of P loading in Manitoba. For example, although the Red River contributes a relatively small portion of the water to Lake Winnipeg, the Red River accounts for 73% of the P loading in the lake. Within the Red River’s contribution of P, approximately 60% of that load is from the U.S. portion of the watershed and 17% of the load is estimated to be a result of present-day agricultural activities within Manitoba.

Livestock manure is an important and growing potential source of P in the environment. For example, Statistics Canada estimates that manure nutrient production (i.e., animal units) by beef cattle and pigs in Manitoba has risen by 35 and 65 percent, respectively over the last ten years, with beef cattle responsible for the largest absolute increase in manure production during this period (+228,000 animal units). As a result of the dramatic growth in Manitoba’s livestock industry, two municipalities in Manitoba (La Broquerie and Hanover) now have the sixth and eleventh highest densities of livestock in Canada, respectively. However, across the province as a whole, manure P accounts for only 19% of total P removed by the province’s crops.

The proportion of P in rivers and streams that is contributed by livestock production is, however, very difficult to quantify. Some of the major deficiencies in our knowledge of nutrient loading in Manitoba’s surface waters include:

• The dominant role played by snowmelt flows in prairie hydrology regimes suggests that most of the P load to rivers and lakes is supplied during the spring. Despite the extensive body of knowledge on soil P processes and on snowmelt runoff, our understanding of P transport from land to water during snowmelt is limited. More research into snowmelt phosphorus runoff processes in Manitoba is required if effective land management policies are to be adopted by the province to deal with the annual P loads to Lake Winnipeg and its upstream rivers.

• Most of Manitoba’s water quality data is based upon a periodic sampling schedule (e.g., every month), rather than on an event sampling schedule (e.g., during and after runoff events), so the major nutrient loading events are not well monitored.

• Analysis of nutrient loading data for the Red River watershed shows that the 40% of the load emanating from within Manitoba is evenly split between the watershed upstream of Winnipeg, and Winnipeg and its downstream environs. More research is needed to establish the sources of the load in and downstream of Winnipeg.

• Estimates of the agricultural loading in the Red River basin are based on inadequate information concerning appropriate export coefficients for calculating P transported from various types of land use and management under local conditions.

• Processes that temporarily retain P within Prairie rivers, streams and lakes (e.g., sedimentation, biological immobilization) are not well understood and may be an important source of error in estimating P loading within parts of the overall watershed.

Methods of Reducing P Discharge from Livestock One of the strategies for reducing the risk of P transfer from agricultural land to surface water is to reduce the amount of manure P that is produced by livestock. The efficiency of utilizing P in feed could be improved, for example, by minimizing any surplus P in the diet; conventional genetic selection; innovative improvements to feed formulation, pre-treatment, mixing, processing and feeder design; phase feeding to match the ration to the requirements of sexes or

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different growth phases; improved attention to nutrient balances; and overall improvements to productivity for the same amount of P fed or excreted. One of the largest obstacles to improved utilization of feed P is the high proportion of feed grain P that is in the form of phytate, a compound that is largely indigestible by pigs and poultry. Therefore, P utilization by monogastric animals can be improved by using feed sources with low concentrations of phytate (Highly Available P or HAP feed grains), adding phytase enzymes to feed, or modifying the genetics of the monogastric animal to introduce phytase into the saliva (e.g., the “Enviropig” developed at the University of Guelph).

Innovative methods of redistributing, recovering or immobilizing the P in livestock manure form another strategy for reducing P loading onto agricultural land. The solids in liquid manure generally contain a relatively high proportion of the manure’s P content; therefore, the P content of the liquid portion is reduced when solids are removed by natural sedimentation or mechanical separation. The P in manure can be treated chemically in order to reduce the risk of P movement into waterways. For example, aluminum, iron or magnesium may be added to precipitate the P; zeolites may be added to adsorb P; or polyacrylamide polymers may be added to encourage flocculation and settling of P in liquid manure. The P in manure can also be immobilized biologically, using anaerobic or aerobic decomposition reactors, aquatic algae or plants, or constructed wetlands. However, the impact of Manitoba’s relatively cold climate and short summers may be an important limitation to these biological methods of P removal.

Livestock management practices that could be promoted and adopted immediately for reducing the amount of manure P produced by livestock include:

• Improvements to overall efficiency of feed use and reduction of feed waste • Reduction in dietary P overformulation and increased use of phase feeding

Further research is warranted on the effects of the following livestock production and manure management practices on the risk of P transfer from land to water:

• Effects of strategies designed to improve the availability of phytate P to monogastric animals on P solubility in feces and elimination of excess plasma P via urine, including:

o phytase supplementation in feeds o feeding "highly available P" corn, soybean and barley varieties o genetic selection for phytate utilization o combinations of the above

• Effects of strategies designed to redistribute, recover or immobilize P on P recovery, crop nutrition and soil properties under Manitoba conditions, including:

o solid separation and composting o chemical additives such as alum, magnesium, zeolites, and polyacrylamide

polymers to retain P in an easily recoverable form o biological immobilization of P

Impact of Soil Type, P Management, Landscape and Climate on P Retention and Release by Soil Soils do not possess an infinite capacity to retain P. Therefore, for long term environmental and agronomic sustainability, P application should not exceed P removal and the siting and density of livestock enterprises should be determined accordingly. However, the balance between P application and removal, alone, is not very useful as an indicator of the risk of P transfer to

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water. For example, P balance alone, does not consider the availability of the P in the added material; nor does steadfast maintenance of a P balance allow for the reduction of pollution risk from a high P soil or the improvement of P fertility on a low P soil. Therefore, a soil’s current concentration of P and a soil’s capacity to retain P must also be considered.

Simple measures of “soil test” or extractable P in soils provide a general indication of how much P is in the soil, but do not reveal the soil’s capacity to retain additional P or the degree to which that retention capacity is already saturated. Therefore, although the relationship between soil test P and the concentration of P in runoff is often consistently related within a soil type or within a group of closely related soil types, the relationship is much less predictable across different soil types. Therefore, if a soil with a low retention capacity for P is used to establish a low environmental threshold for soil test P, agricultural production on soils with a high retention capacity for P is unnecessarily constrained. Contrarily, if a soil with a high retention capacity for P is used to establish a high environmental threshold for soil test P, environmental quality will be impaired by soils with a low retention capacity for P.

A soil’s capacity to retain additional P is an important factor for determining the sustainable loading rates for manure P over the short term (i.e., over a few months or over several years). Retention mechanisms in soil include adsorption onto soil surfaces, precipitation into soil solids, and biological immobilization into soil organisms and organic matter. The characteristics of soil that determine its P retention capacity include: type and amount of exchangeable cations, iron and aluminum oxide content, types and amounts of clay minerals, calcium carbonate content, and pH. Manitoba soils are usually regarded as having a high capacity to retain P because of their high pH and carbonate content. However, this assumption may be unjustified for three reasons: the strength of P retention by carbonates is not nearly as strong as that by iron and aluminum; carbonates are not found at the surface for 65% of soils in Manitoba; and 25% of Manitoba’s soils now have a neutral to slightly acid pH. Furthermore, no measurement techniques or thresholds for P retention capacity and degree of P saturation have been developed for soils in the Northern Great Plains region of North America, or for the landscapes, climate and agricultural management practices that are typical for Manitoba.

One of the reasons why local landscapes, climate and management practices are important is that the loss of P from land to water is influenced greatly by the transport processes that move soil and water. The three main modes of P movement are erosion of particulate P, runoff of dissolved P and the leaching of P below the root zone. Particulate P accounts for 60-90% of P loss from agricultural land in the U.S.; however, a few recent studies in Western Canada have identified much lower proportion of P losses in the particulate form. Dissolved P may account for the majority of P loss in the Canadian Prairies because of our relatively dry climate, the relatively high proportion of P that may be leaching from vegetative material, and the high proportion of runoff that is dependent on snowmelt. Although particulate P losses can be controlled quite readily through the control of soil erosion, the control of dissolved P losses is much more challenging. Losses of P through leaching have never been considered an agronomic concern, but over the last thirty years, researchers have gathered substantial evidence that small amounts of P can be lost to groundwater and tile drains, especially if the soil is cracked or lacking in oxygen. The main concern for P leaching is not its effect on groundwater quality, but rather its effect on the surface water that it will likely enter in the future.

Due to the combination of source and transport factors that affect the risk of P movement from land to water, scientists have developed two strategies for integrating those factors into an

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overall, site-specific risk assessment. One strategy is a semi-quantitative “phosphorus index” or “phosphorus site index.” Various source factors (e.g., soil test P or degree of P saturation, crop residue P, etc.) and transport factors (e.g., erosion and runoff risk) are assigned a numerical value. Then, the ratings for each factor or group of factors is added or multiplied to generate an overall risk score that is then used to estimate the overall risk of P loss as high, medium or low. These “P indexes” were originally developed for extension, not scientific or regulatory purposes, so their use for the latter purposes is frequently questioned. Furthermore, these indexes usually require significant local calibration; recent efforts to use this strategy to quantify P losses in Minnesota, for example, resulted in significant underprediction of P losses in the American portion of the Red River Valley.

Another strategy to integrate the effect of P source and transport factors into an overall risk assessment is to develop a detailed, quantitative, process-based model for P loss. Alberta Agriculture Food and Rural Development began to build such a model, but they have found that the development of this model is extremely challenging and expensive.

Therefore, we recommend the Phosphorus Index as an effective method of assessing the risk of P loss from Manitoba soils. To aid in the development of an effective P index for Manitoba, the following knowledge gaps are identified:

• the phosphorus adsorption capacity (PSC) of representative Manitoba soils should be determined

• the degree of phosphorus saturation (DPS) of these soils must be assessed and a model developed as a tool for estimating both the PSC and DPS from soil physico-chemical properties

• the role of crop residues and vegetative covers as sources of P under Manitoba conditions where runoff is relatively limited and mostly confined to spring snowmelt

• field scale and watershed studies should be conducted on representative Manitoba soils and landscapes to determine the role of snowmelt runoff P and sheet or overland flow on P transport

Other knowledge gaps that deserve further investigation include:

• the role of soil CaCO3 in the retention of manure P in comparison to fertilizer P • the role of cation composition and content of manure on the retention of manure P by the

soil • the role of organic acids in manure on the retention of manure P by the soil • the comparative movement of organic and inorganic P fractions of manure in soils • long-term fate (2 years of more) of retained manure P (i.e, aging and solubility of manure

P in Manitoba soils with time)

While tools are being developed to predict the quantity of P released from Manitoba soils to water, several general types of best management practices should be recommended to farmers, immediately, including:

• avoid applying P at rates that exceed twice the annual withdrawal of P by crops • avoid applying manure onto soils that test very high in P or where transport factors such

as steep slopes or close proximity to waterways create a high risk of water contamination

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• inject or incorporate manure where possible to reduce the loss of P through runoff and where surface application cannot be avoided, such as on perennial forages, manure should be applied after spring snowmelt

• when constructing new livestock facilities, keep in mind that the soil does not have an infinite capacity to retain P and, therefore, for long term sustainability, the density of livestock production should not exceed crop withdrawal of P

Legislation and Regulation Regarding P Management in Other Jurisdictions Public agencies can use three main types of policy instruments to minimize the risk of P transfer from agricultural land to water: regulatory (either regulating the agricultural practices, themselves in a preventative way or regulating according to management performance, after the fact), economic (through taxes and/or subsidies) and extension/education (voluntary, participatory activities). The criteria for selecting and implementing any of these instruments include: the availability of scientific rationale as a basis; equity among farmers, between farmers and non-farmers, and between polluters and non-polluters; efficiency of implementing and enforcing the policy; political acceptability; environmental benefits; and effects on economic competitiveness.

Within Canada, Alberta and Saskatchewan have not implemented regulations that target manure P management, yet. Ontario's Nutrient Management Act (2002) was developed to apply to N and P from manure, fertilizers, and biosolids; however, that Act is currently being revised due to difficulties in implementation. In its current form, the Act sets a soil test P threshold at which a site-specific P index is triggered for determining an appropriate separation distance from water for manure application. Quebec's newest version of “Agricultural Operations Regulation” is focused on nutrient balance for N and P. In cases where excess P might be a concern, the regulation requires a detailed plan of how those concerns will be alleviated, enlisting the producers to work towards the environmental objectives, rather than forcing compliance with specific agricultural practices. However, in areas where the livestock density is high, overall, the construction of additional pork production facilities, in particular, is not permitted.

In the U.S., Iowa, Minnesota and Maryland utilize the site-specific P index as a regulatory instrument, a practice which, as mentioned previously, has generated some technical concerns. A soil test threshold set at 60 and 120 mg Olsen-extractable P/kg soil near and away from waterways, respectively, currently governs Minnesota's regulations on P. The Netherlands uses a Minerals Accounting System (MINAS) to regulate the N and P applied as manure, fertilizer, and biosolids. Those regulations, which are currently being revised, require a detailed accounting of N and P imports and exports from each farm and assess a levy for any nutrient surpluses that occur in excess of predetermined limits for risk of nutrient loss.

Regulatory recommendations for Manitoba include:

• Given that about 60% of phosphorus loadings in the Red River originate from US sources, efforts need to be expended by the Government of Manitoba to work with US jurisdictions to reduce phosphorus loadings before they enter Manitoba.

• There is a need for a more collaborative approach among government departments and agencies in the development of a phosphorus management strategy for Manitoba, especially among Manitoba Agriculture and Food, Manitoba Conservation, Environment

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Canada’s National Water Research Centre, and Fisheries and Oceans Canada’s Freshwater Institute.

• Develop a comprehensive approach to nutrient management, with manure and P as components.

• Invest in research to reduce P in manure (eg. phytase management and other feed additives), before regulating P.

• Monitor and regulate on a watershed basis rather than an individual farm basis, focusing first on regions with high nutrient loads.

• Implement soil phosphorus regulations that include a voluntary education program on best management practices within a regulatory framework.

• Planning of initial siting of ILOs should be a high priority. • Evaluate regulatory tools to ensure the choice is scientifically sound, targeted to

ameliorate environmental concerns while minimizing unnecessary constraints on agricultural activities

• Government has to commit sufficient resources for monitoring and enforcing the regulations for regulations to be effective.

• Legislation and regulation regarding P management should be introduced cautiously to ensure environmental protection without undue hardship to the agricultural industry

Against these recommendations are some additional regulatory cautions including:

• Tighter environmental regulations will impact small-scale farm operators more negatively than large-scale farm operators. Larger farm operations are in a better position to have the financial resources, technical knowledge, and human resources to know and follow increasingly complex regulations.

• Regulations add to the costs of production and decrease the competitiveness of the agricultural sector. The Province of Manitoba should not “get too far out in front” with its regulation of the livestock industry relative to competing jurisdictions in Canada and the US.

Summary Manure is a valuable source of P for crop production. However, applying too much manure P, especially in the wrong place, is not only agronomically wasteful, but potentially harmful to the environment. As a result of recent increases in Manitoba’s production of livestock, the risk of transfer of manure P from agricultural land to surface water is probably increasing. However, the amounts of P that are discharged from livestock production and other agricultural activities are very difficult to determine due to a lack of hydrological data and, more important, a lack of data on the transfer of P from soil to water. Part of this problem is due to the highly variable impact of soil type, P management, landscape and climate on P retention and release by soil, information that is not well documented for Manitoba conditions. Fortunately, livestock producers have a wide variety of potential techniques for reducing P discharge from their operations; however, the technical and economic merit of these techniques is generally not well documented either, for Manitoba conditions. As a result of these challenges, legislation and regulation regarding P management should be introduced cautiously to ensure environmental protection without undue hardship to the agricultural industry.

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Foreword

The original proposal was prepared under the leadership of Dr. Régis Simard, Head of the Department of Soil Science in the fall of 2001. However, when Dr. Simard became ill and subsequently died from an inoperable brain tumour, the rest of the project’s co-applicants (Dr. Kathy Buckley, Dr. Wole Akinremi, and Dr. Don Flaten) assumed full responsibility for the project in the summer of 2002. Dr. Ken Snelgrove and Dr. Ed Tyrchniewicz joined the project team in fall 2002. Research associates (Ian Halket, Grant Penn, and Brian Wiebe) were hired in the fall of 2002 to work under the leadership of the project’s research leaders.

With limited resources for the project, the main emphasis in the literature search was on information that would be appropriate for Manitoba’s soils, waters, landscapes, climate and agricultural practices. Furthermore, most of the literature that was reviewed was from refereed research journals, although some literature from government reports and websites was also reviewed.

If resources, especially time, had not been limited, the review might have been expanded to include more emphasis on such topics as: the selection of water quality standards that are appropriate for Prairie Canadian water bodies; identification of critical source areas for surface water and nutrients; effects of surface and tile drainage on water and nutrient flow to rivers and streams; the role of sheet or overland flow for transporting nutrients in a dispersed drainage system that includes a large portion of land that is affected by tillage and wind erosion, and fertilizer and manure application; the effect of soil texture and landscape on the transport of colloidal soil particles that can travel long distances and which would be regarded as “dissolved P” according to conventional water quality analyses; the role of vegetation as a source, rather than sink for P, especially in a snowmelt-dominated runoff system; the effect of livestock overwintering sites and livestock access to surface water on the risk of P transfer; opportunities to increase the N:P ratio of manure by conserving N; and the opportunity to develop innovative methods of intensive livestock production that result in a lower water content and, hence, lower transportation cost for manure nutrients.

Funding for this review was received from the Manitoba Livestock Manure Management Initiative (MLMMI), the Manitoba Rural Adaptation Council (MRAC), and the Manitoba Conservation Sustainable Development Initiatives Fund (SDIF). In addition, this review was supported by the Agriculture and Agri-Food Research Centre in Brandon and the Departments of Soil Science, Ag Business and Civil Engineering at the University of Manitoba.

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Introduction

Concerns for Phosphorus in the Environment The main concern for excess P accumulation in soil is the risk of contamination of surface water bodies and the subsequent decline in water quality by the process of eutrophication (increasing growth of algae, surface scums, followed by the depleted oxygen concentrations, foul odours, sedimentation, fishkills and release of algal toxins). The P responsible for eutrophication originates from a variety of sources, not only from livestock and crop production activities, but also from natural ecosystems and direct discharge of human and industrial waste.

Eutrophication occurs at very low concentrations of P in water (Correll 1998). Such small amounts of P loss are not agronomically significant, but are very significant from a environmental perspective. In addition, the environmental impacts of these small losses of P occur many miles from the P source (e.g., in Lake Winnipeg) and reflect the cumulative impacts of all sources. As a result, where water quality threshold concentrations for P are established, they are usually set at very low concentrations. For example, in Manitoba Conservation’s latest water quality objectives, standards, and guidelines, the upper limit has been set at 0.05 mg P/L for streams and rivers, and 0.025 mg P/L for lakes (Manitoba Conservation 2002). These standards have been recently reinforced by the announcement of the Manitoba Government’s Water Strategy (Manitoba Conservation 2003). Part of the reason for developing such standards is due to concerns over rising concentrations of P in many Manitoba streams and signs of eutrophication in Manitoba lakes (Jones and Armstrong 2001).

The main sources of P contamination in surface water include point sources from urban and industrial activities and non-point sources from agricultural activity and natural processes such as erosion and wildlife. Small amounts of P may also be deposited from rainfall and snow, especially in areas where atmospheric pollution is a concern. Runoff of P from wildlife waste and leachate from vegetative material also increases the concentration of P in surface waters, especially in semiarid regions such as the Canadian Prairies, where the total volume of runoff is relatively small, too. As a result, the quality of many Prairie water bodies was poor long before the beginning of any significant agricultural activity in the region (e.g., the La Salle River, at the southern edge of Winnipeg was formerly called the “Riviere La Sale” or “Stinking River” by French and English settlers, respectively, in the 1800s (p. 105, Shilliday 1993). However, diatom studies have shown that the introduction of urban and agricultural activity into a region causes a substantial increase in accumulation of P in water (Dillon and Kirchner 1975). In Manitoba, as well, most of the streams where P concentrations are increasing are situated in Southern Manitoba, where urban and agricultural activities are most intense (Jones and Armstrong 2001).

Aquatic loading of P due to agricultural activity is a significant source of water quality impairment in Canada (Chambers et al. 2001) and the U.S. (Parry 1998). The main sources of P that contribute to these losses are synthetic fertilizers, crop residues, and livestock manure. For the latter source, areas that export P into surface water include pastures, manured hayland, manured cultivated fields, wintering areas, feedlots, manure storage systems, and livestock handling yards. As a result of the important role played by agriculture in increasing the concentration of P in water bodies, management of agricultural sources of P contamination has attracted considerable attention from the scientific community in North America and Europe

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(please refer to Sharpley and Menzel 1987, Daniel et al. 1998, Higgs et al. 2000, Sharpley et al. 1994, Sharpley and Rekolainen 1997, Sharpley et al. 2000, Sharpley and Tunney 2000, and Van der Molen et al. 1998 for some excellent reviews on this topic).

The challenge of addressing the issue of P transfer to water bodies is, however, complex. Phosphorus behaviour in soil, in water and at the interface between these two media is very complicated and so is the jurisdictional responsibility for managing these resources. As a result of this complexity, P management is a challenge that requires a multi-disciplinary and multi-agency approach. For example, knowledge of hydrological and soil science is needed to understand the problem; knowledge of these sciences plus animal science, agricultural engineering and agricultural policy is required to develop practical solutions. Similarly, in order to develop and implement a sensible, fair and effective P management strategy, Federal, Provincial, and municipal agencies, along with university scientists and agricultural producer groups must work cooperatively

Implementing Measures to Reduce the Risk of P Contamination of Water Recent increases in livestock production have coincided with increased concern for water quality, especially in Lake Winnipeg and Lake Manitoba. Therefore, pressure on the Government of Manitoba to regulate livestock manure management according the P content in the manure, in addition to the manure’s N content which is the current basis for regulation, is increasing.

In response to such pressure, governments have three main types of policy instruments to choose from: regulatory (controls on production practices or environmental performance), economic (taxes and subsidies) and extension (voluntary, participatory activities). Regardless of which instrument is selected, the basis for the instrument must be scientifically sound, so that the environment will be protected without imposing unnecessary hardship on agricultural producers.

The development of a sound scientific base for reducing the potential for P contamination of surface waters by agricultural sources requires detailed information about many factors. Such factors include the terrain attributes involved in the risk of overland or subsurface transport of P, the amount of P available from all sources and their modes of application on agricultural land. However, most of the information on P behaviour and transfer has been developed for areas in which soils are dominantly acidic and where transport is dominated by overland processes after rainfall (erosion and surface runoff). This information may not be pertinent to Manitoba because it is dominated by calcareous soils, relatively flat agricultural land base, and runoff processes that are mostly associated with snowmelt.

Therefore, before Manitoba can develop sound regulatory or codes of practice or evaluate the potential impact of those practices on Manitoba’s agricultural industry, the “state of the art” scientific and regulatory information that is available should be summarized and gaps in knowledge identified. In addition, the regulations and codes of practice have been proposed and/or adopted to limit the risk of transfer of agricultural soil P to water bodies in other jurisdictions should be assessed for their adaptability to Manitoba conditions.

Factors Affecting the Risk of Phosphorus Transfer from Agricultural Land: An Overview The factors affecting the risk of P transfer from agricultural land are similar to the typical source-receptor-pathway model that applies to most types of environmental contamination. In the case

of nutrients applied onto agricultural land, the source factor can be subdivided into several subfactors including form of nutrient and rate, method of application; rate of nutrient removal by crops and livestock; and the soil's capacity to retain nutrients. Pathway factors relate to the transport of nutrients in particulate and dissolved forms. And, of course, the receptor factor relates to surface or groundwater, where a defined level of water quality is desired. All of these factors interact in the general model described in Figure 1.

With and bharvesustaiP are water

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Figure 1. Factors affecting the risk of nutrient transfer to water; for long-term agronomicand environmental sustainability, nutrient application rates are equal to removal rates.
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respect to P, much of the agricultural land in Manitoba is naturally low in phosphorus (P) enefits from the application of P in livestock manure (Figure 2). If P that is removed as sted grain or forage is not replaced, the fertility of the soil will decline and the nability of agricultural production will fail. However, if excess rates of manure or fertilizer applied, significant amounts of P can move off the land or through the soil and into surface bodies.

f the reasons why manure P is often applied in excess of the crop’s P requirements is that :P ratio of manure is typically 3:1 or less, whereas the N:P ratio of most crops is 4:1 or Therefore, when manure is applied to meet the crop’s N requirements (e.g., according to t regulations in Manitoba), the long-term consequence is a surplus of P in the soil.

al management strategies are being employed to address the challenge of lowering or ating surplus manure P application to soil. One strategy is to improve the efficiency by feed P is utilized and reduce the excretion of P by the animal (e.g., improved genetic ility of livestock to use feed P, feed additives, reducing feed waste and P supplementation, atching the diet more precisely to age and sex of animal). Another strategy is to

ribute, recover or immobilize manure P, after the manure is excreted (e.g., mechanical al of manure solids, biological and chemical immobilization of P).

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Figure 2. The majority of Manitoba soils are rated medium or lower for their concentration of extractable P according to agronomic soil tests and require P application for optimum production. Numbers for each area represent the percentage of fields testing medium or lower in extractable P (Fixen 2002).

A short-term surplus of P, however, does not always result in a significant increase in P contamination of surface water. Each soil type has an inherent capacity to retain P through processes such as chemical precipitation into solid forms, chemical adsorption onto soil surfaces or biological immobilization into organic materials. This retention capacity has a large influence on the ease with which the P in the soil is released for crop production or for water contamination. In a case where P is applied in excess of crop removal, the risk of P transfer from soil that has a high retention capacity is probably low, especially if that retention capacity is far from being saturated (Figure 3a). However, the same nutrient surplus for a soil with a low retention capacity may create a high risk of P transfer, especially if the saturation of that retention capacity is already high (Figure 3b).

Figure 3. The risk of P transfer is low if the soil has a high retention capacity for P and a low degree of saturation; however, the risk of P transfer is high if the soil has a small retention capacity for P and a high degree of saturation.


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The risk of P transfer cannot, however, be predicted by the soil’s remaining capacity to retain P and the balance between P application and removal, alone. The ease with which P can be transported to surface water must also be considered. The main pathways for P movement to surface waters (Figure 1-4) are:

• release of soil-bound precipitated or adsorbed P from eroded soil entering water bodies • soluble P dissolved in surface runoff • flow of P from groundwater contaminated by P leaching

Figure 4. Transport factors must also be considered when evaluating the risk of P transfer from land to surface water (Sharpley et al. 1999).

Environmental problems with P loss do not occur unless source factors (e.g., P balance and soil’s remaining capacity to retain P) and transport factors occur together simultaneously. For example, a soil with a high concentration of P and little opportunity for soil and water movement to a water body is not considered a major environmental hazard. Similarly, the environmental threat of a soil with a low concentration of P and high potential for movement to a water body is also low. However, a P enriched soil in a landscape and climate where movement to water is likely, will pose a significant threat to water quality. As a result of the interaction between these source and transport factors, a high proportion of the P loss from agricultural land is often confined to a relatively small area (Sharpley et al. 1999).

References Chambers, P. A., Guy, M., and Roberts, E. 2001. Phosphorus losses from agriculture: effects on Canadian ecosystems. Pages 1-11 in 38th Annual Alberta Soil Science Workshop Proceedings, Lethbridge, Alberta, February 20-22, 2001.

Correll, D. L. 1998. The role of phosphorus in the eutrophication of receiving waters: a review. J. Environ. Qual. 27:261-266.

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Daniel, T.C., Sharpley, A.N., and Lemunyon, J.L. 1998. Agricultural phosphorus and eutrophication: a symposium overview. J. Environ. Qual. 27:251-257.

Dillon, P. J. and Kirchner, W. B. 1975. The effects of geology and land use on the export of phosphorus from watersheds. Water Research 9:135-148.

Fixen, P.E. 2002. Soil test levels in North America. Better Crops 86(1):12-15.

Higgs, B., Johnston, A.E., Salter, J.L., and Dawson, C.J. 2000. Some aspects of achieving sustainable phosphorus use in agriculture. J. Environ. Qual. 29:80-87.

Jones, G. and Armstrong, N. 2001. Long term trends in total nitrogen and total phosphorus concentrations in Manitoba streams. Manitoba Conservation Report No. 2001-07.

Manitoba Conservation. 2002. Manitoba Water Quality Standards, Objectives, and Guidelines. Final Draft: November 22, 2002. Manitoba Conservation Report 2002-11. Manitoba Conservation, Winnipeg, MB.

Manitoba Conservation. 2003. Manitoba’s Water Strategy. Manitoba Conservation website: http://www.gov.mb.ca/conservation/waterstrategy/pdf/water-strategy.pdf

Parry, R. 1998. Agricultural phosphorus and water quality: a U.S. environmental protection agency perspective. J. Environ. Qual. 27:258-261.

Shilliday, G. 2003. Manitoba 125: A history. Great Plains Publications, Winnipeg, MB

Sharpley, A.N. and Menzel. 1987. The impact of soil and fertilizer phosphorus on the environment. Adv. Agron. 41:297-324.

Sharpley, A.N. and Rekolainen, S. 1997. Phosphorus in agriculture and its environmental limitations. Pages 1-53 in H. Tunney, O.T. Carton, P.C. Brookes and A.E. Johnston, eds. Phosphorus loss from soil to water. CAB International, Wallingford, UK.

Sharpley, A.N. and Tunney, H. 2000. Phosphorus research strategies to meet agricultural and environmental challenges of the 21st century. J. Environ. Qual. 29:176-181.

Sharpley, A.N., Chapra, S.C., Wedepohl, R., Sims, J.T., Daniel, T.C., and Reddy, K.R. 1994. Managing agricultural phosphorus for protection of surface waters: Issues and options. J. Environ. Qual. 23:437-451.

Sharpley, A.N., Daniel, T., Sims, T., Lemunyon, J., Stevens, R., and Parry, R. 1999. Agricultural phosphorus and eutrophication. United States Department of Agriculture, Agricultural Research Service. Publication ARS-149, 42 pp. Website address: http://www.ars.usda.gov/is/np/Phos&Eutro/phos&eutrointro.htm

Sharpley, A.N., Foy, B., and Withers, P. 2000. Practical and innovative measures for the control of agricultural phosphorus losses to water: an overview. J. Environ. Qual. 29:1-9.

Van der Molen, D.T., Breeuwsma, A., and Boers, P.C.M. 1998. Agricultural nutrient losses to surface water in the Netherlands: impact, strategies and perspectives. J. Environ. Qual. 27:4-11.

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1 Relative Magnitude of the Phosphorus Discharged Mr. Ian Halket and Dr. Ken Snelgrove

Department of Civil Engineering, University of Manitoba and

Dr. Brian Wiebe Department of Soil Science, University of Manitoba

1.1 Chapter Summary Eutrophication is a significant problem for Manitoba’s waterways. Phosphorus (P), the limiting nutrient, is at the root cause. Since all the rivers in southern Manitoba drain to Lake Winnipeg, the state of P in the lake is a good barometer for the P problem in the province as a whole. Studies such as those conducted by Stainton et al. (2003) and Pipp (2002) show the lake is under stress. In attempting to deal with this problem, any solution must consider the sources of P to the lake. Of the total P carried by rivers to Lake Winnipeg, the Red River is by far the largest contributor, with approximately 70% of the annual load (Bourne et al. 2002). Unfortunately, there is little, if any, literature or data available which quantifies the apportionment of P sources. It is also important to note that the Red River system is not entirely within Manitoba and only 40% of the river’s P load originates from within the province; the bulk of the river’s P load emanates from outside of the province, and is therefore outside the province’s direct control.

The literature is clear that urban point and non-point sources coupled with agricultural activities are major sources of P in Manitoba. However, how much P each source contributes is a question that remains largely unanswered. The majority of point source effluents such as municipal and industrial discharges can only be estimated because of lack of data in the Red River watershed. The same constraint also applies to estimates of agricultural sources while the role of urban non-point sources, especially along the Winnipeg sections of the river, lacks data entirely. Further scientific investigation of these major sources is required in order to fully quantify their risk of P export. The role of in-stream processes in modifying the total P load in Manitoba is similarly not well understood, and more research is also required on this critical aspect of P delivery.

Of major concern to Manitobans is the timing of the P release from land to water. Research demonstrates that most of the P is released in the spring during the snowmelt period. The literature on rainfall runoff processes is rich, but lacking with respect to snowmelt runoff processes. Without data on this aspect, questions with regard to snowmelt processes remain. Do best management practices applied to rainfall-generated P runoff also apply to snowmelt conditions? The answers are not known.

Another interesting observation, unearthed in the Clear Lake watershed by Manitoba researchers, is the finding of significant concentrations of P in snowfall. Is this P in snowfall a local phenomenon, or is it province wide? Is P also present in rainfall? More research is needed in order for scientists to be able to answer these questions.

The majority of the water quality sampling for P is conducted on a periodic basis. In order to fully assess P runoff processes, snowmelt and rainfall flood data is required. Water quality sampling strategies should be revised to capture such key events. Event-based data would also provide an understanding and possible extension of the historical P concentration record collected for our waterways.

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The data show that P concentrations and loads within our rivers and lakes are higher than recommended levels for the maintenance of healthy aquatic ecosytems. More work should be conducted to determine the sustainable capacity of the aquatic system to P loads in a Manitoba context. However, the questions raised in this chapter require expedient scientific effort to answer, so that Manitobans can assess and explore strategies that are effective in addressing their concerns about eutrophication.

In the context of increasing concerns for P loading in Manitoba’s rivers and lakes, livestock production in Manitoba has been growing. As a result, livestock manure is becoming an increasingly important source of P in the Manitoba’s terrestrial and aquatic environment. For example, Statistics Canada estimates that manure nutrient production (i.e., animal units) by beef cattle and pigs in Manitoba has risen by 35 and 65 percent, respectively over the last ten years, with beef cattle responsible for the largest absolute increase in manure production during this period (+228,000 animal units). As a result of the dramatic growth in Manitoba’s livestock industry, two municipalities in Manitoba (La Broquerie and Hanover) now have the sixth and eleventh highest densities of livestock in Canada, respectively. However, across the province as a whole, manure P continues to account for only 19% of total P removed by the province’s crops. Furthermore, the amount of P applied as fertilizer in Manitoba is approximately six times that applied as manure.

1.2 Introduction Eutrophication, the excessive growth of aquatic plants in a water body, is a significant and growing problem for Manitoba’s waterways. The predominant cause of eutrophication is increased levels of phosphorus (P), which permits plant growth to proceed unchecked. Where is the P in Manitoba’s waters coming from? Chambers et al. (2001) suggested that “changes in human activity in the basin over the past 30 years have increased stress on the Lake Winnipeg ecosystem. It appears there has been a gradual increase in N [nitrogen] and P likely due to changes in agricultural practices, including expansion of the livestock and food processing sectors, and increases in human population.” Determining where the P in Manitoba’s waters is coming from is important in order to effectively deal with the problem of eutrophication.

Recent studies in Manitoba point to several sources of P in our waterways, the major sources being natural sources, municipal and industrial wastewaters, and agriculture. With a focus on southern Manitoba, specifically Lake Winnipeg and the Red River watershed, this chapter will review the literature pertaining to the importance of these P sources in Manitoba and identify where additional research is needed.

This chapter will also review the literature describing the transport of P from land to water. There is an abundance of scientific literature and models on the runoff processes affecting P movement from land to water. However, the bulk of these studies is devoted to the rainfall runoff processes, and on the prairies it is the spring snowmelt flow that dominates the rivers and streams, with 85% of the annual flow occurring in this period (Nicholaichuk 1967). The literature suggests that snowmelt runoff processes are substantially different from rainfall runoff process, and therefore the movement of P from land to water during snowmelt is very different from P movement during rainfall. Understanding the P transport processes during snowmelt is important to understanding P in Manitoba waters.

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1.3 Phosphorus in the Environment – Why P is Important P is the eleventh-most common element in the earth’s crust. P occurs primarily as phosphate in deposits of apatite (Ca5F(PO4)3), a mineral found in igneous, sedimentary and metamorphic rocks. Rocks containing apatite, when exposed to weathering, release the phosphate into the environment. The release of this initial inorganic species of P, usually orthophosphate, depends on the pH of the receiving environment.

There are four reasons for the scientific emphasis on P. First, as an essential element for all plant life, P (or bioavailable P) is generally thought to be the growth-limiting factor for algae and other aquatic plants in fresh water ecosystems. Many studies have shown the high correlation between increasing concentrations of P and algal growth (Ahlgren et al. 1988, Prairie et al. 1989).

Second, it takes only very small amounts of P for aquatic plant growth. For example, the cell stoichiometry of phytoplankton suggests that, on average, for every one part of P, cells contain 15 parts of nitrogen and 100 parts of carbon. Relative to the other major nutrients, therefore, very little P is required for plant growth. The critical concentration of P that accelerates growth of algae and other aquatic plants in lakes is very low. Daniel et al. (1998), suggest that thresholds above which accelerated plant growth in lakes occurs are 0.01 mg/L for dissolved P and 0.02 mg/L for total P. Levels of P released to waterways by human activity usually exceed this level.

That P is easier to control than the two other major nutrients, carbon and nitrogen, is the third reason for the emphasis. P originates from only one original source (earth materials), whereas carbon and nitrogen are found in many sources (earth, air and water).

Fourth, when there is an excess of P, as is purportedly occurring in many of Manitoba’s lakes and rivers, then algae that can fix nitrogen tend to dominate in the water body. These nitrogen fixing species are considered nuisance species because of their often noxious and toxic natures.

For these reasons, P becomes a pollutant when it is introduced to surface waters in abundant quantities. Thoman and Mueller (1987) note that excessive algal levels in fresh water are undesirable and can cause odour, fish kills, habitat destruction, and a general degradation of the aesthetic and recreational value of the water body. Identifying the sources and amounts of P released to Manitoba’s fresh waters is important, therefore, in order to manage P to maintain the health of aquatic systems.

1.4 Forms of P in Water There are many different forms of P in water, all of which are measured as total phosphorus (TP). However, only some forms of P are available to aquatic plants, the key consideration in the analysis of P’s contribution to eutrophication. TP, therefore, is not an accurate reflection of the P that is available to plants.

The only plant-available form of P is orthophosphate, of which there are two types in typical aquatic systems: HPO4

2-, which predominates in alkaline conditions; and H2PO4- in acidic

conditions (Busman et al. 2002). Both of these forms are also referred to as bioavailable P, or reactive P. The bioavailability of other forms of P in water depends on the rate by which they are converted to orthophosphate. Some forms may slowly release orthophosphates while still others are basically inert. Bioavailable P as a proportion of TP varies tremendously depending on the sources and types of P available.

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Analysis of P in water usually divides TP into two fractions: dissolved P (DP) and particulate P (PP). This division is most often accomplished operationally by filtration through a 0.45 micron filter. PP includes phosphates adsorbed to soil particles, P precipitates and living and dead phytoplankton. DP includes truly soluble P, the orthophosphates, polyphosphates and pyrophosphates, as well as colloidal P (soil particles less than 0.45 microns in size), and organic excretions. The fractions of DP and PP are also divided into organic P and inorganic P. Bioavailable P is found in all fractions, although the organic fractions are considered to be mostly inert. The inorganic fractions may, however, become long-term sources of bioavailable P as they slowly metabolize in river and lake sediments. It is the inorganic PP and DP that contain the largest proportions of bioavailable P, and in water quality sampling this total inorganic fraction is considered bioavailable and reported as total reactive P (TRP).

The organic and inorganic particulate and dissolved P fractions are continuously changing in water. Phytoplankton consume dissolved P and convert it to organic P. The phytoplankton are in turn consumed by zooplankton which excrete inorganic P and the cycle continues. Inorganic and organic particulate P (phytoplankton and zooplankton) settle-out of the water in quiescent environments such as lakes, backwaters and reservoirs. Here the particulate P is covered by sediment and, if burial is rapid enough, it may be removed from circulation. On the other hand, the particulate P may be reintroduced to the water column through biological and chemical processes acting in the benthos.

The impact of P in water is directly related to the type of P present and, more specifically, by how much P is convertible to bioavailable P. Bioassays, which determine the amount of bioavailable P directly by measuring algal uptake, are the most reliable of all test methods (Rekolainen et al. 1997). However, biossays take a long time to complete, up to three weeks, and can require complex laboratory set-ups. As a result, surrogate chemical tests have been developed to determine bioavailable P. These tests include the molybdate method, with or without acid digestions, and adsorption on iron-oxide strips. These tests provide an indirect measure for bioavailable P. No particular test is currently accepted as a world or country wide standard; Manitoba uses the molybdate method to test for bioavailable P.

A new scientific line of inquiry is the role of pyrophosphate. Most studies of the role of P in eutrophication have focused on orthophosphate. However, some biologically available chemical forms of P (pyrophosphate and other polyphosphates) which are present in fertilizer and animal feed and in urban wastewater escape detection by standard analysis techniques (Richardson and Sundareshwar 2002). As this new development points out, analytical techniques are continuously improving, in step with the knowledge of the physical chemical and biological impacts of the measured P fractions. Reviews of analytical techniques are found in Boström et al. (1988), Pettersson et al. (1988), Froelich (1988), Engle and Sarnelle (1990) and Haygarth and Sharpley (2000).

Scientific opinion on what constitutes eutrophic status for lakes in terms of TP levels ranges from 0.016 mg/L to 0.390 mg/L with a mean of 0.080mg P/L (Vollenweider and Kerekes 1980). The Manitoba Surface Water Quality Standards, Objectives and Guidelines (MSWQSOG) state that TP should not exceed 0.025 mg/L in any reservoir, lake or pond, or in a tributary at the point where it enters such bodies of water (Williamson 1988). A level 0.05 mg P/L is recommended for streams and rivers. These thresholds may be unattainable for many of the prairie region’s lakes and rivers, as natural background levels are already much higher. The review of the

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MSWQSOG currently underway, is considering setting site-specific criteria to recognize these regional differences in P concentrations in Manitoba’s waters.

1.5 P in Manitoba’s Waters Lake Winnipeg is the central component of the Nelson River system, a huge watershed draining all of Southern Manitoba and a large part of central North America (Figure 1.1). The lake is fed by three major rivers: the Red River, Winnipeg River and the Saskatchewan River. These watersheds span an area of central Canada that encompasses eco-regions ranging from boreal shield to prairie. The Winnipeg River contributes the largest inflow to Lake Winnipeg, about 40%; the Saskatchewan River 22%; while the Red River contributes about 8%.

In a recent study, Stainton et al. (2003) suggested that Lake Winnipeg is becoming increasingly eutrophic. The authors point to trends in algal community, zooplankton community, nutrient concentrations, and chlorophyll levels, compiled from recent studies of the lake, as supporting evidence for increased eutrophication. Sentinel species such as freshwater clams and the Lake Winnipeg snail are disappearing from the lake (Pipp 2002). “Possible triggers for this change are elevated loading of phosphorus from agricultural runoff and/or municipal effluent and/or increases in light penetration and phosphorus recycling brought about by hydroelectric impoundment” (Stainton et al. 2003).

Figure 1.1. Lake Winnipeg Watershed (Manitoba Conservation).

Three recently released reports, two by Manitoba Conservation and one from Canada’s Department of Fisheries and Oceans, have set the stage for discussion of the sources of P to Lake Winnipeg and its tributary waters. Jones and Armstrong (2001) analyzed total nitrogen (TN) and TP data, collected at 46 water quality sampling stations on 33 different rivers and creeks in Manitoba, to determine long term trends. The period of record for each station varies, the longest being from 1973 to 1999, and the shortest from 1990 to 1999. Using a statistical program

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developed by the United States Geological Survey (QWTrend), the authors suggested 15 streams had increasing concentrations of TP over their periods of record, 7 had decreasing concentrations, while 15 showed no long-term trend at all. Thirteen of the streams showing increases in TP concentrations over the long-term are in the Red River system.

Bourne et al. (2002) reported on the TP and TN loading to streams in Manitoba. The authors estimated the total nutrient loads at 41 water-quality monitoring stations in the Province for the period 1994 to 2001, and applied a P export model to estimate nutrient sources. The emphases of the report were the streams of the Red River watershed in Manitoba. McCullough (2001) also reported on the TP, TN and total organic carbon (TOC) loadings to Lake Winnipeg from its major tributaries. The report examined the period from 1968 to 1999, but calculated the mean annual TP loadings on shorter periods of record. For instance the mean annual TP loadings for the Red River are based on 1990-1999 record.

All of these reports provide a foundation for further analysis of P in the Red River basin and both the Bourne and McCullough reports are consistent in their conclusion that the largest contributor of TP to Lake Winnipeg is the Red River. Figure 1.2, prepared from data used in Bourne et al. (2002) and McCullough (2001), for the period 1994 to 2001, shows the estimated average annual TP input to Lake Winnipeg from its major tributaries. Despite some discrepancy between the reports in the calculated annual TP loads for the tributary rivers, due to the different periods of records used, both paint a picture of Lake Winnipeg as receiving a greater annual loading compared to what it is losing through the Nelson River outlets. Consequently, the lake must be storing a large amount of TP in its sediments, which raises the question of whether the unique biogeochemical dynamics of the Lake can sustain this amount. It is also interesting to consider how the Lake would respond if the P load was suddenly curtailed, given the large amount of P currently stored in the Lake’s bottom sediments. Would the Lake show immediate improvement or would the amount stored in the bottom sediments act as a long-term source of P? Further scientific investigation is required to answer these questions.

With just over 4900 tonnes of TP per year, the Red River system is by far the largest contributor of TP to Lake Winnipeg (Bourne et al. 2002, Figure 1.3). Since the period of record, 1994 to 2001, includes some of the wettest years on the Red River relative to other tributaries of the Lake, the percentage TP contribution (73%) from the Red River may be slightly overestimated. In fact, McCullough (2001), using an earlier period of record, 1990-99, reported the annual load at 3900 tonnes TP. McCullough noted that this period of record would produce conservative estimates because it includes some dry years at the beginning of the decade.

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Total Phosphorus Loading to Lake Winnipeg

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/yr)

Figure 1.3. Phosphorus loading to Lake Winnipeg (Williamson 2003).

Of the 4905 tonnes of TP supplied to Lake Winnipeg on an annual basis by the Red River 2876 tonnes, or 59% of the load, originates from outside the province (Bourne et al. 2002). Brigham et al. (1996) support this figure with an estimate of 60% of the TP load in the Red River originating from outside the province. Consequently, these reports have serious implications for management practices, concluding as they do that the major source of P is outside of the province’s direct control. The Bourne report goes on to estimate the proportion of TP from three other sources within Manitoba’s share of the Red River watershed, shown in Figure 1.4. Agriculture is estimated to contribute 17% while natural sources are responsible for 13% and direct wastewater discharges 11%. If McCullough’s annual P load estimate of 3900 tonnes were used the contribution from direct waste water discharges would rise to 14% while agriculture and natural sources would fall to 12% and 15%, respectively.

Figure 1.2. Average annual TP loads and median concentrations in Lake Winnipeg rivers. (Adapted from Jones and Armstrong 2001, Bourne et al. 2002 and McCullough 2001).

8

Figure 1.4. Phosphorus loading to Lake Winnipeg (Williamson 2003).

The Bourne report used two approaches to identify the sources of nutrients in the Red River basin. First, the authors suggested that the total measured stream nutrient P load (TMSNL) is the sum of the P supplied from in-stream processes plus watershed processes. Subtracting the in-stream processes P from the TMSNL would yield the amount of P resulting from watershed sources. The second part of the approach is to independently calculate the amount of P supplied by watershed processes, by applying land use export coefficients. The answers from both approaches are then compared.

The first approach suffers, however, from the lack of meaningful data detailing in-stream processes. The authors acknowledged this problem, but without more data they were unable to solve it. They managed to quantify municipal sewage discharges to the rivers, either through measurement or estimates based on population. But they were not able to assess the natural fluxes of P in river sediments, in groundwater, or from other in-stream processes.

The second part of the Bourne approach, estimating the watershed sources of P in the Red and Assiniboine Rivers, divided the watersheds into four general land-use categories: pasture, cropland, forest, and all other types. Coupling these categories with nutrient export coefficients for the land-use types found in the literature allows for a generalized quantification of the watershed sources of TP. This approach, originally developed for the Lake Mendota watershed in Wisconsin (Bennet et al. 1999), was used by Chambers and Dale (1997) in the Athabasca, Wapiti and Smoky Rivers of Alberta. Two of the export coefficients, pasture and forest, are based on work reported by Chambers and Dale (1997) for the Dakotas, Minnesota and Wisconsin. The export coefficient for cropland is the result of the work by Green and Turner (2002) on the South Tobacco Creek watershed. The “other” category, which includes all other land-use categories and is a diverse group from urban to wetland, uses an export coefficient value that describes atmospheric deposition of TP in Alberta.

The results of the Bourne analysis combined with their estimates of the municipal wastewater treatment plant discharges are summarized in Table 1.1.

Total Phosphorus Loading to the Red River

59 %

13 %17 %

11 %

0

1000

2000

3000

4000

5000

Red River United States Estimated NaturalBackground

Present-DayAgriculture

Manitoba DirectWastewaterDischarges

TP (t

/yr)

9

Table 1.1. Sources of TP in the Assiniboine and Red River Watersheds in Manitoba, compiled from data in Bourne et al. (2002)

Pasture Cropland Forest Other Total TMSNLZ WWTFy TMSNL-WWTF

Assiniboine River TP (tonnes/yr) 123 808 59 49 1039 209 75 134

%TP 11 73 5 4 7 Red River

TP (tonnes/yr) 81 1037 43 48 1209 1731 470 1261 %TP 5 62 3 3 28

ZTMSNL – Total Measured Stream Nutrient Load yWWTF - Wastewater Treatment Facility discharges

The estimated amount of P for the Red and Assiniboine watersheds were then compared to the amounts estimated for these watersheds from the first approach. The estimate of the load from watershed processes (the sum of pasture, cropland, forest and other, above) for the Red River watershed of 1209 tonnes TP compares closely to the estimate of 1261 tonnes (TMSNL minus Waste-Water Treatment Facilities(WWTF) discharges) from Bourne’s first approach. However, the figures for the Assiniboine River, 1039 tonnes of TP versus 134 tonnes (TMSNL minus WWTF) estimated by the first approach suggest that there are other processes not accounted for in the analysis.

The use of nutrient export coefficients to determine P loading from land-use assumes that all land within a particular land-use category contributes equally to runoff. However, this concept is inconsistent with present-day hydrological theory on runoff processes. Runoff is highly variable both within and between catchments. Current hydrological modelling practice recognizes that soil type and moisture content, slope, management practices and other factors besides land-use are also important in determining amount of runoff and therefore TP export to the stream. Additionally, not all areas within a catchment necessarily contribute runoff to a stream. Runoff is generated from various source areas within a catchment, and these areas respond with varying degrees to the intensity of the snowmelt or rainfall event.

A different perspective for determining the source of the TP loads in the Red River system, which adds to the insight gained from the work of Bourne et al., is to break the River into three distinct reaches (Figure 1.2): 1) the confluence of the Red and Assiniboine as they pass through Winnipeg and on to Selkirk, 2) the Red upstream of Winnipeg, and 3) the Assiniboine upstream of Winnipeg. The first section is heavily urbanized, while the latter two are primarily agricultural. This division allows the relative contributions of each section to be assessed, and the shortcomings of the data and our understanding of the river processes outlined. The loading data reported in the Bourne report is used for this analysis.

Bourne et al. (2002) reported that the average annual TP load on the Red River downstream of Winnipeg at Selkirk is 4905 tonnes. Upstream of Winnipeg at St. Norbert it is 3103 tonnes. The Assiniboine River at Headingly conveys another 637 tonnes of TP, for a total of 3740 tonnes entering Winnipeg. Therefore, as the Red and Assiniboine Rivers course through Winnipeg and on to Selkirk, they gain 1165 tonnes of TP, or 24% of the total annual load delivered to Lake Winnipeg. This highly developed and urbanized section of river appears to be, therefore, a significant contributor of P.

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A note of caution concerning the load data reported by Bourne et al. (2002) should be stated here. In 1997, water samples were not collected for the month of May during ‘the flood of the century’. Therefore, the load Bourne et al. (2002) calculated for May used the average of the concentrations measured in April and June of 1997 to estimate this missing sample. This results in an underestimation of the May loading for the 1997 flood. Because the average monthly flow for this May period is the highest flow on record it has an impact on the average annual P load calculated for this site. There are rigorous statistical methods available for estimating this flow. However, averaging the reported May concentrations for the historic record would result in a revision of the annual TP load to 3273 tonnes at St. Norbert, resulting in this section being responsible for 20% of the load to Lake Winnipeg. This example also helps to illustrate the sensitivity of the calculation of annual TP yields when based on a monthly sampling period. This revised annual loading (3273 tonnes) for St. Norbert is used in the following analysis.

The estimated annual TP contribution from the City of Winnipeg’s wastewater treatment plants plus storm and combined sewers is 390 tonnes (Bourne et al. 2002). This leaves 605 tonnes of P unaccounted for in this stretch of river. The Bourne report does not explore this discrepancy. Also noteworthy is the change in the median 2001 concentration of TP along this stretch of river (Figure 1.2). Upstream of Winnipeg the median TP concentration is 0.22 mg/L and 0.21mg/L for the Assiniboine and Red rivers, respectively. Downstream of Winnipeg the median concentration increases to 0.32 mg TP/L. This is an increase in the concentration of approximately 50%. Similarly, Kalkhoff et al. (2000) reported that TP concentrations in the Eastern Iowa Basins Study Unit are among the highest in the corn-belt. The area has a land-use mosaic very similar to the Red River watershed, with agriculture covering 90% of the land, and forest 4%. The authors concluded that the TP concentrations are greatest in the large river basins that contain the largest cities and towns, which seems to parallel the situation in the Red River Valley.

The Lake Champlain Basin is highly forested and, therefore, very different from the Red River basin in terms of land-use. Yet, Hegman et al. (1999) in a study of P sources to Lake Champlain, estimated that 20% of the annual P load originates from wastewater and industrial treatment plant discharges, while an additional 30% is from urban non-point sources with agricultural land contributing another 44% of the annual P load.

There are many possible explanations for the TP loading within this section of the Red River basin. In-stream processes, such as groundwater recharge and bank erosion, non-point sources from the densely developed area adjacent to this stretch of river, especially leaky septic systems, and combined sewer overflows are all potential contributors. The identification of the sources of P in this section of river needs further investigation.

Upstream of Winnipeg, the Red and Assiniboine Rivers contribute annual TP loads from within Manitoba of 566 and 298 tonnes, respectively (Bourne et al. 2002). These figures are calculated by subtracting the loads arriving at the province’s borders from the loads calculated just upstream of Winnipeg. These sections, therefore, contribute 15% and 6% of the TP load to Lake Winnipeg (Figure 1.5).

11

Red R.wat ershed in Manit oba upst ream of Winnipeg

15%

Assiniboine R. wat ershed in Manit oba upst ream of Winnipeg

6%

Load f rom Winnipeg and it s environs

20% Load or iginat ing f rom out side of Manit oba

59%

Figure 1.5. Distribution of the Average Annual P Load in the Red River.

TP generation along the Assiniboine, however, is much more than is portrayed by the annual load figures. This is due to the presence of the two reservoirs, Shellmouth and Portage, which reduce the TP in the system through biological uptake and particulate settling, and the Portage diversion channel which, in times of high flows, diverts water to Lake Manitoba. An additional problem is the role played by plant communities along the Assiniboine immediately upstream of the confluence with the Souris River. Figure 1.2 shows that the Souris delivers 307 additional tonnes of TP annually to the 355 tonnes recorded at Brandon. One would then expect 662 tonnes (or an approximately equal amount) to be recorded downstream at Treesbank. Instead, the average annual load at Treesbank is 380 tonnes. What happens to the 282 tonnes? Part of the answer may lie in the Brandon Riffles, a shallow stretch of cobbled riverbed, downstream of Brandon, which is colonized by periphyton and other floating and rooted aquatic plants during the summer. Cooley et al. (2001) reported that this community is able to fix a great amount of TP in its biomass, resulting in improved water quality downstream. This community disappears in the fall, only to recolonize the Riffles the next summer. The fate of the P tied up in the biomass is unknown.

It is misleading, then, to believe that the annual TP load calculated just upstream of Winnipeg on the Assiniboine River accurately reflects the Assiniboine watershed’s P load. Obviously, in-stream processes acting along the Assiniboine River play a major role in the TP budget for the watershed. The extent and dynamics of the role of in-stream processes need further investigation.

It is worth noting that the Jones and Armstrong, Bourne, and McCullough reports all rely on periodic P sampling of streams and rivers. Nutrient data is derived from samples collected on biweekly, monthly or quarterly periods, while the flow data is continuously monitored. Thus, nutrient loads are calculated by averaging the flow data over the sampling period. The average flow for the period is multiplied by the sampled concentration to find the TP load. These loads

12

are then summed to determine the annual load. Because the nutrient data is not continuous, short-term fluctuations in TP concentrations are not accounted for. Indeed, Rekolainen et al. (1991) pointed out that this method of calculating TP loads may underestimate the loading by up to 40%. The implications for Lake Winnipeg are potentially enormous, considering the large amount of TP that is already estimated to be entering the lake on an annual basis. Rekolainen et al. (1991) suggested that a more accurate picture of TP loads can be obtained by augmenting the periodic TP sampling currently used with increased sampling during snowmelt flood and rainfall stormflow events.

In addition, the Jones and Armstrong, Bourne, and McCullough reports depend on very short, recent periods of record. These records may not be long enough to be indicative of long-term trends within the Lake Winnipeg basin.

Despite these cautions, these reports provide a snapshot of the present day sources of TP to Lake Winnipeg and a foundation for further analysis of P in the Red River basin. The Bourne and McCullough reports are consistent in their conclusion that the largest contributor of TP to Lake Winnipeg is the Red River (73%) and the Jones and Armstrong report (2001) concludes that the TP concentrations within the Red River Basin are increasing. Fifty-nine percent of this TP load on the Red River originates from outside the province. A large part of the P load, 20%, is derived from within Manitoba and originates from the highly urbanized stretch of river between St. Norbert and Selkirk, while the mainly agricultural sections of the Assiniboine and Red upstream of Winnipeg supply the remaining 21% (Figure 1.5).

1.6 Sources of P in Manitoba This section will describe the sources of P in Manitoba using a top-down approach. First, atmospheric sources of P will be discussed. Second, land sources of P are outlined, including natural, urban and agricultural sources. Last, the in-stream processes that effect the transport and storage of P are discussed.

1.6.1 Atmospheric P Sources Schindler et al. (1976) reported that atmospheric deposition is a significant source of P for lakes. There are three forms of P deposition from the atmosphere: dry, wet and phosphine gas deposition. Dry deposition is the P bound to dust and other small particulate matter that settles out of the atmosphere. Wet deposition is the P that is dissolved in precipitation, primarily rain or snow. Phosphine is a gas generated primarily in the natural environment by wetlands.

In studying combined dry and wet deposition in the Experimental Lakes Area of northwestern Ontario, Schindler (1977) estimated an atmospheric P deposition rate of 0.24–0.53 kg/ha/yr. Chambers and Dale (1997), in their study of the Athabasca, Wapiti and Smokey rivers in Alberta, used an atmospheric deposition rate of 0.20 kg/ha/yr based on work by Mitchell (1985) and Trew et al. (1987) on two Alberta lakes, Lake Wabamun and Baptiste Lake. Trew et al.(1987) reported a rate of 0.32 kg/ha/yr.

Looking at dry deposition alone, Bennett (1985) provided an estimate of wind blown phosphate concentrations in the Brandon, Manitoba area ranging from 0.01 - 0.04 µg/m3 with a geometric mean value 0.2 µg/m3. He suggested an atmospheric loading rate of 0.82 kg/ha/yr. Beck (1985) on the other hand, reported a slightly more conservative rate of 0.41 kg/ha/yr in his study of twelve lakes in the Riding Mountain Park area.

13

With regard to wet deposition, a recent discovery of the presence of high concentrations of P in snowfall around the Clear Lake area in Manitoba introduced the prospect that high P concentrations in snow and rain may be a province-wide phenomena. Thompson et al. (2000) and Belke et al. (2001) reported on P in the winter snowpack. Both of these reports measure P in snowfalls in the Clear Lake region, the Thompson study during the winter of 1999-2000 and the Belke study in the 2000-2001 winter. The first study used three sampling sites, collecting samples after each snowfall. The mean TP concentration in the snowfall samples for the year was 1.28 mg/L. This study was followed up the next winter by Belke et al., who reported that the mean phosphate ion concentration in the snowfall samples was 0.191 mg/L, ± 0.130 mg/L. This concentration was significantly less than the mean phosphate ion concentration in the 1999-2000 snowfalls of 1.280 mg/L, but still a very high level when compared to MSWQSOG level of 0.025 mg/L for lakes. The authors contend that a possible reason for the lower concentrations in the 2000-2001 winter, especially in the first snowfall, may be due to an earlier flushing of the atmosphere by autumn rains.

Unfortunately, the information available on this subject is very limited. Indeed, an examination of EPA air quality monitoring data showed that P levels in the air are not commonly monitored.

An interesting aside to the above reports is the hypothesis of Thompson et al. (2000) that the unusually high phosphate concentrations in the 1999-2000 snowfall can be attributed to fine particulates. The researchers employed a two-step filtration procedure in their analysis of the snowfall samples. They first filtered the samples through 0.45 µm filter and then through a 0.25 µm filter, finding that most of the P was retained between the filter sizes. This suggested that the phosphate found in the snowfalls was being deposited in association with colloidal particles. The authors also note that the first snowfall in the autumn had the highest concentration of P, followed by declining P concentration in subsequent snowfalls, until the spring, when P concentration increased. The spring snowfalls occurred after an initial melt period. The authors postulate that colloidal particles aerosolized from the prairie become the preferential nuclei of condensation for snowfall. However as the snow blankets the prairie, entrainment of the colloidal particles is reduced until ground is exposed in the spring. These observations and hypothesis need to be further investigated on the Manitoba prairies because of the implications they may have on P management.

A new development on the P horizon is the role of phosphine gas in the cycling of P through wetlands. Only a few articles are available as yet on this topic, but they are worth noting because of the large area of wetlands in the Manitoba prairies. Ji-Ang et al. (1998), in a study of phosphine gas in Beijing, China, point to the fluctuating diurnal nature of phosphine concentrations in air. The highest levels occur in the early morning and decline through the daylight hours because of interaction with ultra-violet radiation causing oxidation, which converts phosphine to soluble P forms within a matter of hours. The authors report the chemistry of the interaction is not fully understood. They report levels of 41 and 135 ng/m3 near paddy fields and marshlands. In a parallel study, Eismann et al. (1997) noted a similar diurnal trend for the environs around Leipzig, Germany, although levels were not as high as in Beijing. Devai et al. (1988) measured phosphine emissions from a constructed wetland in Hungary and estimated that 17 kg/ha/yr of phosphorus was being lost by this route. Devai & Delaune (1995) later report the first quantification of phosphine emission rates from waterlogged soils, between 0.036 - 0.27 kg/ha/yr for brackish marshes and 0.08 - 0.57 kg /ha/yr for salt marshes in Louisiana. To date,

14

wetland and P balance studies have not incorporated phosphine emissions from waterlogged soils and wetlands.

1.6.2 Terrestrial Sources Urban point and non-point sources are significant terrestrial sources of P to Manitoba's waters, as are natural and agricultural P runoff sources. The mainly agricultural portion of the watershed upstream of Winnipeg contributes approximately half of the Manitoba portion's annual P load to Lake Winnipeg. The other half of this load is from the City of Winnipeg and the portion of watershed downstream of Winnipeg. Urban point and non-point sources contribute significant amounts of TP to this latter section of the river.

1.6.2.1 Urban P sources There are two sources of P in the urban environment: point and non-point sources. Point sources are the traditional end-of-pipe sources, municipal and industrial waste water treatment plant discharges, which are usually the most easily identified and regulated. Non-point sources of P, which are also a major contributor of TP from urban areas, include runoff from lawns, parking lots, construction sites, septic systems, and developed areas that lack sewers.

Bourne et al. (2002) estimated the annual TP discharges from municipal and industrial wastewater treatment facilities as 553 tonnes for the Red River system. This estimate is a mixture of recorded data and estimates of effluent P based on population density of the municipal area. The records from three waste water facilities, Winnipeg, Brandon and Portage la Prairie, and three industrial facilities, are used in the calculation of annual P load. These data are then added to estimates of P discharges, based on population density, for only those municipalities that directly discharge to surface water. Other industrial discharges are not accounted for because of a lack of data.

There are a few problems with this estimate. It ignores the P load from the many industries within the basin that discharge directly to municipal wastewater treatment facilities. These industrial discharges probably surpass any estimate of effluent loading based on the population density of the town in question. It also ignores the contribution from wastewater treatment facilities that do not discharge directly to surface waters. Federally regulated facilities were not included, because as the authors state “data were insufficient to provide a reasonable estimation of loading from these sources” (Bourne et al. 2002). And the calculation itself, based on population density, multiplied by an influent TP loading factor of 3.38 g/capita/day, multiplied by a P removal efficiency factor of 0.59 or 0.65, depending on level of treatment, seems to be too unreliable for serious conclusions. It also relies on just one source for the values of the influent and efficiency factor (Chambers et al. 2001) which, considering the many and diverse systems estimated, probably deserves further verification.

The TP load calculated by the City of Winnipeg authorities is based on ‘dry weather’ flows. However, around 70% of the city is serviced by a combined sewer system. During periods of heavy rain and snowmelt (wet weather flows), the city’s three treatment facilities do not have the capacity to handle the increased flows and, thus evacuate the raw sewage and storm water overflows to the Red and Assiniboine Rivers. This additional TP load in the overflows during a calendar year does not seem to be accounted for.

15

Another distinct characteristic of the effluent from water treatment facilities is that almost all of the TP is soluble P, and therefore its effect on the environment will be much more acute than other sources where the DP fraction of TP is not as high. Currently, waste water treatment facilities within the Manitoba portion of the Red River watershed lack nutrient removal systems. However, many wastewater treatment plants in jurisdictions upstream of Manitoba have installed nutrient removal facilities.

Wastewater treatment facilities also create a sludge by-product, usually referred to as biosolids. This material, a residue from primary and secondary treatment processes, is dewatered to various extents and then land applied, much like manure. The amount of biosolids generated by a wastewater treatment facility depends on level and type of treatment processes. The disposal of biosolids continues to be a major concern for wastewater treatment facilities across the continent. Biosolids contain significant amounts of TP, but it usually is less labile than manure P. Therefore, biosolid P is not as prone as manure P to dissolving into surface water, but is not as effective as a fertilizer supplement. However, biosolids also contain other contaminants not found in manure like heavy metals, which make municipal wastewater biosolids less palatable as a supplement to fertilizer.

In an Issues in Ecology report on the sources of P in the United States, Carpenter et al. (1999) stated that P from non-point urban sources is a significant source of pollution to waterways in the U.S. Urban (non-point) runoff is the third leading cause of lake deterioration in the U.S., affecting about 28% of the lake area that does not meet water quality standards. As an example of municipal loading, Figure 1.6 shows the estimated annual TP yields from urban lands in the Milwaukee region.

Figure 1.6. Annual P loads to surface water from non-point urban areas. Source: Wisconsin Department of Natural Resources (1991).

ANNUAL PHO SPHO RUS LO AD

0 0.5 1 1.5 2

Freeways

Industrial

Commercial

Shopping Centers

Multi-Family

Small Lot Resid.

Large Lot Resid.

Parks

Pounds per Acre

16

Similarly, Winter and Duthie (2000) examined the nutrient budget for Lake Simcoe in Ontario from 1990 to 1998. They found that the mean annual exports of TP were highest from vegetable polders (1.09 kg/ha/yr), followed by the sub-catchment with the greatest proportion of urban land (0.65 kg/ha/yr). Mixed agricultural sub-catchments ranged from 0.11 to 0.27 kg/ha/yr of TP, with the lowest export being from catchments with a large proportion of forest and scrubland, ranging from 0.06 to 0.07 kg TP/ha/yr.

Waschbusch et al. (1999) reported on DP concentrations in runoff in two Madison area residential districts. Runoff concentrations of DP from streets varied from 0.12 to 0.39 mg/L, and from lawns from 0.22 to 0.53 mg/L, an order of magnitude higher than the water quality standard. The authors found lawns and streets to be the most significant sources of P in the test basins, contributing about 80% of the total annual loading. In a supporting study, Bannerman et al. (1993) reported that lawns in residential and industrial areas contributed 14% and 44% respectively of the P load to storm water runoff in Madison, Wisconsin.

In an examination of the Minneapolis-St. Paul region of Minnesota, Barten and Jahnke (1997) found that runoff from urban areas to lakes and streams is a major source of P. Creason and Runge (1992) estimated that approximately 3191 tons of P is applied to lawns in the region annually, the majority of this amount being applied to lawns that do not need P. Indeed, Barten (1994) found that over 67% of lawns in four suburban municipalities in the region had concentrations of soil test extractable P exceeding 50 lbs/acre.

From 1988 through 1994, Thorolfsson and Brant (1996) examined urban runoff in Norway during both summer and winter. The authors found that snowmelt runoff is much greater than typically planned for in drainage designs, resulting in greater winter flooding and combined sewer overflows than during the summer. The authors developed an urban storm-runoff model that accounted for snowmelt as well as rainfall, but they state that due to a notable lack of information regarding urban storm runoff during the winter season they were unable to validate their model. Likewise, in a review of water quality impacts of winter operations on urban drainage in Toronto, Oberts et al. (2000) recognize that data on the water quality impacts of urban meltwater is scarce. In this report, the authors conclude that further research on adaptation of conventional stormwater management techniques to cold climates is needed.

1.6.2.2 Natural P sources Natural land sources of P within the Red River and Assiniboine River watersheds are forest areas which include riparian zones immediate to the stream channels. In a study of 928 watersheds across the U.S. where non-point sources of P were responsible for P generation, Ormernik (1977) showed that as forest cover decreased and other types of land-use increased, TP generation increased. Figure 1.7 shows the P export from the different types of non-point sources investigated by the author. According to the author, P export from forested areas (<0.1 kg/ha/yr) is low relative to developed lands.

17

Figure 1.7. Non-point sources of P loss in the U.S. (Ormernik 1977).

Hegman et al. (1999) identified all sources of P export in the Lake Champlain basin of Quebec, Vermont and New York. Sixty-five percent of the basin is forested, but the authors estimate that forest contributes only 7% of the total annual P load. Bourne et al. (2002) estimated that forest accounts for only 17% of the land-use in the Red River basin, a much smaller proportion than in the Lake Champlain basin, and therefore the forest could be expected to contribute a small portion of the annual load.

1.6.2.3 Agricultural P sources The loss of P from agricultural land is dependent on several factors. These factors include the balance of P inputs into and output from agricultural systems; the forms, quantity, and availability of P in the soil; and the runoff characteristics of the catchment area (Sharpley and Rekolainen 1997). The balance of P inputs and outputs from a field or a region are determined by calculating the difference between P additions (fertilizer, manure, and biosolids application, and atmospheric deposition) and P removals (grain, straw and livestock export and P losses with runoff, leaching, and wind erosion). Soil P and runoff characteristics are discussed in chapter 3.

Fertilizer use has increased steadily in Manitoba since 1965 and although nitrogen (N) has shown the largest increase, phosphorus use has also more than doubled (Figure 1.8). Removal of P in harvested crops was fairly well matched to additions of fertilizer P until the early 1990’s after which additions have exceeded removals (Johnston and Roberts 2001). Despite the increase in P fertilization, 73% of Manitoba soils sampled tested medium or lower in plant available P in 2001 (Fixen 2002). Johnston and Roberts (2001) developed a more thorough P budget for Manitoba which included manure P as well as fertilizer P, based on the 12 agricultural regions of Manitoba as delineated by Statistics Canada for census purposes (Figure 1.9).

0

50,000

100,000

150,000

200,000

250,000

300,000

350,000

400,000

1965 1970 1975 1980 1985 1990 1995 2000

annu

al fe

rtili

zer

purc

hase

s (to

nnes

)

N

P2O5

K2O

Figure 1.8. Annual fertilizer nutrient sales in Manitoba from 1965 to 2000 (adapted from Korol and Rattray 1997; Korol and Rattray 2001).

18

Due to the difficulties associated with obtaining local data, Johnston and Roberts’ (2001) estimated P budgets for 2000 (Table 1.2) were determined as follows:

• Crop and livestock data were obtained from the 2000 issue of the Manitoba Agriculture Yearbook (Manitoba Agriculture and Food 2001).

• Production data for grain corn, all wheat, barley, oats, rye, canola, flax, potatoes, alfalfa, and other tame hay were used (represent in excess of 95% of total crop production in the province). The distributions of alfalfa, grain corn, and potato were not given in the Manitoba Agriculture Yearbook and were estimated from the 1996 Census of Agriculture (Statistics Canada 1997) and in consultation with Manitoba Agriculture and Food (Bill Moon, personal communications).

• The values for nutrient removal per unit of yield were obtained from the Potash and Phosphate Institute (PPI) website (http://www.ppi-ppc.org/ppiweb/canadaw.nsf).

• The provincial value for P fertilizer consumption for 2000 was obtained from Korol and Rattray (2001). The proportion of the provincial P total used in each agricultural region was assumed to be the same as the proportion of total lime and fertilizer purchased in each agricultural region as reported in the 1996 Census of Agriculture (Statistics Canada 1997).

• Manure nutrients were determined using livestock numbers from the 2000 Manitoba Agriculture Yearbook, and recoverable manure nutrients calculated using the USDA-NRCS method (Kellogg et al. 2000).

The estimated budgets for 2001 were prepared using the method of Johnston and Roberts (2001) with the following modifications:

• Crop and livestock data were obtained from the 2001 issue of the Manitoba Agriculture Yearbook (Manitoba Agriculture and Food 2002).

19

• Production data for peas, dry beans, and sunflowers was added. • The distribution of the production of alfalfa, grain corn, potato, peas, dry beans, and

sunflowers were not given in the Manitoba Agriculture Yearbook. They were estimated based on the 2001 provincial production (Manitoba Agriculture and Food 2002) and the planted area in each Agricultural Region reported in the 2001 Census of Agriculture (Statistics Canada 2002). Yields were assumed to be uniform across Agricultural Regions.

• The values for nutrient removal per unit yield were obtained from the Canadian Fertilizer Institute (CFI) Tables for western Canada on the PPI website (http://www.ppi-ppic.org/ppiweb/canadaw.nsf). Dry beans were not included in this table and the values from the CFI tables for eastern Canada were used (http://www.cfi.ca/uploaddocuments/d160%2BNU%5FE%5F01%2Epdf).

• Manure nutrients were determined using livestock numbers from the 2001 Manitoba Agriculture Yearbook (Manitoba Agriculture and Food 2002), and recoverable manure nutrients calculated using the USDA-NRCS method (Kellogg et al. 2000). The same coefficients were used for 2001 as were used for the 2000 data.

In 2000, fertilizer P applications met or exceeded crop removal in eight out of twelve of the regions (Table 1.2). The following year fertilizer P application declined by an average of 1.1 kg/ha and although crop removal of P (i.e. yields) were also lower the net effect was that P fertilizer applications met or exceeded crop removal in only five out of the twelve agricultural regions. Including recoverable manure P in the estimate increases the P balance and results in substantial annual P excesses in agricultural regions 7, 8, 9, 10, and 11 (range from 3.8 to 11.4 kg P/ha in 2000 and 2.3 to 11.3 kg P/ha in 2001). These five regions are centred around Winnipeg and drain into Lake Winnipeg largely via the Red River (Figure 1.9). These regions with the large annual surplus P levels are regions with large livestock populations and/or potato production (Johnston and Roberts 2001).

20

12

6

7

8

5

2

1109

4

11

Flin Flon

Brandon

LEGEND LÉGENDE

Régions agricolesde recensement

CensusAgricultural Region

Centre urbainUrban Centre

SaskatchewanOntario

3

WinnipegWinnipeg

Figure 1.9. Manitoba 2001 Census Agricultural Regions (Adapted from Statistics Canada 2002).

21

Table 1.2. Removal of P by crops, addition of P as mineral fertilizer and manure, and P balance in Manitoba by agriculture census regions (adapted and updated from Johnston and Roberts 2001)

Census Region

Field area 000 ha

Crop Removal of P

Fertilizer P Applied

Recoverable Manure P z P Balance

------------------------------- kg P/ha(lb P/ac) ------------------------------- 2000

1 599.8 8.9 (7.9) 7.7 (6.9) 0.6 (0.5) -0.6 (-0.5) 2 606.9 9.0 (8.0) 10.0 (8.9) 1.0 (0.9) 1.9 (1.7) 3 579.5 8.5 (7.6) 9.9 (8.8) 0.7 (0.6) 2.0 (1.8) 4 263.4 7.1 (6.3) 7.3 (6.5) 0.5 (0.5) 0.7 (0.7) 5 243.3 12.1 (10.8) 10.0 (8.9) 0.5 (0.4) -1.7 (-1.5) 6 453.4 8.6 (7.7) 8.3 (7.4) 0.6 (0.6) 0.4 (0.3) 7 736.1 9.1 (8.1) 13.6 (12.1) 1.6 (1.4) 6.0 (5.4) 8 821.5 10.6 (9.4) 12.8 (11.4) 1.6 (1.4) 3.8 (3.4) 9 380.0 7.4 (6.6) 13.2 (11.7) 5.6 (5.0) 11.4 (10.2)

10 94.9 6.4 (5.7) 7.1 (6.3) 3.0 (2.7) 3.8 (3.4) 11 264.3 8.2 (7.3) 11.3 (10.1) 2.2 (1.9) 5.2 (4.7) 12 335.9 8.3 (7.4) 6.4 (5.7) 2.0 (1.7) 0.0 (0.0)

Average 8.7 (7.7) 9.8 (8.7) 1.7 (1.5) 2.7 (2.4)

2001 1 577.2 9.4 (8.3) 7.0 (6.2) 0.9 (0.8) -1.5 (-1.4) 2 609.8 8.6 (7.6) 7.9 (7.0) 1.2 (1.1) 0.5 (0.5) 3 571.2 8.6 (7.6) 7.9 (7.0) 0.7 (0.7) 0.0 (0.0) 4 260.6 8.8 (7.8) 7.0 (6.3) 0.5 (0.4) -1.3 (-1.1) 5 235.1 10.5 (9.4) 9.1 (8.1) 0.4 (0.4) -0.9 (-0.8) 6 455.4 8.7 (7.7) 7.0 (6.2) 0.7 (0.6) -1.0 (-0.9) 7 741.3 7.6 (6.8) 10.8 (9.7) 1.7 (1.5) 4.9 (4.4) 8 797.3 10.0 (8.9) 11.6 (10.3) 1.5 (1.4) 3.1 (2.7) 9 382.1 5.8 (5.1) 10.2 (9.1) 6.9 (6.2) 11.3 (10.1)

10 98.6 6.3 (5.7) 6.3 (5.6) 2.3 (2.0) 2.3 (2.0) 11 261.7 7.2 (6.4) 9.7 (8.6) 1.9 (1.7) 4.5 (4.0) 12 363.9 7.1 (6.3) 5.1 (4.5) 1.5 (1.3) -0.5 (-0.4)

Average 8.4 (7.5) 8.7 (7.8) 1.6 (1.4) 1.9 (1.7) z manure P available for mechanical spreading on agricultural land.

Manure P is generally highly recoverable in intensive livestock operations (ILOs). In contrast, a pasture-based system results in relatively low recoverability. As a result, beef feedlots and hog or poultry barns have much higher P recoverability than beef cattle which are only confined for a few months each year. Estimates of manure production in Manitoba clearly show that cattle and hogs produce most of the 24 927 000 kg of manure P in Manitoba of which almost 70% is produced by cattle (Table 1.3). However, most Manitoba cattle are not in feedlots and therefore the recoverability of cattle manure P is much lower than for hog manure P. As a result, it is estimated that hogs produce over one half of the 8 751 000 kg of manure P which is available for mechanical spreading on agricultural land in Manitoba.

Between 1991 and 2001 the numbers of most types of livestock have increased in Manitoba (Table 1.4). The coefficients used by Statistics Canada to convert livestock numbers to animal units are given in Table 1.5. Beef cattle increased 35%, hogs by 65%, poultry by 19%, and other livestock by 63%. From a manure P perspective, the density of the livestock units is more important than the numbers of animals and that, too, has increased. Beaulieu and Bédard (2003)

22

reported on changes in Canadian livestock numbers from 1991 to 2001. In 1991, 92% of Manitoba livestock was in low density areas (<25 animal units km-2) and the remainder in medium density areas (25-70 animal units km-2). By 2001, 8.4% of livestock was in high density areas (>70 animal units km-2), 9.3% in medium density areas, and 82.3% in low density areas. Two Manitoba rural municipalities (RMs) in agricultural region 9, La Broquerie and Hanover, showed large increases in animal unit density over the ten years. They were ranked sixth and eleventh for livestock density in Canada in 2001 with 129 and 106 animal units km-2 respectively (Table 1.6). As livestock density increases, land for manure application that is within economic spreading distances becomes inadequate and over application of manure nutrients is more likely to occur. Nicolas et al. (2002) performed detailed P budgets for four Manitoba RMs with varying degrees of livestock density. The RMs of Hanover and La Broquerie were estimated to have annual residual values of 23 and 29 kg P/ha respectively. In contrast, the RMs of Roland (intensive cropping, minor livestock) and Sifton (extensive livestock) had estimated P balances of 0 and 2 kg P/ha respectively.

Crop yields and therefore the actual P removal will vary from year to year largely due to weather differences. Fertilizer and manure application varies with crop and soil fertility level and manure application is also determined by location of the manure source as it is not economical to transport manure large distances. The result is that every year within each of the twelve agricultural regions there will be fields with P applications greater than crop removal and fields with P application below crop removal. However, the proportion of fields with P applications above removal is greater in regions with higher livestock density and/or potato production. The risk of P loss from land to water will increase if a field consistently receives P additions above removal rates and soil P levels are allowed to build up to the point where excessive quantities of water soluble P are available for loss with runoff (Chapter 3).

23

Table 1.3. Estimated annual production of manure, manure nutrients, and recoverable manure nutrients in the Province of Manitoba for the major classes of livestock - 2001 z Number Animal Manure Nutrient excreted x Manure nutrients w Recoverable nutrients v

of animals units y produced N P N P N P tonnes ------------------------------ thousands of kg -------------------------------- Cattle 1 424 427 1 018 732 10 752 735 54 576 17 048 17 618 14 518 4 909 2 998 Hogs 2 540 220 357 797 3 912 981 22 712 6 742 5 671 5 731 4 764 4 814 Chickens 7 985 741 17 479 186 603 2 512 910 1 699 774 1 648 751 Turkeys 694 248 5 181 38 527 585 228 312 194 302 188

Provincial Total 1 399 188 14 890 847 80 385 24 927 25 300 21 217 11 623 8 751

tons ------------------------------ thousands of lbs ------------------------------ Cattle 1 424 427 1 018 732 11 828 009 120 068 37 505 38 759 31 940 10 799 6 595 Hogs 2 540 220 357 797 4 304 279 49 967 14 832 12 477 12 608 10 481 10 591 Chickens 7 985 741 17 479 205 263 5 526 2 001 3 738 1 703 3 626 1 652 Turkeys 694 248 5 181 42 380 1 287 501 686 426 665 414

Provincial Total 1 399 188 16 379 932 176 847 54 840 55 660 46 677 25 570 19 252

z Data from 2001 Manitoba Agriculture Yearbook (Manitoba Agriculture and Food 2002) and calculations according to Johnston and Roberts 2001. Please note that the calculations are based upon those of Kellogg et al. 2000 from USDA-NRCS for U.S. conditions for livestock production and manure collection, storage, handling and application. These calculations may not be representative of Manitoba conditions. y Animals units as per Kellogg et al. 2000 (=1000 lb of live weight animal) x Nutrients as excreted - estimated nutrient content of manure (including urine) as excreted from the animal w Manure nutrients - estimate of nutrients in the manure after normal losses v Recoverable nutrients - those manure nutrients which are estimated to be available for mechanical land application to crop land (less confinement of livestock is associated with lower recovery of manure nutrients)

24

Table 1.4. The distribution of Manitoba livestock by type, 1991 and 2001 (Beaulieu and Bédard 2003)

1991 2001 Beef Dairy Hog Poultry Other Total Beef Dairy Hog Poultry Other Total

--------------------- thousand animal units z -------------------- -------------------- thousand animal units -------------------- 658 111 134 48 70 1021 886 85 221 57 114 1363

Change +35% -23% +65% +19% +63% +34% ----- distribution of animal units by type (percent) ----- ----- distribution of animal units by type (percent) -----

64.5 10.9 13.1 4.7 6.8 100 65.0 6.2 16.3 4.1 8.3 100 Change +0.8% -43% +24% -13% +22% ----- percent of Canadian livestock populations ------ ----- percent of Canadian livestock populations -----

9.5 4.2 12.5 7.6 10.3 8.6 10.0 4.0 19.4 7.2 10.6 9.8 Change +5% -5% +55% -5% +3% +14%

z animal units as described by Beaulieu and Bédard 2003 (see Table 1.5)

25

Table 1.5. Animal unit (a.u.) coefficients used by Statistics Canada 2001 (adapted from Beaulieu and Bédard 2003) Livestock type Coefficient

Beef cattle

cows calves heifers feeder heifers steers bulls

1.000 0.227 0.714 0.714 0.769 1.000

Dairy cattle

cows calves heifers steers bulls

1.333 0.303 1.000 0.833 1.333

Pigs

boars sows nursing pigs growing pigs

0.200 0.200 0.125 0.033

Poultry

broilers pullets laying hens turkeys

0.005 0.003 0.008 0.012

Other livestock and poultry horses goats rabbits mink foxes bison deer elk llamas other cattle wild boars rams ewes and wethers lambs other sheep duck ostrich emu other chicken

1.333 0.143 0.025 0.013 0.025 1.000 0.125 0.600 0.143 1.000 0.250 0.143 0.200 0.063 0.143 0.020 0.143 0.063 0.010

26

Table 1.6. Canadian census regions with the highest livestock animal unit (a.u.) densities in 2001 (Beaulieu and Bédard 2003) Animals in 2001 Farm area Livestock density (a.u. km-2)

Rank Region animal units z % km2 % 1991 2001 difference

1 2 3 4 5 6 7 8 9

10 11 12

Canada B.C. B.C. Quebec N.S. Alberta Manitoba Quebec Ontario Quebec B.C. Manitoba Quebec

Fraser Valley R.D. Greater Vancouver R. D. La Nouvelle-Beauce Digby County Lethbridge County La Broquerie Matawinie Waterloo R.M. Desjardins Cowichan Valley R.D. Hanover La Haute-Yamaska

13 954 500

177 500

71 500 80 800 7 300

427 000 38 300 19 500

113 800 14 200 14 400 76 100 34 600

100

1.3 0.5 0.6 0.1 3.1 0.3 0.1 0.8 0.1 0.1 0.5 0.2

674 800

490 390 510

50 2 980

300 150 910 120 130 720 370

100

0.1 0.1 0.1 0.0 0.4 0.0 0.0 0.1 0.0 0.0 0.1 0.1

304 179 162

77 62 39

135 122

97 84 62 95

365 183 157 145 143 129 126 125 118 107 106

94

61 4

-5 68 82 90 -9 2

22 22 44 -1

z animal units as defined in Table 1.5 Note: Due to rounding, figures may not add up to totals. Source: Statistics Canada, derived from the 1991 and 2001 Censuses of Agriculture.

27

1.6.3 In-Stream Processes In-stream processes refer to the natural processes that govern P storage and transformation within the stream channel. There are primarily three in-stream processes acting in the river channel: erosion, plant uptake and water interactions in the hyporheic zone. This latter process results from the fact that river channels are leaking conduits. Water, flowing in these porous channels, is constantly interchanging and interacting with the waters in the saturated bed and bank materials. This zone of interchange, which depends on the type and porosity of the bed and bank material and local hydraulic gradients, is referred to as the hyporheic zone. The processes of erosion, plant uptake and water interchange within the bed and bank are interlinked, and their combined effect is to store and change the type of P in the channel. The understanding of these processes is crucial to the accounting and modelling of P in the stream channel.

McCallister and Logan (1978), working on the Maumee River in Ohio, reported that the adsorptive capacity of bottom sediments was found to be higher than that of soil or, for that matter, its clay fraction. In a similar vein, McDowell and Sharpley (2001) measured TP in stream bank sediments (417 mg/kg) above the saturated zone and bed sediments (281 mg/kg), but suggest that erosion of bank sediments should contribute less P in the stream system compared with re-suspension of channel bed sediment.

Indeed, Stottlemeyer and Toczydlowski (1991) pointed to an interesting process that happens during the recession limb of flood flows. As flood water rises in the stream the bank soil becomes saturated. As the flood water recedes, a perched saturated soil condition is established in the river banks. The authors claim this exerts a “piston effect” on the ground water flow to the river, increasing solute transport from the hyporheic zone. Kjartanson (1983) reported that the saturated soil condition occurs in the banks of the Red and Assiniboine Rivers after spring high flows and together with decreased support for the bank during the recession of the flood, is a major cause of bank failure along the rivers. In a study of the Blue Earth River in Minnesota, Sekely et al. (2002) reported that bank slumping is a major source of sediment to the river, supplying 31- 44% (28 047 – 43 307 tonnes) of the total suspended sediment and 7-10% (12-17 tonnes) of the annual TP load. The Blue Earth River, with a drainage area of 6300 km2, is much smaller than the Red River (drainage area 278 000 km2). These processes, the piston effect and bank failure, may in part account for the unexplained increased loading along the St. Norbert to Selkirk reach of the river referenced in section 1.5.

In a study of the fluvial sediment of the Winooski River in Vermont by McDowell et al. (2002) reported that impoundment (731 mg TP/kg) and reservoir sediments (803 mg TP/kg) had greater TP concentrations than river sediment (462 mg TP/kg). This was attributed to more fines in impoundments and reservoirs than in river sediments, a fact, the authors state, which may also decrease the ability of impoundment sediments to release P to solution. McDowell et al. (2002) also reported on the role of sediment on the P regime of the Winooski River in Vermont. The authors state that during base flow periods, P concentrations in the river are controlled by the release of P by sediments, whereas in storm flow P concentrations are controlled by overland runoff and erosion of upland areas. In contrast, Kelly et al. (1999), while researching total maximum daily loads (TMDLs) for the Tulatin River in Oregon, suggested that during non-runoff periods the sources of P in the tributaries are considered to be primarily from groundwater. At the same time, the authors conceded that other possible non-runoff period sources are seepage from agricultural fields, releases from in-stream sediment, and anthropogenic sources such as illicit discharges. Both of these studies illustrate the complexity

28

involved in studying the processes of P release, which operate concurrently and coincidently, from the channel bed.

It is important to note that those areas of the river where settling of sediments may be occurring, impoundments, backwaters and shallows, are also the areas where both floating and rooted plant communities are established. In a series of reports commissioned by the City of Brandon, North South Consultants provide an interesting insight into the role of the aquatic plant community in the Assiniboine River. Toews (2002), and Toews et al. (1999, 2000 and 2001) reported on the establishment of a vibrant aquatic plant community downstream of Brandon, as mentioned previously in section 1.4. This community establishes itself during low flow conditions in the early summer and continues to grow and change over the course of the summer in response to flow, temperature and nutrient supply. The diverse community ranges from algae, to periphyton and rooted and floating macrophytes. The authors also reported on a successional sequence through the summer months and that this community is capable of reducing the TP concentration by plant uptake. They note that this community is often uprooted during high flows in the summer and finally in the fall, with observed algal and plant mats drifting downstream. Due to the periodic nature of the water sampling downstream, this organic P load escapes detection and results in better water quality conditions being reported downstream.

1.7 Temporal Distribution of P in Manitoba’s Waters In western Canada, more than 85% of the total annual runoff from agricultural watersheds is snowmelt runoff (Nicholaichuk 1967). Therefore, hydrological regimes of Manitoba’s rivers are dominated by the spring snowmelt high flows followed by gradually declining flows through summer and winter. The declining flows through the summer months are often interrupted by small peak flows associated with rainfall storms. Rainfall flows seldom equal the spring snowmelt discharge. For example, Figure 1.10 showing the 1996 hydrograph for Birdtail Creek, illustrates this regime and clearly shows the dominance of snowmelt runoff.

Birdtail Creek

0

5

10

15

20

25

30

1-Jan

15-Jan

29-Jan

12-Feb

26-Feb

11-Mar

25-Mar

8-Apr

22-Apr

6-May

20-May

3-Jun

17-Jun

1-Jul

15-Jul

29-Jul

12-Aug

26-Aug

9-Sep

23-Sep

7-Oct

21-Oct

4-Nov

18-Nov

2-Dec

16-Dec

30-Dec

1996

Dis

char

ge (c

ms)

Birdtail Creek

Figure 1.10. 1996 Annual Hydrograph of Birdtail Creek. Data adapted from Water Survey of Canada.

29

One would then expect the annual TP budget to parallel the discharge record. And indeed McCullough (2001) reported that the highest annual yields of TP on the Red River are associated with the three highest floods of the last thirty years.

Figure 1.11 shows the annual and monthly TP loads for the Assiniboine River at Brandon for the period 1994 to 2001. The figure clearly demonstrates that the seasonal hydrologic regime is dominated by the spring snowmelt discharge, which usually occurs in the months of April and May. In 2000, a snowmelt low flow year, the April and May discharge only contributed 8% of the annual TP load of 97 tonnes. On the other hand, in 1995, a year with a high snowmelt flood, 75% of the annual TP load of 649 tonnes was delivered in the months of April and May, dwarfing the annual load of 97 tonnes for 2000.

Figure 1.11. Monthly TP loads on the Assiniboine River at Brandon from 1994 to 2002. Data adapted from Manitoba Conservation and Water Survey of Canada.

Figure 1.12 depicts the monthly TP loadings on the Red River at Emerson, St. Norbert and Selkirk. The large TP loads carried in the spring are obvious.

0

1000

2000

3000

4000

1994 1995 1996 1997 1998 1999 2000 2001

Emerson

St. Norbert

Selkirk

Load ( t onnes)

TP Loads on the Red River

3000-4000

2000-3000

1000-2000

0-1000

Brandon: Monthly TP loads

0

50

100

150

200

250

300

350

1994

1995

1996

1997

1998

1999

2000

2001

Load

(ton

nes)

Monthly Load

Figure 1.12. Monthly TP loads on the Red River at Emerson, St. Norbert and Selkirk for 1994 to 2001. Data adapted from Manitoba Conservation and Water Survey of Canada.

30

Unfortunately, the literature on Manitoba’s rivers lacks analysis of the extent to which the spring snowmelt process is responsible for the annual TP load in Manitoba. Nor is there any discussion on the effects of the snowmelt regime on P transport processes in Manitoba. However, the literature on snowmelt transport of P for the U.S. and northern Europe, although not rich, helps to point out what to look for in our data.

Becher et al. (2000) reported a similar seasonal pattern for the concentration and loads of TP with their work on twelve rivers and creeks in eastern Iowa for 1996-97. This study is also interesting from a scale point of view when compared to the Red River watershed. The three largest rivers in eastern Iowa, with agriculture constituting 93% of all land use in these watersheds, transport an estimated 6800 tonnes of P annually to the Mississippi River, an amount paralleling the P export by the Red River watershed to Lake Winnipeg (TP 4900 tonnes/yr). Likewise, Longabucco and Rafferty (1989) reported that half of the annual TP load on the Oak Orchard Creek watershed in western New York state was delivered by spring snowmelt runoff. The authors stated that runoff during the late winter-early spring period is the most important hydrological factor responsible for the annual P load to Lake Ontario. And according to Hatch et al. (2001), Lake Tahoe streams deliver 75% of the annual TP load to the lake during the spring snowmelt period. Moreover, they contended that the effects of human impacts on the watersheds are more prevalent during the high water years.

Johannessen and Henriksen (1978) measured pollutant release from the snowpack at three sites in southern Norway during the winter of 1974. They found that 50 – 80% of the pollutant load, including TP and soluble P, was released with the first 30% of the meltwater. The average concentration of the first 30% of the meltwater was two to two and one half times as high as the concentration in the snowpack itself. Similarly, Rekolainen (1989) investigated the effect of snow and ground-frost melt on concentrations of P and suspended solids in two agricultural watersheds in southern Finland. Unlike rainfall events, where TP concentration is strongly correlated with discharge and suspended sediments, Rekolainen found that the high P concentrations measured during snowmelt were not correlated with discharge. During the first days of the snowmelt period the TP concentrations were very high (0.3 –1.2 mg/L). As the snowmelt continued, with cultivated fields still partly snow-covered, TP concentrations decreased to 0.3 –0.5 mg/L. After the snow in the cultivated fields had melted and soil frost had began to thaw, the TP concentrations increased again (0.3 –0.5 mg/L). This secondary surge correlated with the peak discharge and suspended sediment concentrations. Also, the proportion of soluble reactive P was higher during the snowmelt period than during the frost melt period. This characteristic of the P melt water hydrograph, initially high TP concentrations correlating with initial melt of the snowpack, followed by a dip and a later surge during the ground thaw has been reported by Lake and Morrison (1977) and Lathrop (1986). Uhlen (1989) also reported this characteristic for a fertilized grassland watershed in Norway, and derives a regression equation to explain the amount of P loss during snowmelt. Altogether, snow runoff depth, grass residue in the fall and amount of P fertilizer applied in the previous year accounted for most of the P loss (R2=0.86).

Johannessen and Henriksen (1978) argued that the high TP and DP concentrations in the initial onslaught of snowmelt runoff are created by what they call a ‘freeze concentration’ process: the melting and refreezing of crystals which preferentially causes contaminants to accumulate at the surface of the crystals. Rekolainen (1989), on the other hand, suggested that the high TP concentrations are caused by continuous extraction of P from the soil surface and plant residue

31

cover by low ionic concentration meltwater. Both of these mechanisms have value and therefore merit consideration. Whether similar patterns of DP-TP occur in Manitoba waters is impossible to discern because of the length of time between the regular water quality sampling periods.

Manitoba scientists do not have available to them the snowmelt discharge data that is at the disposal of the Europeans and Americans. However, by looking at the flow and water quality data that does exist for Manitoba, the magnitude of P in the snowmelt floods and the need for more precise event-based data is apparent.

Water quality samples are routinely collected by Manitoba Conservation, generally biweekly, monthly or quarterly, often dependent on projects or public requests for sampling. Larger rivers like the Red are sampled on a monthly basis, while smaller streams like the Boyne River are sampled quarterly. Event-based water quality sampling is not routinely collected by Manitoba Conservation and, therefore, the dynamics of either snowmelt or rainfall floods are not readily determinable. For example, the four samples collected from the Boyne River in 1996 (Table 1.7), completely missed the spring snowmelt flood which occurred between April 12 to June 16. Figure 1.13 illustrates how critical this flood was to annual water budget and the annual P budget. From the few biweekly records available for small streams in Manitoba, the concentrations of TP during this spring snowmelt period are usually greater than concentrations at other times of the year. Therefore, it is impossible to accurately portray the annual P budgets for streams in Manitoba. Also, there is a problem in the coordination of the historical flow and water quality data. Often the water quality data is collected where no flow data exists or vice versa.

Figure 1.13. 1996 Annual Hydrograph for the Boyne River. Data adapted from Water Survey of Canada.

Table 1.7. TP samples collected on the Boyne River, 1996

Date Discharge (cm) TP (mg/L) 15-Jan-96 0.05 0.093 9-Apr-96 0.09 0.155 24-Jun-96 0.08 0.081 7-Oct-96 0.19 0.119

Boyne River

0

5

10

15

20

25

30

35

40

45

50

1-

Jan

15-

Jan

29-

Jan

12-

Feb

26-

Feb

11-

Mar

25-

Mar

8-

Apr

22-

Apr

6-

May

20-

May

3-

Jun

17-

Jun

1-

Jul

15-

Jul

29-

Jul

12-

Aug

26-

Aug

9-

Sep

23-

Sep

7-

Oct

21-

Oct

4-

Nov

18-

Nov

2-

Dec

16-

Dec

30-

Dec

1996

Dis

char

ge (c

ms)

Boyne River

32

There are few small rivers or streams in Manitoba with a continuous flow record and biweekly sampling record that can be used to illustrate the seasonal P loading regime. In 1996, the La Salle River was sampled every two weeks, on average, from May to September at Sanford. The TP loading for the La Salle River during 1996 is shown in Figure 1.14. The spring flood of 1996 is a one in ten year event. Clearly, this shows the importance of the spring flood flows to the TP export regime on the La Salle River.

La Salle River TP load

0

1000

2000

3000

4000

5000

6000

1-Jan

1-Feb

1-Mar

1-Apr

1-May

1-Jun

1-Jul

1-Aug

1-Sep

1-Oct

1-Nov

1-Dec

1996

Load

(kg)

La Salle River TP load

Figure 1.14. 1996 Annual TP load hydrograph for the La Salle River. Data adapted from Water Survey of Canada.

1.8 Conclusions The literature clearly establishes that 60% of the TP load carried by the Red River to Lake Winnipeg originates from outside the province. Of the 40% generated from within the province, agricultural and urban sources contribute major amounts. Although establishing the exactly proportions is difficult, the evidence suggests even splits between agriculture, urban and background amounts. Great uncertainty exists in the ability to estimate total loads in the system as is evident from lack of mass balance across major river confluences. Apportionment of those measured amounts to their source contribution remains an even greater challenge. The majority of point source effluents, such as municipal and industrial discharges, are based only on estimates due to the lack of data collected on these sources within the Red River watershed. The same constraint also applies to estimates of agricultural sources. Finally, urban non-point sources, especially along the Winnipeg sections of river, lack data entirely. Further scientific investigation of these major sources is required in order to fully quantify the loads of P to Manitoba’s surface waters and the apportionment of the source contributions.

Within the portion of P that is transferred from agricultural land to Manitoba’s rivers and streams, only a small part of that P originates from land on which livestock manure has been

33

applied. However, with continuing growth in Manitoba’s livestock industry, the concern for environmentally responsible practices for land application of manure is justified. Of major concern to Manitoba is the finding that a large majority of the P load is carried in the spring by snowmelt driven high flow events. How this will affect management practices remains to be seen. Both agricultural and urban runoff processes during snowmelt are less well understood than during rainfall events. For example, European research shows that the role of leaching of soluble P from vegetative material and crop residues may be relatively large in snowmelt runoff, compared to rainfall runoff. The lack of literature related to the mechanisms of interactions between P with snowmelt runoff indicates that more research is required to determine if management practices adopted to mitigate P runoff from non-point sources are to be effective. What is clear from the data, however, is the strong correlation between P loading to streams and runoff quantity. It is obvious that runoff serves as the major transport vector for P from non-point source land. Increasing the efficiency of agricultural and urban land drainage has had the effect of decreasing the time required for major runoff events to travel to receiving waters and has increased the total quantity of runoff. This perpetuates both nutrient losses and peak flow discharges the latter being responsible for damaging floods. Obviously, requiring land owners to hold back runoff water from drainage ditches and storm sewers would have a negative impact. For farmers, crop yields would be reduced and seeding dates would be advance further into the spring. However, measures designed to decrease the speed at which water travels to receiving water, such as impoundment and constructed wetlands, should have the effect of decreasing peak discharge and P loadings to streams. Increasing the hydraulic efficiency of land drainage through ditch straightening, lengthening and smoothing has the effect of decreasing the distance rainfall has to travel laterally to arrive at a drain. This increase in drainage density has the effect of increasing the contributing area of rainfall-runoff generation. Frozen soil has a similar impact since by decreasing the water holding capacity of soils, the area of the landscape capable of contributing to direct runoff increases. Practices, such as winter spreading of manure, which provide direct contact of P to runoff water under the combined influence of increased hydraulic efficiency and frozen ground should be avoided. However, tempering this conclusion is the need for increased research in the field of frozen soil infiltration. Evidence collected from prairie environments suggests that low fall soil moisture at freeze-up can have a dramatic influence on the quantity of snowmelt runoff that may be retained in the soil. For this reason, hard and fast rules regarding application of nutrients to agricultural lands may not be appropriate and should be based on soil type, frozen soil moisture conditions and aerial snow quantities. Investigating the amount of P in precipitation also requires greater attention, considering the findings of the Clear Lake studies conducted by Thompson et al. (2000) and Chubak et al. (2001). These studies indicate surprisingly high concentrations of P in snowfall. If this holds true for the remainder of the province, then the case for more P research in winter is further underscored. High P levels in early winter precipitation may, in fact, point to management practices that reduce wind erosion of soil and straw ash following crop removal in the fall. In-stream processes also require investigation, in particular stream bank erosion, sedimentation and scour of river channel sediments, the incorporation of P into lake sediments, the role of aquatic plant communities on the seasonal storage of P and water interactions with P in the hyporheic zone. All of these affect P concentrations and forms of P in water due to changes in chemistry and storage-release mechanisms.

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Ultimately it is the forms and quantity of P in the aquatic environment that determine its impact on ecosystem health. Determining, in a more rigorous fashion the quantity/form combinations of phosphorus that can be assimilated in a sustainable fashion should be a research priority. Transferring concentration values from other jurisdictions without adequate validation in a Manitoba context could be detrimental. In a similar vein, the implementation of costly management strategies, aimed at either at-source or end-of-pipe controls, without sound estimates of their ultimate impact on aquatic ecosystems may prove fruitless and even detrimental to the desired impact.

All of the above investigations would be better served if increased water quality sampling frequency for total P and dissolved P was available for snowmelt floods and rainfall storm flows. Manitoba Conservation should strongly consider increasing the sampling frequencies of these data for major streams in the province and support research necessary to characterize a representative sample of headwater basins and stream reaches distributed throughout the province. Not only could these be used to gain a greater understanding of P runoff processes in Manitoba, but may also be useful to extend these distributions of P loadings to the historic record.

1.9 Abbreviations and Definitions Bioavailable P (orthophosphate, citrate soluble, citrate acid extractable) is inorganic phosphorus that is readily “available” for plant use.

DP or dissolved Pis operationally defined as that fraction of P that passes through 0.45 micron filter.

Inorganic P is derived from parent rock during soil formation or is mined; is mostly orthophosphates.

Labile P is the quickly available portion of P that exists in equilibrium between immediately available (inorganic) and very slowly available (organic and precipitated) P where upon a slight change in surrounding conditions (moisture, temperature, or pH, for example) it becomes available to plants.

Nonlabile P is unavailable P that is in organic adsorbed or precipitated forms.

Organic P is formed from gradual uptake and use by plants, animals, and microorganisms converting inorganic P in their cells.

Orthophosphate is commonly used interchangeably, though not quite correctly, with available P. It is an inorganic anion or salt of orthophosphoric acid that contains one P atom (not a group of two or more P atoms which is polyphosphate).

Oxidation Reduction Potential (Redox Potential) a measurement of the ease with which a substance either acquires electrons (reduction) or releases electrons (oxidation). PP or particulate Pis operationally defined as that fraction of P that is retained by a 0.45 micron filter. Reactive P is the same as Bioavailable P. Residual P or Residual organic P is P that is not extracted by preceding treatments.

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Soil test P estimates the amount of P that can be extracted from the soil by plant roots by using acids or bases to cleave all but the most insoluble forms of P, giving an indication of the quantity of P that plants may use during a growing season.

Soluble P or water soluble P is the fraction of P when extracted with water passes through 0.45 µm filters; water or weak acids or salts of weak acids extraction gives an indication of intensity of runoff after a rain. Most commercial fertilizers are readily water-soluble.

Soluble reactive P (SRP) is soluble P (the filtrate, 0.45 µm filter, of water or weak acid extraction) that responds to colorimetry. Usually, this means the extract that reacts with the ascorbic acid and molybdate chemicals in the Murphy-Riley method.

Triple Superphosphate, calcium dihydrogen phosphate, is formed by treating phosphate rock with phosphoric acid to create an inorganic fertilizer containing about 20% P (45% P2O5).

TMSNL is Total Measured Stream Nutrient Load. TOC is total organic carbon, a summation of all organic carbon in a given system. TN is total nitrogen, a summation of all nitrogen compounds in a given sytem.

TP is total phosphorus, a summation of all phosphorus in a given system. WWTF is Waste-Water Treatment Facilities.

1.10 References Ahlgren. I., T. Frisk and Kamp-Nielsen. 1988. Emperical and theoretical models of phosphorous loading, retention and concentration vs lake trophic state. Hydrobiologia 170: 285-303. Bannerman, R. T., Owens, D. W., Dodds, R. B. and Hornewer, N. J. 1993. Sources of pollutants in Wisconsin stormwater. Water-Science-and-Technology 28(3-5): 241-159. Barten, J. M. 1994. Soil fertility level of 181 lawns in four municipalities in the Twin Cities Metropolitan Area. Report prepared for Suburban Hennepin Regional Park District. 8pp. Barten, J. M. and Jahnke, E. 1997. Suburban lawn runoff water quality in the Twin Cities Metropolitan area, 1996 and 1997. Report prepared for Suburban Hennepin Regional Park District. 18pp. http://www.dnr.state.wi.us/org/water/wm/dsfm/shore/documents/lawnrunoff.pdf Beaulieu, M. S. and Bédard, F. 2003. A geographic profile of Canadian livestock, 1991-2001. Statistics Canada, Agriculture Division, Ottawa, ON. [Online] Available: http://www.statcan.ca/english/IPS/Data/21-601-MIE2003062.htm [2003 April 10]. Becher, K.D., Schnoebelen, D.J., and Akers, K.K., 2000. Nutrient concentrations and yields in surface water in Eastern Iowa: J. Amer. Water Resources Association 36 (1): 161–173. Beck, A. E. 1985. Recreational development capacity study of twelve lakes in the south Riding Mountain Planning District. Department of Environment and Workplace Safety and Health. Water Standards and Studies Report 87-5. Belke, S. E., McGinn, R. A. and Rousseau, P. 2001. Anthropogenic nutrient loading in the winter snowfalls of the Clear Lake watershed. Technical Report. Riding Mountain National Park. Wasagaming, Manitoba. 37pp. Bennet, E. M. Reed-Anderson, T. Houser, J.N., Gabriel, J.R. and Carpenter, S.R. 1999. A phosphoprus budget for the Lake Mendota watershed. Ecosystems 2:69-75

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Bennet, M. 1985. personal communication in: Beck 1985. Boström, B., Persson, G. and Broberg, B. 1988. Bioavailability of different phosphorus forms in freshwater systems. Hydrobiologia, 170, 133-155. Bourne, A., Armstrong, N. and Jones, G. 2002. A preliminary estimate of total nitrogen and total phosphorus loading to streams in Manitoba, Canada. Manitoba Conservation Report No. 2002-04: 49pp. Brigham, M.E., Mayer, T., McCullough, G.K. and Tornes, L. H. 1996. Transport and Speciation of Nutrients in Tributaries to Southern Lake Winnipeg, Canada: Abstracts of Presentations from 16th Annual International Symposium on Lake, Reservoir and Watershed Management, Nov. 13-16, 1996, Minneapolis/St. Paul, MN, North American Lake Management Society, p.64. Busman L., Lamb, J., Randall, G., Rehm, G. and Schmitt, M. 2002. The nature of phosphorus in soils. in Phosphorus in the agricultural environment. [Online] Available: http://www.extension.umn.edu/distribution/cropsystems/DC6795.html Chambers P. A., Guy, M., Roberts, E. S., Charlton, M. N., Kent, R., Gagnon, C., Gagnon, G., Grove, G. and Foster, N. 2001. Nutrients and their impact on the Canadian environment.Agriculture and Agri-Food Canada, Environment Canada, Fisheries and Oceans Canada, Health Canada, and Natural Resources Canada. 241 pp. Chambers, P. A and Dale, A. R. 1997. Contribution of industrial, municipal, agricultural and groundwater sources to nutrient export, Athabasca, Wapiti and Smoky rivers, 1980 to 1993. Northern River Basins Study, Edmonton AB. Cooley, M., Schneider-Vieira, F. and Towes, J. 2001. Assiniboine River monitoring study water quality component water quality assessment and model for the open water season. North/South Consultants Inc. Winnipeg, MB. 156 pp. plus appendices. Creason, J.R. and C. F. Runge. 1992. Use of lawn chemicals in the Twin Cities. Public Report Series #7. Water Resources Research Center, University of Minnesota. 21pp. Daniel, T.C., Sharpley, A.N., and Lemunyon, J.L. 1998. Agricultural phosphorus and eutrophication: a symposium overview. J. Environ. Qual. 27:251-257. Devai, I. and Delaune, R. D. 1995. Evidence for phosphine production and emission from Louisiana and Florida marsh soils. Organic Geochemistry. 23: 277-279. Devai, I., Felfoldy, L., Wittner, I and Plosz, S. 1988. Detection of phospine : new aspects of the phosphorus cycle in the hydrosphere. Nature 333: 343-345. Eismann, F. Glindemann, D., Bergmann, A. and Kusch, P. 1997. Soils as source and sink of phospine. Chemosphere 35 (3):523-533. Engle, D.L. and Sarnelle, O. 1990. Algal use of sedimentary phosphorus from an Amazon floodplain lake: Implications for total phosphorus analysis in turbid waters. Limnol. Oceanogr., 35, 483-490. Fixen, P. E. 2002. Soil test levels in North America. Better Crops 86: 12-15. Froelich, P.N. 1988. Kinetic control of dissolved phosphate in natural rivers and estuaries: A primer on the phosphate buffer mechanism. Limnol. Oceanogr., 33, 649-668. Green, D.J and Turner, W. N. 2002. South Tobacco Creek manured watershed runoff study, 1998-2001. Manitoba Conservation and Deerwood Soil & Water Management Association. Manitoba Conservation Report No. 2002-02.

http://www.extension.umn.edu/distribution/cropsystems/DC6795.html

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Hatch, L. K., Reuter, J.E. and Goldman, C.R. 2001. Stream phosphorus transport in the Lake Tahoe basin. Environmental Monitoring and Assessment 69 (1): 63-83. Haygarth, P. M. and Sharpley, A. N. 2000. Terminology for phosphorus transfer. J. Environmental Qual. 29:10-15. Hegman, W., Wang, D. and Borer, C. 1999. Estimation of Lake Champlain basinwide nonpoint source phosphorus export. Lake Champlain Basin Program Technical Report No.31, Grand Isle, VT. Ji-Ang, L., Yahui, C. H., Kuschk, P. Eismann, F. and Glindemann, D. 1998. Phosphine in the urban air of Beijing and its possible sources. Water, Air,and Soil Pollution 116. 597-604. Johannessen, M. and Henriksen, A. 1978. Chemistry of snow meltwater: Changes in concentration during melt. Water Resources Research 14(4): 615-619. Johnston, A. M., and Roberts, T. L., 2001. High soil phosphorus -- Is it a problem in Manitoba? Second annual Manitoba Agronomists Conference, 2001, pp.74-82. Jones, G. and Armstrong, N. 2001. Long-term trends in total nitrogen and total phosphorus concentrations in Manitoba Streams. Water Quality Management Section, Water Branch. Manitoba Conservation Branch Report No 2001-7. 154 pp. Kalkhoff, S. J., Barnes, K.K., Becher, K.D., Savoca, M.E., Schnoebelen, D.J., Sadorf, E.M., Porter, S.D. and Sullivan, D.J. 2000. Water Quality in the Eastern Iowa Basins, Iowa and Minnesota, 1996–98. USGS Water Resources Circular 1210: 45 pp. Kellogg, R.L., Lander, C. H., Moffitt, D. C. and Gollehon, N. 2000. Manure nutrients relative to capacity of cropland and pastureland to assimilate nutrients: Spatial and temporal trends for the United States. USDA-NRCS Publication No. nps00-0579 [Online] Available: http://www.ftw.nrcs.usda.gov/nps/gsa.htm [2003 April 3]. Kelly, V.J., Lynch, D.D., and Rounds, S.A., 1999. Sources and Transport of Phosphorus and Nitrogen During Low-Flow Conditions in the Tualatin River, Oregon, 1991-93: U.S. Geological Survey Open-File Report 99-232. Kjartanson, B. 1983. Geological engineering report for urban development of Winnipeg. Department of Geological Engineering, University of Manitoba, Cantext Publications, Winnipeg, Manitoba. 78pp. plus maps. Korol, M. and Rattray, G. 1997. Canadian fertilizer consumption, shipments and trade 1995/1996. Agriculture and Agri-Food Canada. [Online] http://www.agr.gc.ca/policy/cdnfert/cdnfert9596/text.html [2003 April 3]. Korol, M. and Rattray, G. 2001. Canadian fertilizer consumption, shipments and trade 1999/2000. Agriculture and Agri-Food Canada. [Online] http://www.agr.gc.ca/policy/cdnfert/text99-00.pdf [2003 April 3]. Lake, J. and Morrison, J. 1977. Environmental impact of land use on water quality. U.S. Environmental Protection Agency 905-77-007-B Lathrop, R. C. 1986. A simplified method for obtaining monitored phosphorus loadings. in: Redfield, G. Taggert, J.F. and Moore, L. M., Lake and Reservoir Management 2(5), Proc. Ann. Conc. Int. Symp. N. Amer. Lake Management. 20-26. Longabucco, P. and Rafferty, M. R. 1989. Delivery of nonpoint source phosphorus from cultivated mucklands to Lake Ontario. J. environmental Quality 18 (2): 157-163.

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Manitoba Agriculture and Food. 2001. Manitoba agriculture yearbook, 2000. MAF, Winnipeg, MB. Manitoba Agriculture and Food. 2002. Manitoba agriculture yearbook, 2001. MAF, Winnipeg, MB. Manitoba Conservation. 2003. Water Quality Management Section. Nutrient data. Winnipeg, MB. McCallister, D. L. and Logan, T. J. 1978. Phosphate adsorption-desorption characteristics of soil and bottom sediments of the Maumee River basin of Ohio, J. Environmental Quality 7: 87-92. McCullough, G. 2001. Organic carbon, nitrogen and phosphorus fluxes in rivers flowing into and out of Lake Winnipeg. Canada Department of Fisheries and Oceans, Winnipeg, Manitoba. 47pp. McDowell, R.W., and Sharpley, A.N. 2001. A comparison of fluvial sediment phosphorus (P) in relation to location and potential to influence stream P concentrations. Aquatic-Geochemistry. 7(4): 255-265. McDowell, R.W., Sharpley, A.N. and Chalmers, A.T. 2002. Land use and flow regime effects on phosphorus chemical dynamics in the fluvial sediment of the Winooski River, Vermont. Ecological-Engineering 18(4): 477-487. Mitchell, P. 1985. Preservation of water quality in Lake Wabamun. Alberta Environment, Pollution Control Division, Water Quality Control Branch, Emonton, Alberta. Nicolas, L., Jr., Small, D.,Racz, G., Abbott, D., Hodgkinson, D., Liu, C. and Warkentine, G. 2002. Study of regional nutrient balances in four municipalities in Manitoba. Report to MLMMI Inc., DGH Engineering, St. Andrews, MB.[Online] Available: http://www.manure.mb.ca/projects/completed/pdf/02-hers-04-regional-nutrient-balances.pdf [2003 April 3]. Nikolaichuk, W. 1967. Comparative watershed studies in southern Saskatchewan. Trans. Am. Soc. Agric. Eng. 10(4):502-504. Oberts, G. L., Marsalek, J. and Viklander, M. 2000. Review of water quality impacts of winter operations on urban drainage. Water Quality Research journal of Canada 35(4): 781-808. Ormernik, J. M. 1977. Non-point source stream nutrient level relationships: a nation-wide study. Corvallis Environmental Laboratory, U.S. Environmental Protection Agency, Corvallis, Oregon. 151pp. Pettersson, K., Boström, B. and Jacobsen, O. S. 1988. Phosphorus in sediments-speciation and analysis. Hydrobiologia, 170, 91-101. Pipp, E. 2002. A brief on the downstream impacts of the City of Winnipeg wastewater treatment plant effluents. A submission to the Manitoba Clean Environment Commission Public Hearing on the City of Winnipeg Wastewater Collection and Treatment Systems. 35 pp. Prairie, Y.T., Duarte, C. M. and Kalff, J. 1989. Unifying nutrient chlorophyll relationship in lakes. Canadian Journal of Fisheries and Aquatic Science 46: 1176-1182. Rekolainen, S. 1989. Effects of snow and soil frost melting on the concentrations of suspended solids and phosphorus in two rural watersheds in western Finland. Aquatic Sciences 51(3): 211-223.

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Rekolainen, S., Ekholm, P. Ulen, B. and Gustafson, A. 1997. Phosphorus losses from agriculture to surface waters in Nordic countries. In H. Tunney, O.T. Carton, P.C. Brookes and A.E. Johnston, eds. Phosphorus loss from soil to water. CAB International, Wallingford, UK. 467pp. Rekolainen, S., Posch, M., Kamari, J. and Ekholm P. 1991. Evaluation of the acuracy and precision of annual hosphorus load estimates from two agricultural basins in Finland. J of Hydrology, 128: 237-255. Richardson, C.J. and Sundareshwar, P. V. 2002. Role of Sediment Processes in Regulating Water Quality of the Cape Fear River. U.S. Geological Survey (in press). Schindler, D.W., Newbury, R.W., Beaty, K.G. and Campbell, P. 1976. Natural water and chemical budgets for a small precambrian lake basin in central Canada. Journal of the Fisheries Research Board of Canada. 33: 2526-2543. Schindler, D. W. 1977. Evolution of phosphorus limitation in lakes. Science 195:260-262. Sekely-A.C., Mulla, D. J. and Bauer D.W. 2002. Streambank slumping and its contribution to the phosphorus and suspended sediment loads of the Blue Earth River, Minnesota. J. of Soil and Water Conservation 57(5): 243-250. Sharpley, A. N. and Rekolainen, S. 1997. Phosphorus in agriculture and its environmental implications. Pages 1-53 in H. Tunney, O. T. Carton, P. C. Brookes and A. E. Johnston, eds. Phosphorus loss from soil to water. CAB International, Wallingford, UK. Stainton, M., Salki, A., Hendzel, L. and Kling, H. 2003. Evidence From Ecosystem Research by Fisheries and Oceans Canada For The Need To Protect Lake Winnipeg From Phosphorus Derived From The Red River Basin. A submission to the Manitoba Clean Environment Commission Public Hearing on the City of Winnipeg Wastewater Collection and Treatment Systems, Freshwater Institute, Fisheries and Oceans Canada. Statistics Canada. 1997. Agricultural profile of Manitoba – Part 1. Statistics Canada, Ottawa, ON. Statistics Canada. 2002. Farm data: Initial release. Publication 95F0301XIE, Statistics Canada Internet Site [Online] http://www.statcan.ca/english/IPS/Data/95F0301XIE.htm [2003 April 3]. Stottlemeyer, R. and Toczydlowski, D. 1991. Stream chemistry and hydrologic pathways during snowmelt in a small watershed adjacent to Lake Superior. Biogeochemistry- Dordrecht 13(3): 177-198. Thoman, R. T.and Mueller, J. A. 1987. Principles of surface water quality modelling and control. Harper Collins Publishers Inc., New York, NY., 644pp. Thompson, M.E., McGinn, R.A. and Rousseau, P. 2000. Nutrient Loading in the Winter Snowfalls over the Clear Lake Watershed. Technical Report. Riding Mountain National Park. Wasagaming Manitoba. 36pp. Thorolfsson, S. T. and Brandt, J. 1996. The influence of snowmelt on urban runoff in Norway. Pages 133-138 in Proc. of 7th Int. Conf. on Urban Storm Drainage, Hanover, Germany. Towes, J. 2002. Assiniboine River monitoring study water quality component progress report-October 2002. A report prepared for the City of Brandon. 128pp. Towes, J., Coooley, M. and Schneider-Vieira, F. 1999. Assiniboine River monitoring study water quality component progress report-October 1999. A report prepared for the City of Brandon. 86pp.

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Towes, J., Coooley, M. and Schneider-Vieira, F. 2000. Assiniboine River monitoring study water quality component progress report-April 2000. A report prepared for the City of Brandon. 86pp. Towes, J., Coooley, M. and Schneider-Vieira, F. 2001. Assiniboine River monitoring study water quality component progress report-April 2001. A report prepared for the City of Brandon. 78pp. Trew, D.O., Beliveau, D.J. and Young, E. I. 1987. Lake Baptiste study: a summary report. Alberta Environment Pollution Control Division, Water Quality Control Branch, Emonton, Alberta. Uhlen, G. 1989. Surface runoff losses of phosphorus and other nutrient elements from fertilized grassland. Norwegian J. of Agricultural Sciences 3(1): 47-56. Vollenweider, R.A. and Kerekes, J. 1980. The loading concept as a basis for controlling eutrophication, philosophy and preliminary results of the OECD programme on eutrophication. Prog. Water Technol., 12, 5-38. Waschbusch, R.J., Selbig W.R. and Bannerman, R.T. 1999. Sources of Phosphorus in Stormwater and Street Dirt from Two Residential Basins in Madison, Wisconsin, 1994-1995. U.S. Geological Survey Water Resources Investigations Report 99-4021. Water Survey of Canada. 2003.. Environment Canada. Flow data available on the HYDAT disk. Winnipeg, MB. Williamson, D. A. 1988. Rationale document supporting revisions to Manitoba surface water quality objectives. Water Standards and Studies Section, Manitoba Department of Environment and Workplace Safety and Health Report No. 88-5: 50pp. Williamson, D. A. 2003. Manitoba’s nutrient management strategy. A presentation to the Manitoba Clean Environment Commission Public Hearing on the City of Winnipeg Wastewater Collection and Treatment Systems. Manitoba Water Quality Section, Manitoba Conservation. Winter J.G. and Duthie, H. C. 2000. Export Coefficient Modeling to Assess Phosphorus Loading in an Urban Watershed. J. Amer. Water Resources Association 36 (5):1053-1061 Wisconsin Department of Natural Resources. 1991. in C. D. Johnson, D. Juengst, 1997, Polluted Urban Runoff – A Source of Concern, University of Wisconsin-Extension, Wisconsin, Nonpoint Source Water Pollution Abatement Program, 4pp.

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Reducing Phosphorus Contribution from Animal Agriculture in Manitoba Dr. Kathy Buckley and Mr. Grant Penn

Brandon Research Centre, Agriculture and Agri-Food Canada

1.11 Chapter Summary The growth and development of modern commercial livestock production facilities are being restricted in some countries and will be restricted in other areas of the globe if solutions to the problem of nutrient management are not developed and implemented. It is reasonable to expect that nutrients removed in feed production should be returned to the soil to maintain soil quality and fertility. However, because all of the feed is not produced in close proximity to livestock production and land area surrounding these operations is often insufficient or unsuitable for continued heavy application of nutrients, the potential environmental impact of nutrient contamination is perceived as a major issue. The options in this case are either to reduce the size of livestock production units and distribute them according to land capacity, which is an unlikely measure, or, to reduce excreted nutrients and process the manure where necessary to redistribute the nutrients regionally. Because of its implication in water quality, P in livestock manure has been the focus of much research. As a result of this work a number of strategies for substantially reducing P excretion and management have been identified.

1. Improvement in feed efficiency. Increasing overall feed efficiency (amount of feed per unit of gain) can result in a major reduction in excreted nutrients. Improvements in genetic potential of animals, proper formulation and mixing of ingredients, feed processing methods (grinding, pelleting and particle size) and use of wet feeders can all improve feed efficiency. Inclusion of nutrients in the proper proportion can be critical to efficient use of feed ingredients. For example if vitamin D3 is not present in the diet in sufficient quantities, poultry are unable to use calcium and P in the diet. Considering that chicks consume only a small amount of feed during the first few days, thorough mixing is an important and critical consideration when mixing rations especially for poultry.

2. Improvement in P availability. Phytase addition to swine and poultry diets is especially effective in increasing P availability to meet nutrient requirements and reduce P excretion. The effect of phytase feeding on P solubility in the feces and elimination of excess plasma P via the urine needs further elucidation. The use of “highly-available P” corn, soybean and barley varieties in feed formulation is showing promise in terms of improving P availability to monogastrics. Wide-spread incorporation of these feeds in animal diets depends on improving the agronomics of crop production. Again, feeding trials with these feedstuffs is necessary to determine the solubility of excreted P. The combined effectiveness of phytase and highly available P feedstuffs to reduce P excretion and fulfill requirements for growth also needs to be determined. Breed selection for the natural ability to utilize phytate as well as using genetic engineering techniques to enhance the ability of the animal or the plant material to synthesize endogenous phytases show some potential for reducing P excretion. To advance feed research, methods and instrumentation for rapid or on-farm analysis of P availability of ingredients and total diets needs to be developed.

3. Reduction in over-formulation. Traditionally, the main consideration of diet formulation was to maximize the growth and health of the animal. Diet formulation then

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progressed to formulation of least-cost rations that met the requirement of the animal at the lowest cost. These diets offered better nutrient balance but still included excess macro-nutrients for “insurance”. The excess nutrients were deemed necessary in order to compensate for the variability in nutrient composition of ingredients (normally based on national averages) and to make up for the lack of knowledge concerning the availability of the nutrients in the ingredients. The use of more precise ingredient composition and nutrient availability data and better defined nutrient requirements will reduce P excretion. Reducing the P intake of feedlot cattle posses a particular problem because P content in high-concentrate diets contain P in excess of metabolic and growth requirements. The current best management practice is to remove all supplemental P but this requires that alternative sources of calcium supplement be included in the ration. In livestock production systems where diet P formulation is difficult, a manure treatment system to recover and recycle P might be considered.

4. Phase feeding and feeding for optimum gain. Phase feeding is a way to more precisely meet the nutrient needs of growing livestock and poultry. Nutrient requirements change as animals grow and if the nutrient formulation changes in tandem with growth, optimum gain with high nutrient use efficiency can be achieved. and high P in concentrate. Diets should be formulated so that animals perform at less than maximum, because relative nutrient response decreases as the animal reaches maximum performance.

5. Reduction of feed waste. Reducing feed loss is a simple solution to improve feed efficiency that is sometimes difficult to implement depending on the rearing system. Studies have shown that feed waste accounts for 3 to 8% of the feed nutrients. For example, a 5% level of feed waste can result in an equivalent loss of 82 g of P per hog. Use of proper feeder designs, regular maintenance and careful adjustments of feeders is essential for prevent of excessive feed waste.

6. Treatment systems to redistribute, recover or immobilize P. In cases where diet formulation as a method to reduce P excretion is more difficult to achieve, advanced treatment might offer the opportunity to manage P to increase the sustainability of an intensive livestock production system. Reducing P in liquid manure systems for hog production would not be so critical if it could be recovered and reused by advanced manure treatment systems. Solids separation and composting reduces mass and volume decreasing the cost in transportation and distribution of nutrients. Phosphorus recovery as a precipitate is possible in liquid and poultry manures. Immobilization of P by the addition of alum, magnesium, zeolites and polyacrylamide polymers also have potential as treatment options; however, the long-term effects of this form of nutrient management on soil biological, chemical and physical properties as well the effect on sustainability of crop production has yet to be explored. Biological or natural systems to recover P for redistribution or incorporation into animal feeds may become more feasible and economical in the future.

1.12 Introduction Most of the phosphorus (P) consumed by livestock is excreted in the manure. Any restrictions on the land application of phosphorus from animal manures means that producers must either transport the manure further from the point of production, reduce the amount of P excreted by the animals or treat the manure to reduce the amount or mobility of P applied to the soil.

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Figure 2.1. Phytic acid phosphate cycles in natural and agricultural ecosystems (from Brinch-Pedersen et al. 2002).

Phosphorus management requires an integrated approach which involves a number of steps. As a first step the supply of P in the feed should be in accordance with the animal’s requirements. Adequate knowledge of the P requirements and digestibility of feed P at every stage in the production cycle must be acquired and applied. Secondly, exogenous enzymes can be added to enhance the digestibility of P in feeds and decrease P excretion. A third step would be the use of a treatment system to concentrate the P in a form that is more transportable and economical to use as a P fertilizer. These options have been reviewed and assessed in light of the Manitoba context to identify promising P management strategies and technologies.

1.13 Phosphorus Excretion by Livestock and Poultry Phosphorus is an integral element in biological processes within plants and animals. It not only is important for bone and teeth development, but in the chemical structure of the nucleic acids in DNA, as the foundation for cell wall development, and it acts as the common energy currency in all biochemical processes in a living organism. This ubiquity in biological systems does not necessarily mean that P is a readily available nutrient. Most common forms of P in the environment are not very water soluble, like apatite in rock phosphate, making it more difficult for organisms to meet their P needs by absorption. In natural ecosystems, P is returned to the soil and converted to inorganic phosphate via biological and chemical processes after which it is available for a new cycle of plant growth. In agricultural ecosystems, new cycles of nutrient flow are added by tapping the phosphate cycle to redirect P for animal and human consumption and by reintroducing phosphate in the form of manure or industrially produced fertilizers and supplements (Figure 2.1).

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In livestock feedstuffs, 60-90% of the total P present is in the form of phytic acid (myo-inositol (1,2,3,4,5,6)-hexakisphosphate) which constitutes 1-4% of the total weight of cereal and oil seeds (El-Batal and Abdel Karem 2001). A substantial fraction of all P taken up by crop plants from soil is translocated ultimately to the seed and synthesized into phytic acid. Almost all phytic acid is present as phytin, a mixed salt (usually with K+, Ca2+, Mg2+ or Zn2+) that is deposited as globoid crystals in single membrane vesicles together with protein (Raboy 2001). Excretion of seed-derived dietary phytate can contribute to a major public health problem in the developing world, namely iron and zinc deficiency. In the developed world, the greatest interest in seed-derived feed phytate is its contribution to P in animal waste which can, in turn, contribute to water pollution (Sharpley et al. 1994). Inorganic forms of P are normally added to feed rations to increase the uptake of P. The addition of inorganic P does make more P available to the animal, but the proportion of P absorbed remains relatively constant, resulting in most of the inorganic supplemental P being excreted to the environment. Concerns about phytate utilization are usually confined to monogastric species as the microflora in the ruminant gut produce enzymes that make P available to the animal.

Rates of P excretion can only be estimated since the rate of excretions depends on a number of factors including the amount of P in the feed, the form of P present, the ability of the animal to take up the nutrient, the particular life stage of the animal and the animal products harvested (Table 2.1).

Table 2.1. Estimates of P excretion based on rations and products produced z

Numbers Below Expand from Daily Averages to Years

Numbers Below Based on Life Cycle Grow-Out

Herd or Flock Information Units Dairy Cows

Beef Steer Hens Broilers Turkeys Pigs

Animals/day or animals/grow-out No. 1 1 1000 1 1 1 Average DMIy kg/day kg 21.8 9.53 94.8 3.91 24.3 298.0 Average diet total P% (DM basis) % 0.5 0.4 0.65 0.8 0.8 0.57 Milk yield or egg yield /day kg 27.2 47.6 Milk yield or egg P% % 0.1 0.21 Av. net body weight gain or grow-out kg/day 0.09 1.41 0.839 2.18 10.8 115.2 Average P% of weight gain % 0.7 0.7 0.6 0.6 0.6 0.72 Average diet DM digestibility % 65 80 83 84 82 82 Ratio: Feed DM:(milk, 12 eggs, or gain) Ratio 0.8 6.76 3.16 1.79 2.25 2.59 Daily or grow-out balances for Px Input: g DMI x P/DMI = g 109 38 616 31 194 1699 Export: g milk or eggs x P% = g 27 100 g gain x P/gain = g 1 10 5 13 65 829 Manure DM output, % of input % 40 25 22 21 23 23 Yearly manure (wet) @ est. DM% kg 22734 5435 38062 4.1 28 428 P kg excreted yearly or /animal grow-out kg 30 10 187 0.018 0.13 0.869 Manure P% of DM (excreted) % 0.93 1.19 2.45 2.22 2.32 1.27 z Adapted from Powers and Van Horn 2001 y DMI = Dry matter intake x Expanded from daily averages above to annual or life cycle grow-out balances

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1.14 Phosphorus Compounds in Manure – Terminology and Analysis The collaborative efforts of poultry nutritionists, soil scientists, chemists, biologists, engineers and other researchers may be necessary to effectively minimize manure P losses in the environment. Different conventional methods for referring to specific P forms or chemicals are being used ultimately, because the groups do not have the same objectives. Terminology may differ among scientists for the same type of P. A poultry nutritionist uses the term dicalcium phosphate (“dical”), an inorganic P compound added to broiler, pig and beef diets to meet P nutrient needs. A common form of inorganic P to a soil scientist is the total reactive P that can be determined by a common extraction method. It is important to define some of the P terminology in an effort to link input P with P excreted by the different species of livestock (see section 2.9).

Total P (Pt) consists of inorganic P (Pi) and organic P (Po). Further, additional subcategories and numerous individual compounds are used to identify P forms (Miles and Sistani 2002). Manure or litter samples are subjected to drying and grinding prior to analysis. The sample can then be ashed and acidified before the Pi concentration is determined by inductively coupled plasma (ICP) spectrometry. The Pi can be solvent extracted and the amount of P can be measured colorimetrically by ICP. Soil scientists may refer to Pi as molybdate reactive P, soluble reactive P, total reactive P or soluble P. Litter P analysis is commonly performed for Pt and Pi in addition to water-soluble P (WSP) or Kelowna solution extractable P (van Lierop 1986). The WSP can contain some Po, but primarily contains Pi. Phosphorus forms are usually defined by the method of analysis, and the problem is specifically identifying the forms that are determined by the different methods (Pierzynski 2000). An accurate method for the determination of Pi is important because this is the form most susceptible to run off. Historically the Po portion has been calculated as the difference between Pt and Pi. Organic P forms are inositol phosphates, nucleic acids, phospholipids, trace quantities of sugar phosphates and phosphoproteins. The organic part of P in the manure may vary from 10 to 80% of the total and decreases as the manure ages, but not all organic compounds are converted to Pi at the same rate (Barnett 1994). Organically bound P is suspected to change readily during extraction analyses and is difficult to quantify (Greaves et al. 1999). To add to this complexity, the concentration of P in livestock waste varies with the diet, housing and animal management, waste handling procedure, amount and type of bedding, water and feed spillage and climate. In preparing manure for analysis, freeze drying should be avoided as it results in higher P, Ca, Mg, K, Fe, Mn and Zn losses compared to air or oven drying (Sistani et al. 2001).

To gain a better understanding of the true digestibility of phytate by livestock and poultry it would be useful to have an analytical technique for manure phytate content. Phytic acid determination in feeds is relatively easy. Most methods are based on the insolubility of ferric phytate and colorimetric determination of the iron or phosphate content of the precipitate, or on the decrease in iron in the supernatant. Chromatographic and capillary isotachophoresis techniques have also been commonly used to measure feed phytate (Xu et al. 1992; Blatny et al. 1995). Sequential P fractionation and enzymatic characterization of organic phosphorus in animal manures (He and Honeycutt 2001) and Near-Infrared Reflectance Spectroscopy (NIRS) to predict total and phytate P (Smith et al. 2001c) content of excreta show promise as analytical techniques. Alternatively Dušková et al. (2000) have described a capillary isotachophoresis method for analysis of phytic acid in feeds and feces of pigs and poultry. The technique involves injecting the sample as a large plug into the anodic end of a capillary column, and then the high voltage is applied. The components in the sample plug will separate out on the basis of their

46

relative electrophoretic mobilities as they travel along the column. As the sample plug exits the column, each band of ions eluting will cause a dramatic drop in current, which can be monitored and used to quantify, and to some extent identify, species on the basis of migration time and band (peak) width. Results of the application of this technique to the analysis of phytic acid in feces of pigs and poultry indicates that capillary isotachophoresis coupled with a phytic acid precipitation technique is precise, reproducible and much simpler than the common high performance liquid chromatographic technique (Dušková et al. 2000). Furthermore, they concluded that phytic acid contributes significantly to the total P in feces of pigs and hens and enzyme addition to a feed mixture for pigs to make phytic acid available is not sufficient to completely eliminate phytic acid from feces.

1.15 Feed Treatments, Additives and Novel Ingredients At lease five different strategies have been devised to improve the P bioavailability in animal feed and to reduce the environmental P load. 1) The addition of phytase to feed will significantly enhance the release of phosphate from phytate. The commercial potential of microbial phytases has stimulated a large body of research and development activities to identify microbial phytases with high thermostability and better catalytic properties. 2) An old feed treatment method is to incubate the feed in water (steeping), whereby the endogenous phytase potential is activated. A newer variation on this method is to ferment the grain with bacterial spores which produce phytase, then dry the grain and prepare the diet. 3) Mutation breeding for impaired phytic acid biosynthesis is proving to be useful in corn, rice and barley. 4) Recently, pigs have been engineered to produce heterologous phytase in their salivary glands. 5) Plants can be transformed for increased phytase production in the seeds. Schőner and Hoppe (2002) have recently published an excellent review of the effects of phytase on poultry nutrition and performance.

1.15.1 Addition of Phytase to Diets The bioavailability of P in cereal grains and products from oilseeds is generally very low for pigs and poultry, because they have limited capability to utilized phytate P. Bioavailability estimates of P in maize and soybean meal for pigs and poultry range from 10-30% (Nelson 1967; Calvert et al. 1978; Jongbloed and Kemme 1990). Phytate concentration in feedstuffs may vary, depending on the stage of maturity, degree of processing, cultivar, climatic factors, water availability, soil factors, location and the year of growth (Reddy et al. 1982; Ravindran et al. 1995). Phytate P must be hydrolyzed into inorganic P before it can be utilized by pigs and poultry. Phytase is a special kind of phosphatase that catalyzes the stepwise removal of inorganic orthophosphate from phytate. Two classes of phytases are recognized by the International Union of Pure and Applied Chemistry and the International Union of Biochemistry: 3-phytase (E.C.3.1.3.8) first removes the orthophosphate from the 3-position of phytic acid, whereas 6-phytase first removes the orthophosphate from the 6-position. Both classes of phytase then successively remove the remaining orthophosphates, which results in intermediates ranging from free myo-inositol and mono- to tetra-phosphates of inositol (Gibson and Ullah 1990). Phytases are known to occur widely in microorganisms, plants and certain animal tissues (Reddy et al. 1982). Contents of the stomach and intestine of pigs (Jongbloed 1997) and crop, stomach and small intestine of chickens (Liebert et al. 1993) have negligible phytase activity. Although phytase has been detected in other animal tissues and resident bacteria, the activity is thought to

47

be negligible for improving the availability of phytate P in nonruminant animals (Kornegay 1996).

Dietary addition of microbial phytase or the inclusion of high phytase ingredients in pig, poultry and fish diets is now well documented to release a large portion of the naturally occurring phytate P. Phytase activity can increase P digestibility, reduce the amount of P excreted in the feces of monogastic livestock and perhaps reduce the amount of inorganic P that must be added to the diet to meet dietary requirements (Beers and Jongbloed 1992, 1993; Adeloa et al. 1995, 1998; Carter et al. 1999 (see Figure 2.2); Kornegay 1996; Kornegay and Qian 1996; Han et al. 1997, 1998; Harper et al. 1997; Zimmermann et al. 2003; Li et al. 1998; Rice et al. 2002; Sands et al. 2001; Hill et al. 2002; Traylor et al. 2001; Valencia and Chavez 2002; Leske and Coon 1999; Hatten et al. 2001; Jalal and Scheideler 2001; Żyła et al. 2001; Yan et al. 2001; Lan et al. 2002; Rama Rao et al. 1999; Li et al. 2001; Edens et al. 2002; Um and Paik 1999). Research results displayed in Table 2.2 indicates that while there is a net benefits to feeding phytase, reduction of P excretion in the feces is variable. In addition to this obvious benefit it is becoming clear that the use of adequate amounts of phytase in most pig and poultry diets results in improved availability of Ca, Zn, protein/amino acids and energy (Kornegay 2001). Phytate is known to complex with other nutrients reducing protein and mineral absorption, and increasing excretion of essential nutrients. In addition to the reduction in P excretion as a result of feeding phytase, Smith et al. (2001a) found that phytase feeding reduced manure (total feces and urine) soluble reactive P concentrations in swine manure by 22%. It is not clear from the scientific literature how phytase affects plasma levels of P and subsequent excretion of P in the urine. Furthermore, information is not available on the influence of phytase on P metabolism of swine fed diets of varying P digestibility.

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Table 2.2. Effect of phytase on P digestibility, retention and excretion in various species and classes of livestock.

P Digestibility % a P Retention % P Excretion %

Species Type

Source/ Amount of Phytase

Feed P Content %

% Added Dietary Pi & form

Non-Phytate Control

Phytate Treatment

Non-Phytate Control

Phytate Treatment

Non-Phytate Control

Phytate Treatment Reference

600 PU/kgc 0.38 38.65 55.10 27.65 45.18 72.56 54.72 Sands et al. (2001)

A. niger 1500 PU/kgc

0.50 con 0.52 tmt

0.14 diCaP

56.98 68.64 56.79 63.29 Adeola et al. (1995)

498 PU/kgg 0.387 33.8 44.6 32.7 43.5 67.43 56.47 Hill et al. (2002)

1000 PTU/kgh 0.47 con 0.48 tmt

0.82 diCaP

41.7 60.2 58.72 39.75 Valencia & Chavez (2002)

A. niger 1269 PU/kgc

0.34

52.2 66.7 52.1 66.1 47.83 33.86

Wheat bran 627 PU/kgc

0.34 con 0.35 tmt

52.2 59.5 52.1 59.4 47.83 40.54

Han et al. (1997)

A. oryzae 750 U/kgi

0.38 37.9 53.2 37.2 56.0 67.0 47.4

A. oryzae 750 U/kgi

0.56 con 0.38 tmt

1.06 diCaP

56.3 53.2 52.9 56.0 46.2 47.4

Li et al. (1998)

A. niger 454 U/kgk

0.405 59.59 72.24 Rice et al. (2002)

A. niger 75 pU/kgd + wheat 75 pU/kgd

0.363 con 0.377 tmt

29.0 46.9 28.6 46.5 71.0 53.1

Swine Feeders

A. niger 75 pU/kgd + rye 75 pU/kgd

0.36 con 0.348 tmt

28.0 43.0 27.6 42.4 72.0 57.0

Zimmerman et al. (2003)

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Species Type


Feed P Content %

% Added Dietary Pi & form Non-Phytate

Control Phytate

Treatment Non-Phytate

Control Phytate

Treatment Non-Phytate

Control Phytate

Treatment Reference Wheat 75 pU/kgd + rye 75 pU/kgd

0.374 32.4 38.2 32.0 37.7 67.6 61.8 Zimmerman et al. (2003)

A. niger 500 U/kgk

0.40 1.0 diCaP 54.0 68.5

A. niger 1000 U/kgk

0.40 1.0 diCaP 54.0 71.1

Feeders

A. niger 1500 U/kgk

0.40 1.0 diCaP 54.0 74.0

Traylor et al. (2001)

Grower A. niger 500 U/kgk

0.43 0.16 diCaP 45.9 56.7 43.29b

Swine

Finisher A. niger 500 U/kgk

0.37 30.1 44.2 55.8b

Harper et al. (1997)

Broiler Starter

A. niger 10478 PTU/kgh

0.72 0.7 diCaP 100 % of basal diet

91.0 % of basal diet

Broiler Grower


0.72 0.8 diCaP 100 % of basal diet

97.8% of basal diet

Broiler Finisher


0.41 100 % of basal diet

101.0% of basal diet

Edens et al. (2002)

A. niger 600 FTU/kge

0.215 27.0 58.0 Broiler


0.177 34.8 40.9


0.273 36.8 53.4

Poultry

Layers


0.183 28.6 44.7

Leske and Coon (1999)

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Species Type


Feed P Content %

% Added Dietary Pi & form

Non-Phytate Control

Phytate Treatment

Non-Phytate Control

Phytate Treatment

Non-Phytate Control

Phytate Treatment Reference

A. niger 500 U/kgk

0.70 con 0.69 tmt

1.4 TriCaP 31.8 48.0 77.5 68.2 Um & Paik (1999)

Layers A. niger 300 fU/kgf

0.34 0.10 DiCaP 32.9 28.3 Jalal and Scheideler (2001)

Broiler 750 FTU/kg m + 3156 AcPU/kgm

0.71 con 0.41 tmt

1.6 DiCaP con

86.5 91.5 Zyla et al. (2001)

Broiler 1-21 d

M.Jalaludinii 1000 FTU/kge

0.477 53.1 73.1 46.9 26.9

Poultry

Broiler 22-42 d

M. jalaludinii 1000 FTU/kge

0.471 48.7 67.2 51.3 32.9

Lan et al. (2002)

Con = Control diet tmt = treatment diet a Calculations were performed on the some of the published data as necessary to express the units of digestibility, retention and excretion as %. Digestibility is estimated from absorption data if required. b Results combined two treatment effects c One Phytase Unit (PU) is calculated as the amount of enzyme that liberates 1 µmol of inorganic phosphorus per minute from sodium phytate at pH 5.5 and 37oC. (Sands et al.

2001; Han et al. 1997) d One unit of phytase activity (pU) is defined as the amount of enzyme that releases 1 µmol of inorganic ortho-phosphate from 32.6 mmol/L sodium phytate per minute at pH 5.5

and 37oC. (Zimmerman et al. 2003) e One Phytase Unit (FTU) is defined as the quantity of enzyme that liberates 1 µmol of inorganic phosphorus per minute from 0.0051 mol/L sodium phytate of pH 5.5 and 37oC.

(Leske and Coon 1999; Lan et al. 2002) f One phytase unit (fU) is the quantity of enzyme that releases 1 µmol of inorganic phosphorus per min from 0.15 mmol sodium phytate/L at pH 5.5 and 37oC. (Jalal and Scheideler

2001). g One Phytase Unit (PU) was not defined (Hill et al. 2002). h One Phytase Unit (PTU) was not defined (Valencia and Chavex 2002; Edens et al. 2002). i One Phytase Unit (U) is defined as the quantity of enzyme that liberated 1 µmol of inorganic phosphorus per minute for 5.1 mmol of sodium phytate at 37°C and Ph 5.5 (Li et al.

1998). k One Phytase Unit (U) was not defined (Rice et al. 2002; Traylor et al. 2001; Harper et al. 1997, Um and Paik 1999). m One Phytase Unit (FTU) is defined as the quantity of enzyme that liberates 1µmol of inorganic phosphorus from 2 mmol sodium phytate in 1 minute at 40°C and pH 4.5. One

unit of acid phosphatase activity (AcPU) was equal to 1 µmol per minute of p-nitrophenol liberated from 5.5 mmol diisodium p-nitrophenylphosphate at 40 C, pH 4.5.

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Treatment of the diets or dietary ingredients with phytase following processing, but just prior to feeding, may be beneficial in terms of phytate reduction in the gastro-intestinal tract of the animal. Postpelleting application of dry phytase to broiler diets results in significant reductions in litter P accumulation while providing the safety factor associated with the bactericidal effect of high pelleting temperature (Edens et al. 2002). The dry phytase has the added advantage of being more stable and more economical than liquid phytase.

Figure 2.2. Effect of phytase on P balance of finishing pigs (from Carter et al. 1999).

1.15.2 Pre-Treatment of Feed Heat treatment of feed was adopted as a standard practice when feed-borne Samonella became prevalent enough to constitute a serious health risk. Steam pelleting reduces dust problems in the barns and prevents fractionation of the feed ingredients during handling. Exposure to high temperatures during feed processing has a negative effect on P digestibility since a substantial amount of endogenous phytases are denatured. Skoglund et al. (1997) found that steeping a mixed pig diet for 9 h at room temperature rather than steam pelleting the diet reduced the phytate content by 45% and increased phosphorus absorption in the stomach and small intestine of pigs. In earlier work, Kemme and Jongbloed (1993) found that soaking of a phytase-supplemented pig diet in water at room temperature for 8-15 days increased P digestibility to a greater extent than soaking the feed without added phytase. A similar observation was made by Näsi et al. (1995) who found that microbial phytase and soaking with whey improved the utilization of phytate P in barley-rapeseed meal diet and reduced the amount of P excreted in pig feces. A novel investigation by Valaja et al. (1998) revealed that microbial phytase supplementation in a pig diet containing wet barley fibre by-product from the starch-ethanol process did not improve phosphorus digestibility. This was due to the already highly degraded state of the phytate in the wet barley protein fiber. More recently Carlson and Poulsen (2003) performed fermentation studies to examine the effect of grain heat treatment, fermentation temperature, fermentation time and addition of microbial phytase on phytate degradation. They concluded that endogenous phytases are sensitive to heat during fermentation and maintain their activity better at 10°C rather than 20°C while the microbial phytases are unaffected by fermentation at the higher temperature. Since phytate breaks down faster at higher temperatures,

52

reducing the fermentation time, there appears to be an advantage to adding microbial phytase to the fermentation mixture (Carlson and Poulsen 2003). High temperature soaking of wheat, rye, barley and oats, even in the absence of exogenous sources of phytase, has been shown to significantly reduce grain phytate content (Fredlund et al. 1997). Endogenous phytases were destroyed by heat treatment of the grain indicating that soaking alone was beneficial. An investigation by Matsui et al. (1996) demonstrated that wet fermentation of soybean meal with microbial spores of Aspergillus usami prior to drying and pelleting improved phosphorus availability in chicks. They concluded that dietary supplementation of inorganic P was not necessary for chicks fed fermented soybean meal. Dephytinization of soy-based fish feed by fermenting soy proteins with phytase prior to pelleting shows promise as a method to lower the P load into water from farmed fish (Vielma et al. 2002). With the development of new liquid feeders it has become more common to feed soaked and fermented feed to pigs (Carlson and Poulsen 2003), although the persistent but relatively low risk of Salmonella typhimurium DT104:30 contamination of fermented feed still remains (Beal et al. 2002). Further discussion on feed processing will appear in later sections in the discussion on management of individual livestock species.

1.15.3 Low Phytate Feed Ingredients Genetically enhanced, low-phytate corn (more commonly known as high available P corn – HAPC) is now commercially available. This type of corn, which contains the lpa1 mutant gene that inhibits phytate synthesis, has essentially the same amount of total P, but less than one-half as much of the P in the form of phytate. As a result, the inorganic P is several times greater in low-phytate corn (0.18%) than in normal corn (0.05%) (Spencer et al. 2000). Similarly, investigations of the results of feeding modified corn to pigs indicated that low-phytate corn contains at lease five times as much available P as normal corn. The use of low-phytate corn can greatly reduce the amount of P excreted by the pig and increase the N:P ratio in the manure (Pierce and Cromwell 1999; Spencer et al. 2000). In both of these studies, phosphorus excretion was lowered 35-40% by feeding low-phytate corn. These observations were confirmed by Veum et al. (2001) who found that feeding low-phytate corn in pig diets reduced P excretion by 50% and 18.4% in semipurified and practical diets, respectively, compared to normal corn. When phytase is added to the low-phytate corn pig diets, the effect appears to be additive and P excretion can be further reduced (Sands et al. 2001; Hill et al. 2002). Phosphorus retention in broiler chickens was also improved by feeding low-phytate corn compared to normal corn (Li et al. 2000). An in vitro digestion study performed by Li et al. (2000) also demonstrated that 65% (1420 mg/kg) of the total phosphorus in low-phytate corn was released compared to 23% (543 mg/kg) from normal corn. The major conclusions of similar research conducted by Saylor et al. (2001) were: using phytase and reducing nonphytate P by 0.1% using low-phytate corn will have no negative effect on broiler performance, use of low-phytate corn alone reduced inorganic P additions by 16-19% and combining phytase with low-phytate corn resulted in a reduction of 52-63% in dietary inorganic P. Dietary modification resulted in significant reductions in broiler litter P (total P, water soluble P and the ratio of water soluble P:total P), and finally, broiler litters produced with modified diets caused smaller increases in soil test P and water soluble P than litters from normal corn diets but a greater increase in post-harvest soil test P and water soluble P than inorganic fertilizer P.

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Figure 2.3. Metabolic processes of myo-inositol, the precursor of phytic acid (from Brinch-Pedersen et al. 2002).

Unfortunately the lpa1 mutation appears to have an impact on vegetative processes as well as seed phosphorus chemistry because the pathways of inositol to phytic acid are active in most tissues of the whole plant (Raboy 2001) as shown in Figure 2.3. The block in phytic acid metabolism during corn seed development appears to reduce seed dry weight, which leads directly to lower yields (Raboy et al. 2000).

Using a conventional mutagenic technique employing ethyl methansulfonate, Wilcox et al. (2000) produced a soybean mutation showing reduced phytic acid P and increased seed inorganic P. They predict that the available P in these low phytic acid soybean mutants would be about 75% of seed total P compared to nonmutant soybean seed which has an approximate P availability of 25% of seed total P.

Low-phytate barleys have recently been developed by USDA scientists (Larson et al. 1998; Raboy 2001) from chemically mutagenized Harrington barley. Grain from seed stocks of low phytate genotypes, LP422 (hulled, 68% of normal phytate), LP 635 (hulled, 41% of normal phytate), LP955 (hulled, 3% of normal phytate) and LP 422H (hulless, 54% of normal phytate) were compared with normal varieties of hulled and hulless barley in feeding trials with finishing pigs (Thacker et al. 2003, in press). Selection for reduced phytate content did not appear to appreciably alter the chemical composition of barley and had no significant effect on apparent fecal digestibility of dry matter, crude protein or grass energy. Phosphorus digestibility increased as the level of phytate in the barley declined (Table 2.3). In this study, barley was the only dietary component so production data is still necessary before the real value of feeding low-phytate feeds on total phosphorus excretion can be assessed.

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The use of low-phytate barley in pig diets reduced P excretion in swine waste by 55% and 16% in semipurified and practical diets respectively, compared with feeding normal barley (Veum et al. 2002). Low-phytate barley was found to be nutritionally equal to normal barley supplemented with inorganic P to equal the estimated available P in low-phytate barley. Using in vitro procedures designed to mimic the digestive system of the pig, Veum et al. (2002) found that the availability of P for pigs was estimated to be 52% for low-phytate barley and 32% for normal barley. Studies to determine the effect of a Western Canadian low-phytate barley fed with added phytase on P digestibility in pigs are currently in progress (B. Rossnagel – private communication).

Table 2.3. Phytate levels and digestibility coefficients for P in low-phytate varieties of barley*

Barley Variety NBa LP 422 68% of NB

LP 635 41% of NB

LP 955 3% of NB HNBb

LP422H 54% of NB

Phytate content (g/kg) 2.2 1.49 0.92 0.07 2.86 1.54 P digestibility (%) 43.13 49.74 54.81 67.53 30.49 46.62 a Normal barley b Hulless normal barley * adapted from Thacker et al. (2003, in press)

1.15.4 Augmentation of Endogenous Phytase in Livestock The feasibility of transgenic augmentation of salivary phytase production in pigs is currently being studied. Golovan et al. (2001) has recently described the development of transgenic pigs producing salivary phytase as a result of the insertion of the parotid secretary protein promoter linked to the E. Coli appA phytase gene. Saliva samples from the 33 transgenic founder (G0) piglets indicated that 14 produced 5 to 6000 U/mL of phytase in the saliva at 7-11 days of age, 15 produced less than 5 U/mL and four lacked detectable salivary phytase activity. Transgenic G1 progeny have been obtained from 13 founder lines. Of the six litters sired by one of the founder boars, 25 of the 53 piglets were transgenic and showed phytase activities ranging from 341 U/mL to greater than 10 077 U/mL, with a median of approximately 2000-3000 U/mL, indicating a potential for phytate digestion. Nutritional trials with soybean as the phytate source indicated that true digestibility of P in the diets approached 100% for both weanling and growing-finishing transgenic pigs whereas the digestibility for non-transgenic pigs was approximately 50%. The P content of fecal matter from transgenic weanling and growing-finishing pigs was reduced by as much at 75% and 56%, respectively, compared with that of non-transgenic counterparts (Golovan et al. 2001). Phosphorus levels in the urine of transgenic pigs were not described in this study.

It has been suggested that there is a genetic component for bio-availability of phytate P in chickens. The inheritance of the trait for increased bio-availability of phytate P was determined in a randomly mating population by Aggrey et al. (2002). The results of this study suggest that phytate P utilization in chickens is heritable and genetic selection for the trait is possible. Sires excreting low phytate P produced progeny that were significantly heavier at 16 days of age and had better bone mineralization than progeny sired by males expressing poor phytate utilization.

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1.15.5 Enhancement of Phytase Activity in Feed Phytase activity has been reported in a wide range of seeds, such as rice, wheat, barley, corn, rye, soybean and other leguminous and oil seeds (Reddy et al. 1982; Gibson and Ullah 1990) although phytase activity of seeds varies greatly among species of plants (Table 2.4). With the exception of wheat, rye and triticale, most dormant seeds contain very low phytase activity. Diets formulated using ingredients with high phytase activity, such as wheat bran, wheat, triticale, rye bran and wheat middling, may promote greater absorption of phytate P but may not be economically or practically feasible (Kornegay 1996). Table 2.4. Phytate phosphorus content and phytase activity of some common feed ingredientsa Ingredient

Phytate P (g/100g DM)

Phytate P (% of total P)

Phytase activityb (units/kg)

Cereals and by products Corn 0.24 (0.17-0.29) 73 (61-85) 15 Wheat 0.27 (0.17-0.38) 68 (61-78) 1190 Barley 0.27 (0.19-0.33) 58 (55-62) 580 Oats 0.29 (0.22-0.35) 69 (48-78) 40 Millet 0.19 70 Wheat bran 8.8 (6.0-12.7) 76 (68-93) 2960 Grain legumes Lupins 3.0 (2.9-3.0) 55 0 Peas 0.17 45 115 Chick peas 0.21 51 - Oilseed meals Soybean meal 0.39 60 8 Canola meal 0.70 59 16 Sunflower meal 0.89 77 60 a Data adapted from Ravindran et al. (1995) and Kornegay (2001) b One unit is defined as that amount of phytase which liberated inorganic phosphorus from a 5.1 mM Na-phytate solution at a rate of 1 µmol/min at pH 5.5 and 37°C. Research results have documented that Aspergillus niger phytase can be synthesized efficiently in transgenic plant cells (Verwoerd et al. 1995). Since 1995, transgenic wheat (genetically modified with the Apergillus niger phytase encoding gene phyA) and transgenic canola (Phytaseed – canola seed genetically modified with Aspergillus ficuum phytase gene, BASF Corp., Offenbach, Germany) have been produced (Brinch-Pedersen et al. 2000; Zhang et al. 2000). It is believed that compared with microbial phytase, a genetically modified plant has some advantages that may lower the cost of phytase: 1) the foreign genes can be easily transferred and expressed in plants, 2) plants use solar energy and have large biomass accumulation and 3) the phytase in the plant has no contamination with animal pathogens. Although it is early in the development of phytase-containing wheat, there is some indication that only limited amounts of

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transgenic wheat are required in a compound feed to ensure proper degradation of phytate (Brinch-Pedersen et al. 2000).

1.16 Feeding and Management of Poultry and Livestock to Reduce P Excretion

1.16.1 Feeding and Management of Poultry

1.16.1.1 Phosphorus requirements The most important phosphorus requirements are those associated with the production of meat. In adult birds phosphorus balance is achieved at two principal levels, being the bone and the kidney. Both calcium and phosphorus are mobilised from the medullary bone through the action of parathyroid hormone (PTH). The action of PTH is through 1, 25-dihydrocholecalciferol (1,25(OH)2D3), an active derivative of vitamin D (Larbier and Leclercq 1994). The net result of PTH activity is an increase in phosphate secretion, which is associated with the presence of phosphates in the urine. Phosphorus secretion from the kidney itself is not constant as variable levels of reabsorption may be observed. Bone mobilization and urinary excretion are susceptible to variation as a result of dietary factors including intestinal absorption, level of available phosphorus and form of vitamin D3. Generally phosphorus requirements necessary for maximum growth rate are lower than those needed for maximum bone mineralization. Phosphorus deficiency is associated with loss of appetite, slower growth rate, serious problems of locomotion and death. In growing birds, symptoms are more evident the younger they are. Requirements are determined by taking into account growth rate and levels within bone ash. In poultry, ash levels within the tibio-tarsal bones are used to establish the requirement. Often bone (tibiae) strength is used as a measure of phosphorus sufficiency (Figure 2.4).

Figure 2.4. Bone (tibia) breaking strength determination (from Angel 1999).

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1.16.1.2 Feeding and management strategies In recent years the emphasis has focused on feeding phytase to reduce P excretion in all classes of poultry, broilers and laying birds. There are several other strategies that may be used to decrease P excretion. The first, is to feed birds closer to P and calcium (Ca) requirements and minimize safety margins that traditionally have been high for P in order to avoid leg weakness and preserve egg shell quality. This strategy has greater feasibility if feed mills have the capability to rapidly analyze organic and total P content and formulate the ration based on the availability of P in ingredients. Near-Infrared Reflectance Spectroscopy shows promise as a tool for real-time analysis of feed ingredients. Second, feed birds according to declining requirements as the birds age. Most research on P requirements has focused on 3 phases based on age, similar to those cited by NRC (1994). Increasing the number of age phases will help decrease P in litter. Third, feed birds according to their genetic potential (gender and strain) and finally, use feed additives alone or in combination with phytase and low phytate feeds. Although the feeding of phytase and low phytate feeds significantly reduces P excretion, there is some indication that this strategy may not significantly reduce soluble reactive P concentrations in litter runoff (Moore et al. 1998).

Studies were conducted by Edwards (2002) to determine the effect of dietary supplementation with the analogues or precursors of Vitamin D namely, cholecalciferol (D3), 1,25-dihydroxycholecalciferol [1,25-(OH)2D3], 1α-hydroxycholecalciferol (1α-OHD3), and 25-hydroxycholecalciferol (25-OHD3) on utilization of phytate P by broiler chickens. High levels of D3 (110 µg and 220 µg/kg of diet) increased phytate P utilization, but the increase was not as great as that obtained from 1,25-(OH)2D3 supplementation. Supplements containing 1,25-(OH)2D3, D3 and 1α-OHD3 were consistently effective in increasing phytate P utilization as measured by plasma Ca and P, incidence of P rickets, bone ash, and retention of Ca, P, and phytate P. Supplementation with 25-OHD3 in general gave smaller and more inconsistent responses to these criteria, indicating some inconsistency in its ability to improve phytate P utilization.

As described in previous sections, the availability of phytate-P can be improved by dietary phytase supplementation. However, the efficiency of dietary phosphorus utilization will be reduced at a suboptimal Ca/aP (available P) ratio. Phytate P degradation in the gastro-intestinal tract as well as P absorption from the small intestine will be reduced at high levels of calcium, while on the other hand, absorbed P will be excreted in the urine if dietary calcium levels are too low (van der Klis and Versteegh 1999). Figure 2.5 illustrates that at low dietary calcium levels, P absorption from the small intestine was maximal, but due to the lack of a proper counter ion it was retained with the lowest efficiency. At increasing dietary calcium levels, P absorption was reduced and the efficiency of P retention improved. The optimum Ca/aP ratio will be affected by the amount of phytate in the diet.

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Citric acid at levels of 4 to 6% of the diet was found to be markedly efficacious for improving the utilization of phytate-phosphorus in broiler chicks although these levels of citric acid would not be currently economical for poultry (Boling et al. 2000). Further research with broilers chickens indicates that citric acid increases P utilization in corn-soybean meal diets and reduces the available P requirement by approximately 0.10% of the diet (Boling-Frankenbach et al. 2001).

Several trials have been done to determine more accurately the phosphorus needs (on a nonphytate phosphorus basis) of broilers grown in a four phase feeding system (Angel 1999). The four phases were: starter (hatch to 18 days of age), grower (18 to 32 days of age), finisher (32 to 42 days of age), and withdrawal (42 to 49 days of age). This research has shown that available phosphorus in the diet can be reduced by 5% in the grower diet and by 15% in the finisher diet without affecting bone strength or performance. Withdrawal phase requirements are currently being determined. Having more accurate phosphorus requirement (nonphytate basis) information may see a reduction of at least 10% in the amount of available phosphorus fed to broilers which would mean about a 10% decrease in litter phosphorus (Angel 1999). A faster response to deficient nonphytate P levels in older birds seems to indicate that phytate P utilizing ability of laying hens seems to decline with age (Sohail and Roland 2002). This indicates that it may be more difficult to optimize inorganic P additions to reduce manure P during the entire lifecycle of the laying bird.

Figure 2.5. The absorption and retention of phosphorus in four week old broilers, fed a diet containing 2.5 g available P/kg feed and 2.0 g phytate P/kg feed at variable Ca levels (from van der Klis and Versteegh 1999).

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Feed manufacturing protocols, including improved feed formulation, grinding methods to provide proper particle sizes and feed uniformity is important to maximize nutrient use efficiency. Feed processing methods, such as expanding and pelleting techniques, need to be studied to improve feed efficiency and reduce nutrient excretion (Nahm 2002). Proper forming of feed into crumbles and pellets will reduce wasteage substantially. A 2% reduction in feed wastage can reduce the N and P in manure approximately 3% (van Heugten and van Kempen 1999).

1.16.1.3 Summary of strategies for P decreases Based on the literature there are a number of suggestions for potential decreases in P excretion from poultry production

• Flock management o Use superior stock that grows fast and converts the dietary nutrients efficiently. o Separate sex feeding. o Use a phase feeding approach.

• Improving P digestibility in feedstuffs o Formulate diets as close as possible to bird requirements. o Use supplemental enzymes, particularly phytase, in the feed to increase digestion

of nutrients and thus reduce their content in the manure. o Use highly digestible feed ingredients. o Assure uniform feed mixing.

• Use formed feeds to avoid feed wastage.

1.16.2 Feeding and Management of Swine

1.16.2.1 Phosphorus requirements The requirements for P, which depend on an adequate supply of Ca, have been summarized by NRC (1998) in Table 2.5. The suggested Ca:P ratio for swine diets is between 1:1 and 1.25:1. These requirements for total Ca and total P are based on a fortified, corn-soybean meal diet and take into account the fact that available phosphorus from corn is relatively low. The recommendations do not reflect the increased availability of P in other feedstuffs more commonly used in swine diets in Western Canada. This higher availability is attributed to the presence of naturally occurring phytase in plant sources (Table 2.6). On a whole farm basis, the efficiency with which phosphorus is retained in pigs is typically low (de Lange et al. 1999) as illustrated in Table 2.7. It is well known that P deficiency results in reduced growth and abnormal bone mineralization. It has been a common practice to include a plentiful supply of P into pig diets to ensure that P deficiency would not occur. The concern over the long-term accumulation of P in soil from pig manure has fostered much recent and ongoing research on P nutrition of pigs in attempt to improve P utilization without affecting pig health and productivity.

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Table 2.5. Daily calcium and phosphorus requirements of growing pigs allowed feed ad libitum (90% dry matter)a Body weight (kg) b 3–5 5–10 10–20 20–50 50–80 80–120 Average weight in range (kg) 4 7.5 15 35 65 100

Requirements (%/kg of diet) Calcium (%)c 0.90 0.80 0.70 0.60 0.50 0.45 Phosphorus, total (%)c 0.70 0.65 0.60 0.50 0.45 0.40 Phosphorus, available (%)c 0.55 0.40 0.32 0.23 0.19 0.15

Requirements (amount/day) Calcium (g)c 2.25 4.00 7.00 11.13 12.88 13.84 Phosphorus, total (g)c 1.75 3.25 6.00 9.28 11.59 12.30 Phosphorus, available (g)c 1.38 2.00 3.20 4.27 4.89 4.61 a from NRC 1988 bPigs of mixed gender (1:1 ratio of barrows to gilts). The requirements of certain minerals and vitamins may be slightly higher for pigs having high lean growth rates (>325 g/day of carcass fat-free lean), but no distinction is made. cThe percentages of calcium, phosphorus, and available phosphorus should be increased by 0.05 to 0.1 percentage points for developing boars and replacement gilts from 50 to 120 kg body weight.

Table 2.6. Bioavailability of P for pigs a Feedstuff

Bioavailability of P for pigsb %

Bioavailability of P for pigsc %

Cereal grains Corn 12 14 Distillers corn grain with solubles 77 Corn, high moisture 53 Oats 23 22 Barley 31 30 Triticale 46 46 Wheat 50 50 High protein meals – plant origin Canola meal 21 21 Soybean meal, dehulled 25 23 Soybean meal, 44% protein 35 31 Alfalfa meal (dehydrated, 17%CP) 100 Fish meal (Menhaden, mech. extr. 94 Whey 97 Milk, dried skim 91 a Relative to the availability of P in monosodium phosphate, which is given a value of 100. b adapted from Kornegay (1996) c NRC 1998

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Table 2.7. Typical mineral balances (kg/animal) on Dutch pig farms (adapted from de Lange et al. 1999) Nitrogen Phosphorus Potassium I. Growing pigs (25-106 kg live weight) Dietary levels (%) 16.7z 0.52 1.22 Intake (kg/pig) 6.36 1.23 2.90 Excretion (kg/pig) 4.48 0.83 2.73 Retention (kg/pig) 1.88 0.40 0.17 Recovery (%) 29.5 32.5 6.0 II. Sows, including nursing piglets Dietary levels (%) 15.7z 0.59 1.32 Intake (kg/pig) 27.57 6.53 14.52 Excretion (kg/pig) 22.50 5.5 14.0 Retention (kg/pig) 5.07 1.03 0.52 Recovery (%) 18.4 15.8 3.6 III. Starter pigs (9-25 kg live weight) Dietary levels (%) 18.4 z 0.67 1.25 Intake (kg/pig) 0.94 0.21 0.40 Excretion (kg/pig) 0.56 0.13 0.36 Retention (kg/pig) 0.38 0.08 0.04 Recovery (%) 40.5 39.4 10.0 z Total protein (N x 6.25) rather than N

To reduce mineral excretion, the P digestibility and requirements for each species, production type and stage of development must be known. Current P recommendations for swine vary widely in different countries. The reason for this is likely due to differences in housing, genotype, level of feeding, major ingredients and energy content of the diets and criteria of adequacy (Beers and Jongbloed 1993).

1.16.2.2 Feeding and management strategies There is evidence that excess calcium not only decreases the utilization of P but also increases the pig’s requirement for zinc in the presence of phytate (NRC 1998). Little is known about the Ca availability to swine from natural feedstuffs. Because of the phytic acid content, the bioavailability of calcium in cereal grain-based diets, alfalfa, and various grasses is relatively low so calcium must be supplemented. Calcium availability of calcitic limestone, gypsum, oystershell flour, aragonite and marble dust is high but the calcium availability in dolomitic limestone is only 50 to 75%. Little data exists for the availability of dicalcium phosphate, tricalcium phosphate, defluorinated phosphate, calcium gluconate, calcium sulfate and bone meal to swine but these are all highly available calcium sources in poultry diets. Because of the close association between P and Ca availability in the diet and the influence of phytate on the availability of both of these minerals, the effect of added phytase on P digestibility is complex and requires more research in sows and grower/finishers.

In general, as illustrated in Table 2.8, most of the excreted P can be found in the feces of sows and finisher pigs (Poulsen et al. 1999; de Lange et al. 1999). For sows, the category of swine more likely to receive excess dietary P, 23-32% of P intake may be excreted via the urine while

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only 7-11% of P intake was excreted in the urine of finisher pigs (Poulsen et al. 1999). Although the absorption of P in the intestinal tract is somewhat regulated, excess P beyond that required for growth and maintenance, can be absorbed by the pig (Finco 1989). The surplus will be excreted either as endogenous P via feces or eliminated by the kidneys as inorganic P. Adequate phosphorus intake is critical for first-parity sows. Lack of adequate phosphorus in the sow diet can result in posterior paralysis or paralysis of the hind legs, a problem occurring most frequently in sows producing high levels of milk toward the end or just after termination of lactation. High phosphorus is often included in the diet because voluntary feed intake by sows may be reduced by high environmental temperatures (NRC 1998). Understandably, there is a great reluctance by producers to limit phosphorus in sow diets. Although porcine Somatotropin (pST) reportedly increases P retention and reduces P in the fecal output, use of the hormone may have detrimental effects on bone mineralization in finishing pigs fed the recommended levels of Ca and P (Carter et al. 1999).

Table 2.8. Phosphorus (P) consumption, retention and losses. Mean values from France, The Netherlands and Denmark (Poulsen et al. 1999)

France Netherlands Denmark Absolute Relative Absolute Relative Absolute Relative Sows (incl. piglets until weaning): Consumption (kg/year) 6.98 100 5.39 100 8.03 100 Retention (kg/year) 1.56 22 1.35 25 1.11 14 Excretion: Total (kg/year) 5.42 78 4.04 75 6.92 86 Fecal (kg/year) 3.84 (55) 2.58 (48) 4.42 (54) Urine (kg/year) 1.58 (23) 1.46 (27) 2.50 (32) Weaners (piglets from weaning): Consumption (g/pig) 270 100 157 100 310 100 Retention (g/pig) 120 44 97 62 120 39 Excretion: Total (g/pig) 150 56 60 38 190 61 Fecal (g/pig) 130 (48) 53 (34) 170 (55) Urine (g/pig) 20 (8) 7 (4) 20 (6) Growing Pigs: Consumption (g/pig) 1400 100 1160 100 1110 100 Retention (g/pig) 480 34 430 37 390 35 Excretion: Total (g/pig) 920 66 730 63 720 65 Fecal (g/pig) 770 (55) 650 (56) 610 (55) Urine (g/pig) 150 (11) 80 (7) 110 (10)

A strategy to reduce excreted P during the growing/finishing period is to use phase feeding to balance the nutrient requirements to the size of the animal. A survey by the U.S. Department of Agriculture found that the number of diets used in a phase feeding program was dependent on size of the producer. For producers marketing more than 10 000 pigs per year, 74% used four or more diets in their phase-feeding program. Only 28% of the producers marketing less than 2000 pigs per year used four or more diets. For producers marketing 2000 to 10 000 pigs per year, 57% used four or more diets. Very few producers used only one diet for the grow-finish period (<5% of the smallest producers and <1% for larger producers). Phase-feeding reduces feed-cost

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by decreasing the quantity of protein and phosphorus supplied in the late-finisher diets (Han et al. 1998). Application of phase-feeding has been made easier with the use of all-in, all-out technology.

Fine grinding and pelleting are also effective ways in improving feed utilization and decreasing dry matter (DM) and nutrient excretion. When particle size is reduced to less than 600 µm in sorghum, corn and barley based diets, feed performance is enhanced but fine grinding does not improve the digestibility of wheat based diets (Nahm 2002). The effect of particle reduction on P digestibility is not well documented in the literature.

Improper mixing of feeds results in reduced uniformity of the diet, leading to poor animal performance and increase nutrient excretion into the environment. Nahm (2002) presents evidence that excess nutrients are used in feeds to compensate for variability in ingredient composition and accuracy of diet mixing in order to prevent nutrient deficiencies. Inaccuracies in weighing and mixing may account for some of the discrepancies found in the efficacy of phytase addition. There is still some debate about the effect of pig age/weight on the response to phytase (de Lange et al. 1999).

Controlling feed wastage improves herd feed conversion and reduces nutrient losses. Feed wasted in the manure pit can add considerably to the nutrients that need to be applied to cropland (de Lange et al. 1999). Wet-dry feeding systems can significantly reduce wastage of both feed and water. Current research has shown that manure volume per pig can be reduced from 30 to 50% using wet-dry feeding systems. However, the nutrient concentrations in the manure from a wet-dry feeding system are generally significantly higher therefore routine manure analyses are needed to adjust application rates of such manure to cropland. Maintaining pigs under comfortable environmental temperature and humidity conditions will improve feed utilization and can also reduce nutrient excretion. Cold temperatures increase caloric requirements for body maintenance, increasing feed intake and nutrient excretion. Likewise, extremely hot temperatures reduce feed intake, decrease growth rate and increase time to market, thereby also increasing maintenance requirements and nutrient excretion. Raising genetically lean pigs (rather than fat ones) on diets that meet their requirements, controlling diseases and parasites, and using good management practices are further examples of how to improve feed conversion efficiency and reduce nutrient excretion. Van der Peet-Schwering et al. (1999) cite a Dutch report of the estimated effect of different feed and management measures on P and N excretion of sows and growing pigs (Table 2.9).

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Table 2.9. Estimated effect of several feed and management measures on P excretion per sow per year (kg) and per growing pig per year (kg) * Effect on P-excretion Sow Growing pig Reducing mineral content in the dieta -0.55 -0.29 Phase feedingb - -0.07 Optimize feed intakec -0.22 -0.05 Reducing feed wastaged -0.22 -0.09 Reducing initial live weighte - -0.07 Reducing weight at slaughterf - -0.03 Critical temperatureg -0.13 - a P- content reduce from 4.8 to 4.4 g/kg and 4.8 to 4.3 g/kg for growing pigs and sows, respectively. b Three-phase feeding instead of two-phase feeding. c The feed intake optimized by split-sex-feeding, restricted feeding to barrows and optimizing feed intake of sows which will result in a reduction in feed intake by 20 kg/sow/yr. d Feed wastage can be reduced by 0.05 kg per growing pig per day and by 20 kg/sow/yr. e Reducing live wt at start of growing period by 3 kg. f 2% instead of 5% of the growing pigs are too heavy at delivery. g Maintaining the building at the lower critical temperature for sows. * Adapted from van der Peet-Schwering et al. 1999.

According to Lynch and Caffrey (1997) although pig feeding has the potential to utilize food industry by-products with economic benefit to the pig producers and reduction of tipping fees for food processors, formulation of diets with lower dietary P will render some of these materials, e.g. cereal by-products, less attractive.

To improve P distribution, conserve water and reduce odour and ammonia emissions, one consideration might be to adopt a different rearing system for feeder pigs and dry sows. Low cost Biotech shelters bedded with wheat straw have bee used successfully to commercially raise feeder pigs in Manitoba. The manure-straw mixture can be composted and applied to meet the phosphorus requirements of annual and perennial crops. Because the N in compost is primarily organic, leaching of N, even in light soils, is not a concern.

1.16.2.3 Summary of strategies for P decreases According to Poulsen (2000) there are several ways to improve P utilization in pig production:

• Definition of precise recommendations o Digestible P instead of total P o Phase feeding

• Improving P digestibility in feedstuffs o Addition of phytase (intrinsic or microbial) o Changing feeding practices (soaking or liquid feeding) o Processing (alternatives to heating and pelleting which inactivate phytase)

• Better choice of inorganic P supplements o Feed phosphates with high P digestibility when supplementation is needed

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• Herd management o Improving lean growth potential as a result of feeding entire males rather than

barrows will reduce P excretion by 15% o Control of parasites that damage intestinal mucosa

• Plant breeding o Lower phytate content o Greater phytase activity o Increase thermostability of phytase

1.16.3 Feeding and Management of Ruminants

1.16.3.1 Phosphorus requirements Ruminants differ from monogastrics in their requirement for P since the microbial population of the rumen increases the availability of phytate P to the animal rendering nearly all the phytate phosphorus (>97%) available for absorption. Phosphorus is an essential mineral of microbes therefore this population has a discrete requirement for P. Low dietary P is associated with reduced microbial protein flow into the small intestine even when food intake is not reduced. This suggests that the long-term effects of digesting low P diets may be similar to the effects of low nitrogen (protein) diets (Ternouth 2001). A particularly large pool of P in ruminants is associated with the gastro-intestinal tract, particularly the rumen. The size of this pool is the result of the large volume of saliva secreted by ruminants and the ability of the glands to concentrate P (commonly 10-fold) from plasma to serous saliva. It has been estimated that the entire plasma pool of P may be secreted into the rumen in a period of a few hours. The majority of P in the rumen (as much as 90%) may be of salivary rather than dietary origin (Challa et al. 1989). Not only is this pool substantial but it complicates our understanding of the measurement of the P requirements of cattle and the determination of dietary deficiency of P. It is possible to determine the amount of mineral present in milk and body gain with a high degree of accuracy. Maintenance requirements and absorption coefficients are more difficult to accurately measure and as a result, studies estimating these two factors are scarce for most minerals. Maintenance requirement are estimated from endogenous fecal losses comprised of phosphorus from saliva, microbial phospholipids, nucleic acids and sloughed off intestinal cells (Spears 2002). Effectively the available dietary and salivary P enter a single gastrointestinal pool so they both follow the same kinetic pathways and have the same true absorption coefficients.

The mechanism of adaptation to alterations in dietary phosphorus is mediated through saliva P which is directly correlated with P in plasma. A low-P diet leads to a decrease in salivary P which results in low fecal P, alternatively, excess P in the diet leads to elevated concentration of P in plasma thereby increasing salivary P secretion, decreasing the adsorption efficiency of salivary P and increasing fecal losses (Figure 2.6). Typically, 95 to 98% of total phosphorus excretion is in the feces (NRC 2001). Very little phosphorus is excreted in the urine in ruminants (Satter and Wu 2001; Chase 1998). Damage to the intestinal wall by parasites (round worms) can reduce the absorptive capacity of the gut and reduce efficiency of P recycling.

The 2001 NRC adopted research results developed since the release of its previous version (1989 NRC) and made several modifications in estimating the P requirement for lactating dairy cows. The requirement is calculated as the sum of absorbed P utilized for maintenance, lactation and pregnancy. Absorbed P is then divided by the availability of feed P in the gut.

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Absorbed P Maintenance: 1.0 g/kg of feed intake Lactation: 0.9 g/kg of milk Gestation for 190-280 days: 2.5 g/d

Feed P availability Forages: 0.64 Grains: 0.70 Dical: 0.75 Mono Phos: 0.90

The requirement for growth (g/kg average daily gain) is described by an allometric equation from data in the literature regarding growing cattle:

P (g/day) = (1.2 + (4.635 x MW-0.22)(BW-0.22)) × WG where MW = expected mature live body weight (kg), BW = current body weight and WG = weight gain. The absorbed phosphorus requirements (g/kg average daily gain) ranges from 8.3 g at 100 kg live BW to 6.2 g at 500 kg.

The 2001 NRC recommended an overall lower dietary P amount than the previous recommendations. Apparently it is still a common practice for nutritionists or consultants to formulate rations with P levels exceeding the animal’s requirements (Wu and Ishler 2002). According to Wu and Ishler (2002) some reasons why rations are over formulated for P include: 1) added safety margins to account for animal variation and variation in P content of feeds, feed delivery accuracy, feed bunk management, 2) uncertainty about P availability, and 3) the view that addition of P to the ration improves reproduction (Klopfenstein et al. 2002). The NRC recommendations already contain a safety margin and is adequate for maximum performance.

Alternatively Valk and Beynen (2003) have advanced a proposed assessment of phosphorus requirements of dairy cows. The basis for their recommendation is that fact that the current requirements in the Netherlands were developed based on studies with non-lactating sheep and goats. Furthermore the relationship between the P requirement for producing milk, intestinal absorption and secretion of P in saliva has not been described under P deficient conditions which they indicated is necessary to establish true absorption of P. They suggest that P requirements of lactating and non-lactating dairy cows can be calculated as follows:

P requirement (g/day) = [19 + (0.14 × kg milk)] + (kg milk × 0.9)/(0.70)

The constant of 19 was determined by adding an endogenous P loss of 10.0 g, a safety margin of 0.9 g and 2.3 g for fetal growth. Because the true absorption coefficient of dietary P is only 70%, dry cows would need 19 g P/day to meet the net requirement for maintenance and fetal growth of 13.2 g P/day.

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Figure 2.6. Reabsorption of P from bone is increased by the parathyroid hormone (PTH). Phosphorus absorption from the gastro-intestinal tract is increased by the hormone 1,25 di-hydroxycholecalciferol (1,25(OH)2D3) which is a metabolite of Vitamin D3. Phosphorus absorption from the intestinal tract is also related to P intake (Challa et al. 1989).

The level of P found in cattle rations is quite variable due to diverse crop management, regional crop differences and environmental conditions (Table 2.10). The variation in P content of forage varieties grown on the same farm may be small but available P in dairy diets might vary considerably as producers try to take advantage of low cost opportunity feeds like distillers grains or other crop residues. The P level of pastures varies with fertility management, maturity of the stand and the ability of the plant to extract P from the soil (Ternouth 2001). If the P content of the forage is high, there may be difficulties in providing enough energy and to lactating animals without providing a large oversupply of P (Valk et al. 2000). The effect of limiting the amount of P fed to dairy cows and subsequent P excretion can only be precisely managed in intensive farming systems which rely heavily on the use of premixed diets (or total mixed rations [TMR]). However the problem then becomes one of managing the vast amounts of imported nutrients in an environmentally sound manner.

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Table 2.10. Phosphorus content of selected feedstuffs z Feed No. of samples Average (% of

dry matter) S.D.y Normal

range Legume hay 11 962 0.26 0.06 0.21-0.32 Legume silage 3 384 0.32 0.06 0.27-0.38 Grass hay 2 136 0.24 0.08 0.16-0.32 Grass silage 2 030 0.31 0.07 0.24-0.38 Corn silage 16 992 0.23 0.03 0.10-0.26 Beet pulp 106 0.10 0.03 0.06-0.13 Brewers grains 27 0.62 0.06 0.56-0.68 Corn grain 306 0.32 0.11 0.20-0.43 Corn gluten meal 36 0.90 0.21 0.68-1.11 Distillers grains 183 0.82 0.12 0.71-0.94 Feather meal 6 0.28 0.06 0.22-0.33 Fish meal 31 3.39 1.14 2.25-4.53 Molasses 60 0.68 1.20 Up to 1.88 Soybean meal 277 0.68 0.11 0.57-0.79 Wheat 31 0.47 0.23 0.24-0.69 Wheat middlings 90 0.88 0.21 0.67-1.08 z Source: Chase 2000 y S.D. = standard deviation

1.16.3.2 Feeding and management strategies for dairy cattle Potential P loss on dairy farms may be related not only to how much P is excreted in manure and applied to fields, but also how easily the manure P is dissolved in rainwater and subject to potential runoff loss. That is, the chemical forms of P and their relative proportions in manure play an important role. Analysis of fecal samples collected from three feeding trials (Dou et al. 2002), indicated that increasing dietary P levels through the use of P minerals not only led to a higher concentration of acid-digest total P in feces, but more importantly increased the amount and proportion of P that was water soluble and thus most susceptible to loss in the environment. In diets containing 3.4, 5.1 or 6.7 g P/kg feed dry matter, the water soluble fraction of fecal P was 2.91, 7.13 and 10.46 g/kg fecal dry mater respectively, accounting for 56, 77, and 83% of acid digest total P (Dou et al. 2002). The implication of feeding excess P in dairy diets on runoff P was demonstrated in work by Ebeling et al. (2002). Concentrations of dissolved reactive P in water runoff were almost ten times higher from soil receiving manure from cows fed a diet containing 0.49% P compared to manure from cows fed 0.31% P in the diet. Since available P in dairy diets is difficult to measure, Dou et al. (2002) proposed a fecal P indicator concept in which the readily soluble P in excreta is used as a reflection of P consumption and utilization by animals. Readily soluble P would be compared to benchmark values for recommended P intake to provide feedback for lowering dietary P intake.

Diet formulation and management skills are paramount to the reduction of excreted P from cattle. Since cattle can utilize organic sources of P efficiently it should be possible to formulate dairy rations using only plant sources. Factors such as dietary protein content, starch

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degradability and grain content of the diet may affect availability of organic feed P (Knowlton et al. 2001). Knowlton et al. (2001) concluded that organic sources of P, such as phytate P in wheat bran can be used to provide a substantial portion of the P needed to supplement dairy cattle diets after peak lactation. However, feeding wheat bran during early lactation should be limited to prevent suppression of dry matter intake and reduced P retention. Phosphorus could be reduced in the ration by using protein supplements that are lower in P. Reduced use of bone and meat meal due to the risk of bovine spongiform encephalopathy has already led to the use of plant proteins which contain less P.

Several new technologies have the potential to increase milk production efficiency and reduce relative P excretion per unit of milk produced. One such technology is the manipulation of photoperiod by the provision of artificial lighting. It has been shown that increasing day length can increase milk production in dairy cattle by up to 8%. Nutrient intake required by such light-stimulated herds increased by 4.1% and P excretion increased by 2.8% when compared with similar herds under natural day length. The administration of bovine growth hormone (BST) can increase milk production by as much at 30% in certain cows within a herd although milk production in the entire herd may only increase by 14%. The nutrient requirements of a herd treated with BST would increase by approximately 7 to 8% and manure P may be increased by 5%. The nutrient loss per unit of milk produced would decrease by 8 to 10%. Research has also demonstrated that milking three times per day instead of twice per day can increase production per cow by an average of 11%. This greater production per cow results in the consumption of 5% more protein with 3.5% more nutrients excreted in manure. The extra milking per day reduces the amount of manure nutrient losses per unit of milk produced by 7% (FASS 2001).

1.16.3.3 Feeding and management strategies for beef cattle The rearing of large numbers of cattle in feedlots presents considerable environmental challenges one of which is redistribution of manure nutrients back onto crop land. Normally feedlot cattle consume large amounts of feed grain that may be shipped from distant parts of the country. Diets that are grain-based often supply phosphorus in excess of the needs of cattle even in the absence of supplemental P. In fact, balancing feedlot diets to contain the appropriate Ca:P ratio of 2:1 is very difficult because of the high phosphorus levels in feed concentrates (Shannon Scott, AAFC – personal communication). Loss of nitrogen in the feedlot by ammonia volatilization results in an organic fertilizer with a low N:P ratio which is not well utilized by fast-growing crop plants at least in the first year of application. Historically, manure has been perceived as an unpredictable source of fertilizer for crop production. Even today many grain producers will not accept manure as a fertilizer because of the perceived difficulties in balancing nutrients to achieve acceptable crop yields.

The phosphorus requirement of feedlot cattle and calves (0.14-0.20%) is low (NRC 2000; Erickson et al. 2002). With the current technology, it is not clear how to practically reduce P excretion in feedlot cattle below that provided in feed grain other than to revolutionize feeding and production strategies to include substantial amounts of low phosphorus forages in the diets. The current best management practice is to remove all supplemental P in the ration, but this requires that alternative sources of calcium supplement be included to balance the Ca:P ratio. Phase feeding can reduce P intake by 51% for yearlings and 41% for calves and subsequent P in the manure by 59 and 38%, respectively (Klopfenstein and Erickson 2002). In summarizing earlier literature, Klopfenstein et al. (2002) suggest that P requirement for growth and

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bioavailability of P are not well established for beef cattle and are factors worthy of further research.

As with dairy cattle, P balance studies in beef cattle for the purpose of developing recommended levels of P intake are difficult due to extensive cycling of P in the saliva. In vivo sampling using ultrafiltration membrane probes may provide a tool for investigating the chemistry of P as well as other minerals (Janie et al. 2001). Membrane probes which can be implanted in bone, muscle and subcutaneous tissues to sample interstitial fluid, permit the investigation of Ca, Mg and P concentrations in locations not previously accessible for direct in vivo measurement. One of the benefits of the membrane probe sampling is that different tissues can be sampled simultaneously so the device may be valuable in the study of differences in distribution of these minerals and in the effect of diet and management on P balance.

1.16.3.4 Feeding and management strategies for small ruminants Most feeding and management considerations for large ruminants apply to small ruminants with the exception of the effect of P on the formation of urinary calculi. The composition of urinary calculi varies with geographic location, but the most common uroliths are calcium apatite and phosphatic calculi (calcium hydrogen phosphate dihydrate, magnesium ammonium phosphate). Formation of phosphatic calculi is encouraged by high concentrate, low roughage, low Ca:P ratio, high magnesium diets and alkaline urine (Stewart et al. 1990). Normally, phosphorus is recycled through saliva and excreted via feces in ruminants. High grain, low roughage diets decrease the formation of saliva and increase the amount of phosphorus excreted in the urine. High phosphorus diets overwhelm the salivary excretion mechanism and result in high urinary excretion of phosphorus. High calcium diets are effective at reducing the absorption of phosphorus from the gastrointestinal tract. Increased urine output and decreased urine pH may prevent the formation of uroliths by decreasing urine retention time and by dilution of solutes. Some breeds of sheep may be predisposed to urinary excretion of phosphorus versus fecal excretion (Texel, Scottish Blackface).

1.16.3.5 Summary of strategies for P reduction and management for ruminants

• Herd management o Increase amount of milk produced per kg of P excreted. o Maintain a parasite control program.

• Feed management o Test forages for P as an input for ration formulation. o Include more low-P forages in feedlot rations. o Improve the determination of P requirements of high-producing dairy cows. o Improve the determination of P requirements for growth in beef cattle.

• Determine water-extractable P in manure and composts. • Rapid analysis of P bioavailability of various feed ingredients.

1.17 Manure Treatment and Utilization Technologies to Manage Excreted Phosphorus Some of the manure treatment alternatives currently available or under development offer the potential for removing P from manure or rendering manure P less mobile in the environment. Generally a manure treatment option is selected based on the species, housing system,

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consistency of the manure produced and the price. The three main nutrient issues for on-farm treatment of manure are: N loss through ammonia volatilization, excess P, and the balance of N to P (N/P ratio). Some of the treatments to decrease the solubility of P also improve nitrogen retention but very frequently manure treatments increase N loss.

1.17.1 Manure Treatment Alternatives

1.17.1.1 Solids removal In cold climates, slurry animal wastes are often stored for several months in lined earthen basins, storage tanks (above or below ground) or pits beneath slated floors, until conditions are favourable for land application. Liquid-solid separation has been viewed as a method to improve the pumping and irrigation characteristics of liquid manure, to reduce the sludge accumulation which decreases total storage volume or to generate solids for composting. Liquid-solid separation via gravity settling has been used extensively to reduce the solids content in feedlot runoff and flushed dairy manure. Mechanical separation techniques have been widely used with flush manure handling systems in dairy housing facilities but are generally less popular in swine manure handling systems. The major hindrance of using this technique rests with the high capital and operational costs and low separation efficiency (Zhu et al. 2000). The interest in mechanical separation techniques stems from the desire to limit sludge accumulation and decrease objectionable odours emanating from storages, as well as decrease frictional resistance during long distance pumping. Recent progress with combination treatments, which include natural, mechanical and chemical treatment, offer the greatest potential to reduce manure phosphorus concentrations and produce a more balance nutrient for land application. The following is a brief overview of solids removal methods and phosphorus removal capability.

1.17.1.2 Natural sedimentation The easiest and least energy consumptive method for solid-liquid separation is to let stored waste slurries settle. Within a storage container, the suspended and dissolved solids (total solids, TS) will stratify depending upon particulate size and the specific gravity of those particles (Ndegwa et al. 2002). There appears to be a general trend in TS and nutrient stratification patterns (see Figures 2.7 and 2.8) although specific quantities of sediment will depend on the type of operation and management scheme (Ndegwa et al. 2002). Sedimentation is usually the first operation in the waste removal processes. The supernatant (liquid portion on top) can be applied to fields, or after further processing, used as wash water. Reducing the water content of the settled sludge usually results in reduced mass making handling easier, but this also concentrates the nutrients making handling more environmentally risky. Sludge can be mechanically dewatered, digested anaerobically, composted or directly applied to land. Although most land application of sludge will be in the form of a well-mixed slurry to ensure even nutrient spreading.

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1.17.1.3 Mechanical separation A recent literature review of options for mechanical solid-liquid separation of livestock manure revealed some deficiencies in the scientific literature with regard to standardization of protocols for measuring and reporting the relative efficiencies of three categories of separators namely, screens, centrifuges, and presses, used alone or in combination (Ford and Fleming 2001). Few of the cited researchers actually reported the percent removal of total phosphorus and, where phosphorus removal efficiency was reported (Chastain 2001a; Holmberg et al. 1983 ; Chastain et al. 1998; Converse et al. 1999; Converse et al. 2000; Fernandes et al. 1988), the forms of phosphorus present were not described. Ford and Fleming (2001) concluded that phosphorus is not easily concentrated in the separated solids portion. In all but one of the studies reviewed, less than 30%of the TP was removed in the solids fraction for swine and dairy manure. Research by Burns and Moody (2002a) indicates that total phosphorus was preferentially partitioned into the solids after passage through a screw press, with a greater proportion of soluble phosphorus remaining in the press cake from those manures with higher initial total solids. In a more recent study, Møller et al. (2002) found that removal efficiency for total phosphorus was highly variable for different manure types. For freshly collected pig and cattle manure, the removal efficiencies for the screw press were 7.12 and 15.46, respectively while the removal percentage obtained with the decanting centrifuge was 62.28 and 82.00, respectively. Storage reduced the DM content of manure significantly but separation of total phosphorus was only slightly reduced. Recovery of total phosphorus from an anaerobically digested mixture of manures and organic waste was more efficient using the decanting centrifuge compared to the screw press,

Figure 2.7. P-content of manure with depth in pig finishing and nursery barn pits (from Ndegwa et al. 2002).

Figure 2.8. Total solids in manure with depth in pig finishing and nursery barn pits (from Ndegwa et al. 2002).

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52.35-90.95 vs 8.68-10.30, respectively (Møller et al. 2002). Biological decomposition during storage or during anaerobic digestion contributes to the transfer of nutrients, especially nitrogen and phosphorus, between different fractions and chemical forms in manure (Henze 1997). More soluble phosphorus could be expected to be recovered in the separated liquid fraction after storage or digestion of manures.

1.17.1.4 Nutrient inactivation Nutrient inactivation, as a component of stormwater treatment and lake restoration, has been used for over thirty years. Binding of phosphorus into a form that is not bioavailable can be an essential part of watershed management. Algae have no roots and thus require all nutrients to be soluble. The difference between a eutrophic water and “pristine” water is as little as 30-ppb phosphorus—thus a small reduction in phosphorus loading can have a dramatic impact on water quality (Lind and Gribble 2002).

The best phosphorus-binding chemical produces a precipitate that makes the P insoluble over a large pH range. Aluminum phosphate is the least soluble of the three common precipitants over a pH range of 2 to greater than 9. Neither aluminium nor phosphorus is released back into the environment under normal soil or water conditions. (Stumm and Morgan 1981). The US EPA performed the first lake treatment project in 1970 with alum (aluminum sulphate (Al2(SO4)3) and since that time alum has been used for lake rehabilitation, nutrient inactivation, stormwater treatment, and phosphorus precipitation. Iron is also a suitable P binding chemical and its presence in a soil or sediment is largely responsible for much of the “background” P binding present in the environment. Ferric phosphate (FePO4) is also quite insoluble, being about one order of magnitude more soluble than aluminum phosphate. However, under anoxic conditions, ferric phosphate will be reduced to soluble ferrous phosphate thereby releasing the phosphorus back into the water or soil as a nutrient. The oxygen content of the soil or water does not affect aluminum compounds. Calcium binds phosphate compounds best at high pH (>10). At pH values normally found in soil and water (<8.5) they are soluble and used as fertilizers and feed additives. Calcium phosphate compounds are several orders of magnitude more soluble than iron and aluminum phosphates and are not as efficient forms in which to bind phosphorus.

In animal growing operations producing liquid waste streams, the challenge of handling strong waste with high but intermittent flows demands innovative technology not common to traditional waste systems. Ideally a system for treating dairy flush water would be durable, simple to install and use, capable of recovering fiber while removing phosphorus and clarifying the water. Additional physical or biological treatment components can be added to effect solids disposal, nitrification, denitrification, ammonia and BOD removal. Numerous field trials and commercial application of alum in poultry houses have been known to result in soluble phosphorus reductions of up to 84%. In liquid waste applications like swine slurry and cattle feed lot runoff, the reduction in soluble phosphorus can exceed 95% (Lind and Gribble 2002). The salts that effectively precipitate dissolved phosphorus also aid in the flocculation of suspended solids, which enhances settling. Polyacrylamides (PAM) are high molecular weight long-chain water-soluble polymers. The polymers may be cationic, anionic or non-ionic depending on their net charge. Cationic polymers have been shown to be much more effective than anionic or neutral polyacrylamide polymers in removing organic solids from swine manure wastewater (Vanotti et al. 1996).

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Alum (commercial grade, 48.5% Al2(SO4)3·14 H2O) is commonly used to remove P during sewage treatment; hence it is widely available at a reasonable cost. It is currently used as a litter amendment in the poultry industry, primarily to chemically sequester P (Sims and Luka-McCafferty 2002; Maurice et al.1998; Moore and Miller 1994). However, alum seems to have the added benefit of reducing ammonia emissions and substantially increasing available nitrogen in broiler houses (Worley et al. 1999). To reduce soluble P in poultry litter, alum is commonly added at a rate of 10% of manure dry weight. Application of one-half the recommended rate of alum to houses that have not been cleaned out after the previous flock, has been found to be a viable alternative to either full rate application or top dressing with fresh shavings, although soluble P and ammonium N losses were higher at the half-rate (Worley et al. 2000). It was thought that a reduction in alum addition rate might make the technology more attractive to the growers, thus increasing the use of alum in poultry production systems. While the environmental benefits of one farm using less alum may be reduced, the effect industry wide may be increased if more growers will use it. In search of a more inexpensive replacement for alum, Codling et al. (2000) demonstrated that the addition of aluminum-rich residues from well water treatment with alum and calcium hydroxide, or the addition of an iron-rich by-product from TiO2 production containing 62 and 13 g/kg aluminum and 204 and 238 g/kg iron, respectively, to poultry litter reduced both the water and acid extractable P. These two residues were dried and ground before adding different amounts to poultry litter. The general ranking of the residues in terms of their ability to decrease extractable P levels in poultry litter was aluminum-rich > iron-rich (Table 2.11).

Table 2.11. Percent reduction in water-soluble P and Bray and Kurtz No.1-extractable P in poultry litter amended with Al-rich or Fe-rich residues z

Incubation time Residues

Addition rate Week 2 Week 4 Week 7

g/kg Reduction in water-soluble P, % Aluminum 25 26 25 48 50 59 63 72 100 58 87 88

Iron 25 13 14 39 50 24 10 41 100 43 42 62 Reduction in Bray and Kurtz No.1-Extractable P, % Aluminum 25 24 34 38 50 51 60 60 100 81 84 82

Iron 25 3 20 19 50 13 24 22 100 30 45 44

zAdapted from Codling et al. 2000. Ndegwa et al. (2000) observed 77 and 85% P removal in sediment following addition of 1.5 and 2.0 g/L Al2(SO4)3, respectively, to swine manure containing 1 and 2% total solids. Settling removed 42% of P in the control samples. Adding ferric chloride (FeCl3) at these same concentrations resulted in P removals of 84 and 92%, respectively (Ndegwa et al. 2000).

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Similarly, Smith et al. (2001a,b) found that treatment of swine manure with alum and aluminum chloride (AlCl3) could reduce total dissolved phosphorus as well as soluble reactive phosphorus in swine manure. Relatively high levels of AlCl3 (0.75% of final volume) must be added to the manure to be effective (Smith et al. 2001a). Low rates of alum additions to liquid swine manure (Al:P ratio of 0.5:1) reduced soluble P levels to 30 mg P/L from 200 mg P/L (Smith et al. 2001b). High rates of alum additions (Al:P ratio of 1:1) reduced soluble P concentration by two orders of magnitude (near 1 mg P/L). Powers and Flatow (2002) evaluated the use of chemical flocculants in settling solids and phosphorus in liquid swine manure. Manure was generated from pigs fed diets containing varying amounts of non-phytate phosphorus and dietary amendments aimed at improving phosphorus utilization over a 3-phase feeding period. Data pooled across diets and feeding period indicated that (Al2(SO4)3) at an addition rate of 625-mg/L was the most effective addition for high P removal but was no more effective than 250 mg/L for total solids removal (Table 2.12).

Table 2.12. P removal from liquid swine manure across dietary treatments and feeding periods (3-level phase feeding), as affected by different flocculants and flocculant concentrations z

Flocculant concentration Flocculant

40 mg/L 250 mg/L 625 mg/L Phosphorus Removal (%) Al2(SO4)3 21.0 73.9 90.9 CaCO3 15.0 17.6 21.2 CaO 26.7 55.4 66.7 FeCl3 30.0 82.7 78.2 FeSO4 21.8 41.7 49.1 Control 19.0 17.8 14.6 p-values <0.0001 <0.0001 <0.0001

zAdapted from Powers and Flatow (2002). Barrow et al. (1997) reported P removals in 1% total solids dairy flushwater ranging from 78 to 88% when FeCl3 was added at concentrations ranging from 69.5 to 278 mg/L Fe. Only 40% of the P was removed in the control samples. Sherman et al. (2000) demonstrated in simulation studies that reduction of P in dairy manure wastewaters to low levels is possible with flocculants such as alum or FeCl3 (Figure 2.9). The results of their laboratory studies indicated that the cationic polyacrylamide polymers (PAM) when used at low levels (1, 2, and 4 mg/L) were not as effective as the chemical flocculating agents. The economics of the procedures tested in their laboratory were not favorable for general use but they indicated that there may be certain situations where flocculants could provide producers with a method to help them achieve regulatory compliance. Sherman et al. (2000) used an example of a dairy that produces 800 000 L/d of dilute flushwater (approximately that generated from a dairy having 500 cows). If a 0.9 mL/L alum addition was chosen, the dairy would use 720 L alum/day with a total weight of 961 kg. If the liquid alum that contains 48.5% dry aluminum sulfate is obtained for $80 US per tonne, 961 kg would equate to $77 US/day. A year of treatment at the 0.9 mL/L level would cost $28 000 US. Based on field tests with a continuous flow flocculating tank, the amount of P expected to be recovered would be 10 337 kg/yr. If P were valued at $1.60 US per kg, actual P

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based on equivalent fertilizer P would be valuable ($16 500 US/yr) but still much less than the cost of flocculant.

The cost of equipment and labour would be in addition to the capital needed to operate the system. Sherman et al. (2000) also indicated that more research is needed on the bioavailability of P contained in the precipitated solids. Low bioavailability of P would reduce the use of the precipitated solids as a crop fertilizer. On the other hand, using a simple alum injection system for dairy wastewater, Jones et al. (2002) found that only 0.47 mL of liquid alum/L lagoon water (27.6 mg Al/L) was required to reduce soluble reactive P from 11.3 to 1.0 mg P/L in the supernatant. Estimated increases in soil aluminum content in the surface 15 cm of soil during 20 yr of supernatant application at 39 600 L/day to 8.1 ha is only 0.33%. Estimated chemical cost for reduction of soluble reactive P to 1.0 mg P/L was $8.04 US/kg P removed.

0

10

20

30

40

50

60

original 0 0.9 1.8

alum added (mL/L)

Efflu

ent P

(mg/

L) Batch, fieldBatch, lab3-fill, field3-fill, lab6-fill, field6-fill, lab

Figure 2.9. Reductions in P concentrations in low-solids dairy wastewater collected under field conditions and treated with alum solution. Batch, field = alum solution metered into wastewater while 3500 L tank filled; three-till and six-fill (continu08s flow), field = alum addition to manure wastewater as 2100 L tank filled and overflowed through skimmer pipe until three or six volumes were completed, lab treatments were with the same influents in 1L batch volumes in the laboratory (adapted from Sherman et al. (2000).

In a series of complex experiments Dao and Daniel (2002) tried to develop an improved understanding of key manure characteristics that control particulate aggregation in dairy slurry. Their goal was to determine whether organic water treatment polymers and mineral P immobilization chemicals altered P release from the treated manure solids as well as amended soil. Working with dairy manures altered to achieve total soluble solids content of 30 (TSS30) and 100 (TSS100)g/L simulating manure from flushed and scraped manure collection systems, they tested 3 cationic polymers, aluminum sulfate, ferric chloride and coal-combustion ash

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(Class C) at 3 rates of addition. Fly ash consistently reduced solution-phase dissolved reactive P at all rates of addition by 52 and 71% in the TSS30 and TSS100, respectively. Fly ash reacted with phosphate anions to reduce solution dissolved reactive P and suppressed P solubility in the particulate fraction. Their findings with the metal salts suggested than an adequate amount of metal salt promoted both aggregation of the manure solids and P immobilization. However over-addition reversed and negated the benefits of P reduction by metal salts in the presence of organic polymers. These results illustrate the complexity of the chemistry of Al3+ and Fe3+ in manure wastewaters. Dao and Daniel (2002) concluded that the role of particulate and dissolved organic matter in metal phosphate and polyphosphate solubility needs further investigation. Upon examining P solubility of these treated manure suspensions on soils, they found that dissolved reactive P was reduced by 63 to 96% in soil amended with manure treated with mineral additions alone, and 40 to 76% in soil amended with manure treated with a mixture of polymer-mineral amendments. Fly ash-treated manure solids remained stable and sequestered dissolved reactive P (Dao and Daniel 2002). Although the addition of these chemical and organic agents are effective in reducing solubility of P there is still concern that this process may cause secondary environmental pollution by raising the concentrations of chlorides and sulfates in soil (Zhu et al. 2001b).

Polyacrylamide (PAM) polymers have been tested in furrow irrigated agriculture for erosion control and increased infiltration (Lentz et al. 1992; Lentz and Sojka 1994; Sojka et al. 1998a,b). Soil PAM-treatments reduce sediment loss rate over time with improvement of the runoff water quality parameters ortho-P, total-P, NO3

- and biological oxygen demand (Lentz et al. 1998, 2000). Entry et al. (2003) attempted in-field swine manure treatments using water soluble PAMs developed for use in erosion control and known to be safe for a variety of food, pharmaceutical and sensitive environmental applications (Barvenik 1994). The treatment application consisted of the creation of furrows measuring 0.3 m wide x 0.2 m deep x 50 m long. A mixture of 35 g PAM + 750 g Al2(SO4)3 or a mixture of 35 g PAM + 750 g CaO was surface applied 1 m from the water inflow in a furrow area 0.2 m wide x 1.0 m long at a depth or 0.2 m deep (Figure 2.10). An

Figure 2.10. Diagram of rill showing irrigation swine waste-water flowing over manure then the PAM treatment (Entry et al. 2003).

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untreated (control) treatment was included. Irrigation water from swine wash-water was pumped to the furrows at a rate of approximately 60.0 L/min. Soils in the field plots were heavy clay, loam and light sandy loam. Entry et al. (2003) found that the PAM + Al2(SO4)3 or PAM + CaO treatments did not consistently reduce NH4

+, NO3-, ortho-P or total P concentrations in

wastewater flowing over a range of soil types compared to inflow wastewater or the control treatment.. Apparently due to the lack of mixing of swine wastewater and the soil treatments, there was little value in applying PAM mixtures in the field to control run-off of ortho-P or total P. However, Entry et al. (2003) suggested that pretreatment of manure with PAMs would decrease vertical movement of total P in susceptible soils. Furthermore, PAM mixtures with mineral acids may be used along with other techniques and management strategies such as riparian vegetation and denitrification walls to reduce the input of pollutants from animal confinement area to water sources.

In addition to the use of acidifying agents like alum, iron salts and mineral acids, another sequestering agent other than the flocculation polymers are zeolites. Zeolite is an alumina-silicate clay mineral widely available in the western US and Mexico. This cation-exchange medium has been used to reduce ammonia in water (Kithome et al. 1998) and in products such as kitty litter to reduce ammonia volatilization and odour from cat urine. Lefcourt and Meisinger (2001) found that zeolite reduced soluble P by 50% in dairy slurry (9 to 11% dry matter). They hypothesize that this reduction was due to adsorption of P on the zeolite mineral or the formation of Ca- or Mg-phosphate mineral intermediates such as octacalcium phosphate or struvite. When alum was reacted with the dairy slurry, soluble P was reduced about 75% compared to the control which is consistent with research with other types of manure.

Another approach to removing phosphorus from swine manure prior to land application is the forced precipitation of magnesium ammonium phosphate hexahydrate (MgNH4PO4·6H2O), commonly called struvite. Struvite frequently forms in the pipes of flush water recycle components of waste management systems causing blockages and restricted flow (Doyle and Parsons 2002). Beal et al. (1999) found that 92% of the soluble reactive phosphate could be removed from the raw liquid swine manure with an addition rate of 0.83 g/L of magnesium oxide (MgO) at 35oC, while 98% of the reactive phosphate was removed with the addition of 1.66 g/L to the digested waste. Similarly, research by Műnch and Barr (2001) and Jaffer et al. (2002) indicated that 94% of ortho-P in digested sewage and 97% of total P in centrifuged sewage liquor, respectively, could be removed by precipitation as struvite. Studies by Burns et al. (2001) and Burns and Moody (2002b) describe a method for precipitation of struvite from liquid hog manure. The results of these two pilot studies indicated that MgO may not be the best source of Mg in struvite formation and that magnesium chloride (MgCl2-6H2O) is more soluble and reduced the reaction time. According to Burns and Moody (2002b), the next step in the development of this technology is the development of a field-scale recovery unit at a commercial animal production unit. A cost-effective magnesium source and a fast, low-cost method of pH adjustment are needed to successfully implement this technology at the farm-scale. The amount of plant available phosphorus which could be released from struvite is unknown. In an earlier pilot study, Schuiling and Andrade (1999) introduced struvite technology to manure treatment plants in the Netherlands as a treatment to reduce the sludge volume and improve the economics of sludge transportation. It was expected that a reduction of 40% of the sludge volume could be achieved and the P-content of the effluent discharged to the sewage treatment plant could be kept below the limit of 30 mg/L. Apparently there are at least 3 potential outlets for struvite: as a secondary phosphate ore (although it has a relatively low P content of 21%), as a source of

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phosphate in other industrial processes (raw material in Mg-phosphate cements in building materials) and as a slow-release fertilizer. It was observed that if struvite is sterilized and dewatered at 120°C to reduce pathogens, it will release its components more easily than unheated struvite (Schuiling and Andrade 1999). Not all of the precipitate that forms during Mg addition is struvite. Brushite [CaPO3(OH)·2H2O] will also precipitate while incorporating P from the manure (Burns and Moody 2002b).

1.17.1.5 Combined treatments Chastain et al. (2001b) experimented with a combination of solid/liquid separation, sedimentation and lagoon treatment (Figure 2.11) as well as a series of settling times and addition of flocculating agents. Results from the on-farm manure treatment system indicated that 53.1% of the total phosphorus from the dairy manure could be removed with the inclined stationary screen. The screen, combined with the settling treatment, resulted in a phosphorus removal rate of 59.9%. Further reduction treatment in the lagoon resulted in an overall removal of phosphorus in the resulting effluent by 86.1% when compared to the nutrient content of the flushed manure entering the total treatment system. Settling of flushed manure (TS = 41.8 g/L) for 60 minutes without the addition of a flocculating agent removed 37.7% of the total P. Addition of 250 to 400 mg PAM/L (cationic polyacrylamide polymer with a 20% charge density) to screened and unscreened dairy manure significantly increased the overall removal of total P (86.0% and 61.8%, respectively) after settling. The optimum amount of PAM to add was 300 mg/L. The addition of alum (liquid aluminum sulfate) was ineffective for settling of unscreened dairy manure. Treating screened manure with 3194 mg alum/L followed by a 60 minute settling period resulted in removal of 99.6% of total P as well as 89.1% total solids, 99.7% of total soluble solids, 74.4% of total Kjeldahl-N and 96% of the COD. The authors judged that the effluent from this system could be stored and recycled through the flush system as wash water (Chastain et al. 2001b).

Figure 2.11. Flow chart of the on-farm combined treatment system. (Chastain et al. 2001b).

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1.17.1.6 Biological phosphorus removal In a process known as biological phosphorus removal, under aerobic conditions bacteria can take up large amounts of phosphate. The phosphate is stored as polyphosphates to be used by the bacteria as an energy reserve in substrate digestion under anaerobic conditions. Regeneration of the phosphate reserve can take place in anoxic as well as under aerobic conditions. It is a cyclical process where the bacteria alternately release and take up phosphate. In order for the bacteria to accumulate large amounts of phosphate (up to 50% of cellular mass) two conditions must exist (Henze 1997):

• alternating anaerobic/aerobic conditions • no nitrate in the anaerobic period.

The anaerobic condition is important to increase the selection pressure in the reactor in favour of phosphorus accumulating bacteria (thought to be largely Proteobacteria with small numbers of Actinobacteria) such that these bacteria form the larger part of the biomass. During the anaerobic period, nitrate has two negative effects on these bacterial species: denitrification removes some of the easily degradable organic matter which was supposed to be stored in the phosphorus accumulating bacteria, thereby reducing the efficiency of phosphorus removal. Secondly, nitrate influences the metabolism of the phosphorus accumulating bacteria so that no polyphosphate is being stored (Henze 1997). The advantages of using a treatment systems such as this is that the treatment is simple, electrical energy useage is low, and supplemental chemical addition is not required to achieve enhance pollutant removal. However variability in real operational conditions (influent flow, load and composition) indicates that there is difficulty in applying models developed under controlled laboratory conditions to full scale biological P removal systems designed for incorporation into animal wastewater treatment systems (Mudaly et al. 2000). Raw piggery wastewater treated in a pilot biological removal system by Ra et al. (2000) achieved removal efficiencies of organic materials, nitrogen and phosphorus of greater than 96, 97 and 95%, respectively. The ortho-P removal by chemical reactions was negligible compared to the biological removal process (Figure 2.12).

Figure 2.12. Phosphorus removal by biological and chemical reactions in aerobic phase of treatment (Ra et al. 2000).

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Using a more complex reactor system (sequencing batch reactor) Obaja et al. (2003) monitored oxidation reduction potential of wastewaters from an anaerobic digester to distinguish the distinct phases of the cycle (anaerobic, aerobic and anoxic). Regulation of the oxidation reduction potential within the reactor allows greater control over the N of the effluents by adapting aeration to the amount of reducing matter present in the sludge. Control over the N is important for efficient P removal. In wastewater containing 1500 mg/L ammonium and 144 mg/L phosphate, a removal efficiency of 99.7% for nitrogen and 97.3% for phosphate was obtained at 30 °C. To retain good removal efficiencies (>95%) the reactor had to be operated at temperatures higher than 16°C. This kind of P removal treatment from dilute wastewater was intended for the treatment and release of municipal sewage. The application of this technology to animal manures could allow the recovery of concentrated nutrients which would be more economical to transport, the reduction of odours and the generation of water of suitable quality for crop irrigation.

1.17.1.7 Aeration systems for phosphorus removal Removal of soluble P by utilizing existing Ca and Fe in the manure without additional chemicals may be of economic and environmental interest. On a laboratory scale, Zhu et al. (2001a,b) were able to raise the pH of manure to 8 using aeration which resulted in the removal of 80% of the soluble P in liquid swine manure. Initially, they observed that during aeration no new soluble P was generated in the liquid. There are three possible mechanisms for pH increase due to aeration: phosphoric acid removal by aerobes, CO2 purging (formation of carbonates would raise the pH) and ammonia production. No difference was found in terms of P removal between continuous and intermittent aeration regimes leading the authors to suggest that intermittent aeration is preferred because of the energy saving advantage (Zhu et al. 2001b; Luo et al. 2002). Closer examination of the transformations of P during continuous and intermittent aeration process revealed that the efficacy of soluble P removal after the first day of aeration could be improved by solid separation of the manure (Luo et al. 2001). As the organic and soluble P decreased, insoluble inorganic P increased due to chemical precipitation of the ortho-P under the conditions of elevated pH. Solid separation reduced the inorganic P levels and prevented the conversion of inorganic P into the soluble P form. This observation of decreased efficiency of soluble P removal from manure with high solids explains why soluble P removal was only 80% in the earlier publications (Zhu et al. 2001a,b).

1.17.1.8 Natural systems of phosphorus removal Although the natural systems of nutrient removal described in the following text have little application in the harsh climate of Manitoba they are worth mentioning for future reference. These studies were conducted on a small or laboratory scale. Research involving the development and operation of wetlands treatment has been performed at small plot to medium farm-size scale.

Garcia et al. (2002) describes a high rate oxidation system for nutrient removal where organic matter removal is achieved by a mutualistic relationship between bacteria and phytoplankton. The oxygen required for aerobic bacterial decomposition of organic matter is provided by photosynthesis, whereas carbon dioxide, ammonia and orthophosphate needed for phytoplankton growth are supplied from bacterial decomposition and from the wastewater. Mixing the contents of the reactor promotes the growth of the phytoplankton and prevents its settling. The biomass

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produced in the reactor is due mainly to phytoplankton which must be subsequently separated to complete the wastewater treatment. Nutrient uptake by phytoplankton and subsequent biomass separation results in nutrient removal. The phytoplankton photosynthetic activity raises pH in the reactor, resulting in ammonia stripping, which is not entirely desirable, and orthophosphate precipitation. During the course of their experimentation, Garcia et al. (2002) observed that environmental factors such as the lowering of solar radiation and temperature in autumn and winter is only marginally compensated by increasing the hydraulic retention time (time the material remains in the reactor) to obtain a significant total P removal. The applicability of this type of removal system would need to be tested for livestock wastes and would likely only be operable under more controlled climatic conditions.

In some climatic regions, production of aquatic plants to recover nutrients from livestock wastewater has promise as an alternative technology to convert the nutrients into potentially useful products and offer an alternative to land application. One particularly useful species seems to be duckweed which is a member of the Lemnaceae family. Duckweed can tolerate high wastewater nutrient levels and is able to take up copious amounts of nutrients, showing a preference for ammonium ions. The duckweed biomass has a high protein content of up to 30% (dry weight) and has been demonstrated to be a highly digestible source of protein for livestock (Leng et al. 1995). Recent research with the L. minor isolate 8627 in North Carolina indicates that while the duckweed grew well on swine lagoon liquid, there was a lag phase caused by nutrient shock loading (Cheng et al. 2002). This observation lead the authors to suggest that wastewater should be diluted to at least 100 mg/L TKN and 50 mg/L total P to maintain rapid growth and nutrient uptake. Once the duckweed mat is established and nutrient levels are lowered in the pond, more concentrated liquid swine manure from an anaerobic lagoon could be introduced.

Recovery of dairy manure nutrients by benthic freshwater algae has also been investigated (Wilkie and Mulbry 2002) using a nutrient film concept to grow and harvest the algae. Although there was significant nutrient uptake by the algae in this study, significant levels of nutrients remained in the effluent. Due to the small scale of this application (a total of less than 2 square meters of growing area) it is difficult to assess the applicability of this technology to a concentrated livestock facility.

Adsorption to soil was identified as a key wastewater P removal mechanism in treatment wetlands in Pictou County, Nova Scotia (Jamieson et al. 2002). In a constructed wetland that had been receiving wastewater from dairy runoff since 1996, Jamieson et al. (2002) determined from calculating the P adsorption maxima, that the expected life span of this particular wetland was only another 8 years. Generally when a wetland has exceeded its absorptive capacity it must be dredged to remove the phosphorus and expose unsaturated soils. Prantner et al. (2001) experimented with a combination treatment of soil infiltration and wetland treatment for liquid hog manure. The soil treatment acted like a sediment trap to remove the solids prior to the effluent entering the wetland treatment area populated by cattails (Typha spp.). The study, though only on a microcosm scale, showed that phosphorus removal levels of 80% were possible with pre-treatment of the liquid swine manure. More recent research by Hunt et al. (2002) and Stone et al. (2002) with various wetland cell combinations indicated that nitrogen was effectively removed at mean monthly loading rates of 3 to 40 kg N/ha/day with nitrogen removal rates as high as 85%. However, these systems had little capacity for phosphorus removal from pretreated swine manure (pretreated in an anaerobic lagoon).

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At the time when West Nile Virus is presenting a potential health threat in all parts of Canada, the wisdom of maintaining wetlands for manure nutrient management is questionable.

1.17.1.9 Solid manure treatments Composting of animal manure results in a 30 to 50% reduction in mass and produces a material more uniform in nutrient composition. Mineralization of labile C fractions results in an enrichment of the P concentrations of the composted material. An imbalance between N and P in the compost can result in a high loading of P when compost applications are made to meet the N demands of crops. High concentrations of soluble P and organic P also may result in offsite discharges. Research by Tiquia et al. (1998) indicated that the solubility of P increased during decomposition of a mixture of pig manure and sawdust bedding. To explain the increase in P solubility, Tiquia et al. (2002b) examined enzyme activities during aerobic composting. They found that alkaline and acid phosphatases reached their maximum activity 14 days after composting began, likely due to aerobic bacterial growth and the subsequent synthesis of phosphatases and other hydrolytic enzymes. In contrast, Dao et al. (2001) found that mass loss from composting resulted in an enrichment of P concentrations, but did not change the solubility of mineral supplements nor the relative concentrations of extractable dissolved reactive P. However, the concentration of dissolved reactive P in composted manure may depend on the resistance of the carbon source to degradation. Larney et al. (2002) found that composted manure/wood chip-bedding had more P in the available form compared to a composted manure/straw bedding mixture.

The solubility of P after soil application of composted manures from different species could also be variable. Gagnon and Simard (1999) observed that P release from soils treated with composted poultry litter was higher than that from composted beef and dairy manures, and fresh beef manure. A similar observation was made by Sharpley and Moyer (2000) who studied P runoff from composted poultry and dairy manures. An explanation for this phenomenon is offered by Cooperband and Good (2002) who observed differing extractabilities of water soluble P on soils receiving either dairy or poultry manure at various levels of soil test P. They found that water-soluble P release did not appear to be linked to manure C decomposition. In poultry manure-amended soil, water-soluble P concentrations appeared to be controlled by a mineral phase of P, while in dairy manure-amended soil, sorption-desorption processes were predominant in controlling soluble P. Calcium and magnesium-P rich minerals were found in the incubated poultry manure but similar minerals were not found in the incubated dairy manure (Cooperband and Good 2002). These mineral-P complexes are referred to as biogenic phosphates. It is possible that they could form in the bird or at environmental temperatures soon after excretion and may include whitlockite, struvite and amorphous Mg-Ca-phosphate. The other possibility is that the excreta from poultry contains large amounts of phytate which may precipitate with Ca-phosphates preventing the crystallization of apatite (a form of Ca-P) of which is more resistant to dissolution (Cooperband and Good 2002). In spite of differences in soil adsorption characteristics, Sharpley and Moyer (2000) suggest that the simple water extractable method for soluble P may provide analytical laboratories with a tool to incorporate information on the properties of P in different manures and composts into land-application guidelines that protect water quality.

Petersen et al. (1998) conducted experiments looking at the effects of simple passive solid manure storage of cattle and pigs over periods of 9-14 weeks. They found no significant losses in

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P while C and N levels were reduced by 48-49 and 48-57% respectively. In contrast Tiquia et al. (2002a) found that runoff P losses during active or passive composting were significant, with losses of 32% and 29.7% of total P, respectively. Although Eghball et al. (1997) found P losses to be relatively small during composting of beef cattle feedlot manure (<2%), they were concerned that runoff and leaching from composting sites may be a significant contributor of N, P, K and Na to surface and ground waters. In an effort to reduce run-off losses of P, Dao (1999) investigated the effect of compost amendments on P extractability. Caliche, alum and fly ash reduced water extractable P by 50, 83 and 93%, respectively, in compost and 21, 60, and 85% in stockpiled manure.

1.17.2 Strategies to Recover and Use Excess Phosphorus

1.17.2.1 Manure as a feed ingredient Although using manure as a feed ingredient is not a popular notion in Canada, new technologies are emerging for preserving wastes and transforming them into new feed ingredients. El Jalil et al. (2001) describe a process for fermentation of poultry wastes by dilution with water, adding 10% molasses and a culture of Lactobacillus plantarum and Pediococcus acidolactici then incubating the mixture at 30°C for 10 days. Results indicated that the product obtained from the wastes had low counts of enterobacteria and enterococci. Nutritional trials indicated that the incorporation of the treated poultry manure at a rate of up to 40% gave laying performance similar to those obtained with conventional diets (El Jalil et al. 2001).

1.17.2.2 Biological tissues as hyperaccumulators of P Long-term applications of manure have led to accumulation of high levels of soil phosphorus in many regions of the country. Nutrient accumulations may occur in part because of over-application of manure but also because the concentrations of N and P in manure are not in proportion to crop requirements. When manure is applied to the land every year based on plant N requirements, there will be a gradual accumulation of total and available P. In livestock production areas where available P accumulation in excess of 200 mg P/kg soil has occurred, decades may be required to deplete available P to agronomic threshold levels of 20 to 24 mg P/kg soil. A review of the literature reveals that plants typically have between 0.1 and 0.9% P (dry matter basis) in stems and leaves and regardless of the total available P in the soil, the plant is restricted in its ability to take up P. Nevertheless, one approach to decrease soil P contents is to cease applying manure to fields and allow crops sufficient time to remove P to a target level. Some plant varieties can hyperaccumulate P (defined as % P contents between 0.8 and 1.45 in dry matter) and use of such plants could be considered as a crop management approach to decrease soil P concentrations (Novak and Chan 2002). It is possible that isolating P nutritional traits from the germplasm of these P-hyperaccumulator plants and using modern transgenic techniques to transfer these traits to common row and forage crops may have a place in P management. However, as previously mentioned in this review, high phosphorus content of animal feedstuffs already presents a challenge in formulating diets for ruminants, therefore increasing the ability of the plant to take up more phosphorus is counter productive.

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1.17.3 Potential Problems Associated with Adoption of Treatment and Utilization Strategies

Although alum (aluminum sulfate) and zeolites (silicate clay minerals) have the potential to reduce ammonia volatilization from liquid manures, Lefcourt and Meisinger (2001) found that alum additions of 2.5% and 6.25% resulted in a significant increase in soluble Al, which if spread on land could affect soil acidity levels. The rapid adoption of the use of alum by European farmers to bind phosphorus in poultry litter, without adequate agronomic studies to investigate the sustainability of the practice, is of great concern. Some zeolites have the capability to permanently sequester phosphorus which is not a desirable outcome if the intention is only to decrease the solubility of P and reduce runoff from land application. Apparently there are many types of zeolites with varying properties so perhaps careful selection of a form of zeolite, appropriate to the manure treatment or handling system, will offer solutions to managing soluble P.

Until recently the earth’s phosphorus cycle was believed to be restricted to phosphorus species which are transported in the liquid or solid phase. In contrast to carbon, nitrogen, sulphur, oxygen and some other elements, phosphorus was not acknowledged as participating in the gaseous atmosphere in its reduced form (phosphine, PH3). Prior to the early seventies, the phosphine occasionally found in the anaerobic biosphere was thought to be of biogenic origin although it was assumed that microorganisms were unable to reduce phosphate to phosphine (Roels and Verstraete 2001). With the advent of gas chromatography, the evolution and sorption of phosphine by soils was confirmed. Recent research shows the presence of phosphine at a large number of locations. Tables 2.13 and 2.14 provide an overview of the concentrations and/or emissions of phosphine in natural and agricultural environments. The half-life time of phosphine in the atmosphere, in the presence of sunlight is only a couple of hours. A major part of reduced phosphorus present in the biosphere is bound to soil, sediments, sludge and manure (Roels et al. 2002). This is referred to as matrix bound phosphine. Interest in the mechanisms leading to phosphine formation has developed as a result of research indicating that free phosphine formed during anaerobic digestion of manure could inhibit the fermentation process (Eismann et al. 1997a). On the other hand, the matrix-bound phosphine is reduced by more than 50% during anaerobic batch-fermentation. In light of this observation, Eismann et al. (1997b) consider that manure may be more of a sink for phosphine rather than a source. However, the close relationship between phosphine in the feed and the manure of the swine, imply that phosphine residues in the feed represent an important source of phosphine which may later affect the functioning of manure treatment technologies (Eismann et al. 1997b).

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Table 2.13. Survey of the presence of free phosphine in the biosphere (adapted from Roels and Verstraete 2001) Origin Concentration Emission Biogas from sediments of sewage plant and shallow lakes

11.6-382 mg/m2 0.09-9.17 mg/m2 h

Brackish marsh 0.42-3.03 ng/m2 h Salt marsh 0.19-6.52 ng/m2 h

Animal Slurry Treatment Biogas 0-295 ng/m3 Putrefaction gas 24-20300 ng/m3

0.041-0.885 ng/m3a 1300 ng/m2 hb Atmospheric phosphine at locations around the world 0-157 ng/m3c

Landfill gas 0-24646 ng/m3 a North Sea b Wadden Sea (assumption) c Germany, Argentina, Tunisia, Seychelles, Israel, Namibia Table 2.14. Free and matrix bound phosphine content in biogas and sludge collected from various sources (adapted from Roels et al. 2002) Free Phosphine Matrix bound phosphine

Type of digestion Concentration (mg/m3) Sludge Type Concentration

(ng/kg) ng/kg DM Manure 42.3 ± 3.5 Manure 22.9 ± 0.8 639 Paper 4.5 ± 1.1 Paper 2.4 ± 0.3 56 Potato <4 Potato nr* <10 Slaughtery waste 179.2 ± 19.5 Slaughtery waste 12.2 ± 1.4 164 Sewage sludge 10.2 ± 0.2 Sewage sludge 101.4 ± 5.8 2373 Landfill 6398 ± 685 Pig manure 71.4 ± 20.3 1245 * nr no reproducible results could be obtained.

1.18 Conclusions The problems of nonpoint pollution from agriculture do not lend themselves to a “one size fits all” solution. Regulators should keep in mind that flexibility is important if goals are to be reached at the least possible cost to farmers and society. Individual farmers may be able to use several incremental changes to reach an overall solution to their particular problems. It is likely that in the near future, manure application will be restricted to the amount of nutrients that can be removed by crops. In making more precise decisions about nutrient management, the livestock industry needs a more integrated approach which involves livestock genetics, housing, herd management, selection of feed ingredients and additives, feed technology, manure treatment and handling, method of nutrient distribution, crop selection, crop management, residue recycling, management of soil nutrient reserves and land management. A narrow focus on one aspect of nutrient management does not well serve a multi-species livestock industry situated across diverse landscapes and eco-regions. Nevertheless, there are promising individual approaches in P

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reduction strategies which can contribute to overall P nutrient management and these deserve the attention and support of the livestock industry.

Over the last few years there has been only been a small effort in Canada to develop feed grains (low-phytate) with improved availability of P to monogastric livestock species. Although low-phytate corn is now commercially available through breeding programs in the USA, low yields and subsequent low export of soil P, are factors in the acceptance and utility of this crop. Similarly, improved agronomic characteristics of low-phytate barley would significantly increase the potential for this crop. Currently, the data supporting improved P digestibility of low-phytate barley in Western Canada is limited to a single trial with a 3-day fecal collection period. Significantly more research is required to assess the true benefit of this feed grain and the forms of P excreted in the feces and urine.

All grain types contain sufficient P to meet the requirements of all livestock species. In fact, the abundance of P in grains creates problems for the beef cattle industry in their efforts to balance rations to obtain optimum calcium: phosphorus ratios. Efficiency of P utilization increases with the addition of phytase to the diets of poultry and pigs. Enhancement of natural phytases in the feed ingredients also has been shown to be beneficial. The addition of phytase to diets comprised of low-phytate grains may offer further improvement in P digestibility and retention. However, the body of literature indicates that there is a wide variability among locations and diets in P digestibility of phytase-amended feedstuffs. There is sufficient evidence that phytase improves the availability of Ca, Zn, protein/amino acids and energy as well as P which may explain some of the variability in observations. The most obvious lack of knowledge is the effect of phytase on solubility of P excreted since there is little information on urinary P excretion from animals fed phytase supplemented diets. Information on P digestibility of amended diets common in Canadian pork production systems would be of value since genetics, climate, local feed ingredients and other management practices affect overall digestibility of nutrients. Augmentation of endogenous phytase in swine may play a role in reduction of P excretion. Further research and development is required to enhance the rate of expression of the trait in litters produced from founder lines and to maintain that trait throughout the life-time of the pig. The solubility of P in the excrement of transgenic pigs has not been described.

The development of comprehensive animal management and feeding programs for each species of livestock would raise awareness of the impact of management decisions on nutrient loss. For those producers who utilize home-grown feedstuffs, the importance of crop production practices, particle size, diet constituents, proper mixing of feeds and feed waste on the efficient use of P by livestock could have a significant impact on nutrient excretion and loss. For more vertically integrated livestock production systems, increasing the number of diets in a phase-feeding system might be beneficial environmentally and economically. Likewise, protection from disease, parasites and environmental elements are also important to increase nutrient efficiency. Increasing the amount of milk or lean meat produced per kg of P excreted is also considered as part of a nutrient management plan. Inclusion of alternative low-P Ca supplements or low-P forages in feedlot diets as well as the development of a P-excretion index, would assist dairy and beef producers in developing targets for P excretion.

When crop and livestock management issues have been addressed, manure treatment strategies offer a means to recover nutrients in a form which reduces nutrient variability, improves handling characteristics, allows distribution of nutrients over a wider geographical area, or produces an off-farm marketable product. Dewatering, composting, nutrient inactivation by

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chemical treatment and natural biological phosphorus removal all have potential to limit the impact of phosphorus on water quality. Treatment would be especially useful in livestock rearing systems where limiting P in the diet is more difficult to achieve. However, the implication of various treatments on the value of nutrients for crop production is not well investigated or understood. Before adoption of treatment plans, the implication of treatment on transformation of P as well as N and the impact of the processed material on soil quality and processes must be elucidated.

1.19 Abbreviations and Definitions Total P is a summation of all phosphorus in a given system.

AGRONOMIC/SOIL TERMINOLOGY:

Acid soluble organic P is inositol hexaphosphates and small amount of phospholipids.

Applied P is the amount of phosphorus applied to land, usually given on a weight of P per hectare basis (kg P/ha).

Available P (orthophosphate, citrate soluble, citrate acid extractable) is inorganic phosphorus that is readily “available” for plant use.

Biological Oxygen Demand (BOD) The amount of oxygen used for biochemical oxidation by a unit volume of water at a given temperature and for a given time. BOD is an index of the degree of organic pollution in water.

Chemical Oxygen Demand (COD) The quantity of oxygen used in biological and non-biological oxidation of materials in water; a measure of water quality.

Citric acid extractable P is phosphorus extracted from fertilizers or amendments with citric acid to provide a measure of availability; citrate soluble = available to plants.

Inorganic P is derived from parent rock during soil formation or is mined by humans; is mostly orthophosphates.

Labile P is the quickly available portion of P that exists in equilibrium between immediately available (inorganic) and very slowly available (organic and precipitated) P where upon a slight change in surrounding conditions (moisture, temperature, or pH, for example) it becomes available to plants.

Nonlabile P is unavailable P that is in organic adsorbed or precipitated forms.

Organic P is formed from gradual uptake and use by plants, animals, and microorganisms converting inorganic P in their cells.

Orthophosphate is commonly used interchangeably, though not quite correctly, with available P. It is an inorganic anion or salt of orthophosphoric acid that contains one P atom (not a group of two or more P atoms which is polyphosphate).

Oxidation Reduction Potential (Redox Potential) a measurement of the ease with which a substance either acquires electrons (reduction) or releases electrons (oxidation). Residual P or Residual organic P is P that is not extracted by preceding treatments.

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Soil test P estimates the amount of P that can be extracted from the soil by plant roots by using acids or bases to cleave all but the most insoluble forms of P, giving an indication of the quantity of P that plants may use during a growing season.

Soluble P or water soluble P is the fraction of P when extracted with water passes through 0.45 µm filters; water or weak acids or salts of weak acids extraction gives an indication of intensity of runoff after a rain. Most commercial fertilizers are readily water-soluble.

Soluble reactive P (SRP) is soluble P (the filtrate, 0.45 µm filter, of water or weak acid extraction) that responds to colorimetry. Usually, this means the extract that reacts with the ascorbic acid and molybdate chemicals in the Murphy-Riley method.

Triple Superphosphate, calcium dihydrogen phosphate, is formed by treating phosphate rock with phosphoric acid to create an inorganic fertilizer containing about 20% P (45% P2O5).

NUTRITION TERMINOLOGY:

Available P is inorganic phosphorus that is readily “available” for use by the animal.

Denaturation a partial physical disruption of the internal structure of a protein molecule inactivating enzymatic activity, usually reversible, caused by extremes in temperature, pH, radiation, salt concentration, also through exposure to organic solvents, stirring or drying. Dicalcium phosphate, in pure form, is a moderately insoluble form of phosphate (CaHPO4); commercial “dical” is commonly used as a source of supplemental dietary P and is not defined chemically by one compound, but is a mixture of varying quantities of dicalcium and monocalcium phosphates, phosphoric acid, calcium carbonate and impurities.

Dietary P is the phosphorus in the diet of livestock and poultry, in both inorganic (from supplements and grains) and organic (the phytate P from grains or forage) forms.

Enzyme is a protein that facilitates biochemical reactions. The physical shape of the protein is important to enzymatic function.

Endogenous derived or originating internally

Exogenous derived or originating externally

In Vitro in an artificial environment outside the living organism

In Vivo within a living organism

Phosphatase any of a group of enzymes that catalyzes the hydrolysis, or breaking up, of organic phosphates

Phytic acid (phytate-bound P, phytic acid, inositol hexaphosphate, inositol hexphosphytic acid, phytin) is an organic phosphorus compound that makes up the majority of phosphorus in plant feedstuffs and is generally unavailable for digestion in nonruminant species.

1.20 References Adeola, O., Cline, T. R., Orban, J. I., Ragland, D. and Sutton, A. L. 1998. Supplementation of Low-Calcium and Low-Phosphorus Diets with Phytase and Cholecalciferol. [Online] Available: http://www.ansc.purdue.edu/swine/swineday/sday98/16.pdf [18 March 2003].

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Adeola, O., Lawrence, B. V., Sutton, A. L. and Cline, T. R. 1995. Phytase-induced changes in mineral utilization in zinc-supplemented diets for pigs. J. Anim. Sci. 73: 3384-3391.

Aggrey, S. E., Zhang, W., Bakalli, R. I., Pesti, G. M. and Edwards, H. M., Jr. 2002. Genetics of phytate phosphorus bio-availability in poultry. Proceedings of the 7th World Congress on Genetics Applied to Livestock Production. August 19-23, 2002, Montpellier, France Abstact No. 10-21. 2002. Angel, R. 1999. Update on practical approached for decreasing phosphorus in poultry litter: summary of current research. [Online] Available: http://www.agnr.umd.edu/AGNRNews/Article.cfm?ID=949 [20 March 2003].

Barnett, G. M. 1994. Phosphorus forms in animal manure. Biores. Technol. 49:139-147.

Barrow, J. T., Van Horn, H. H., Anderson, D. L. and Nordstedt, R. A. 1997. Effects of Fe and Ca additions to dairy wastewaters on solids and nutrient removal by sedimentation. Appl. Eng. in Agric. 13: 259-267. Barvenik, F. W. 1994. Polyacrylamide characteristics related to soil applications. Soil Sci. 158: 235-243.

Beal, J. D., Niven, S. J., Campbell, A. and Brooks, P. H. 2002. The effect of temperature on the growth and persistence of Salmonella in fermented liquid pig feed. Intern. J. Food Microbiol. 79: 99-104.

Beal, L. J., Burns, R. T. and Stalder, K. J. 1999. Effects of Anaerobic Digestion on Struvite Production for Nutrient Removal from Swine Waste Prior to Land Application. 1999 ASAE Annual International Meeting; Toronto, ON.

Beers, S. and Jongbloed, A. W. 1992. Effect of supplementary Aspergillus niger phytase in diets for piglets on their performance and apparent digestibility of phosphorus. Anim. Prod. 55: 425-430.

Beers, S. and Jongbloed, A. W. 1993. Phosphorus digestibility and requirement of pigs. Feed Mix. 1: 28-32.

Blatny, P., Kvasnicka, F. and Kenndler, E. 1995. Determination of phytic acid in cereal grains, legumes, and feeds by capillary isotachophoresis. J. Agric. Food Chem. 43:129-133.

Boling-Frankenbach, S. D., Snow, J. L., Parsons, C. M. and Baker, D. H. 2001. The effect of citric acid on the calcium and phosphorus requirements of chicks fed corn-soybean meal diets. Poultry Sci. 80: 83-788.

Boling, S. D., Webel, D. M., Mavromichalis, I., Parsons, C. M. and Baker, D. H. 2000. The effects of citric acid on phytate-phosphorus utilization in young chicks and pigs. J. Anim. Sci. 78: 682-689. Brinch-Pedersen, H., Olesen, A., Rasmussen, S. K. and Holm, P. B. 2000. Generation of transgenic wheat (Triticum aestivum L.) for constitutive accumulation of an Aspergillus phytase. Mol. Breeding. 6:195-206.

Brinch-Pedersen, H., Sorensen, L. D. and Bach Holm, P. 2002. Engineering crop plants: getting a handle on phosphate. TRENDS Plant Sci. 7: 1-8.

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Burns, R. T. and Moody, L. B. 2002a. Performance testing of screw-press solid separators: comprehensive solids analysis and nutrient partitioning. In: Proceedings of the International Symposium Addressing Animal Production & Environmental Issues. October 3-5, 2002. Raleigh, NC. 4 pp.

Burns, R. T. and Moody, L. B. 2002b. Phosphorus recovery from animal manures using optimized struvite precipitation. In: Proceedings of Coagulants and Flocculants: Global Market and Technical Opportunities for Water Treatment Chemicals. May 22-24, 2002, Chicago, IL.

Burns, R. T., Moody, L. B., Walker, F. R. and Raman, D. R. 2001. Laboratory and in-situ reductions of soluble phosphorus in swine waste slurries. Environ. Technol. 22:1273-1278.

Calvert, C. C., Besecker, R. J., Plumlee, M. P., Cline, T. R. and Forsyth, D. M. 1978. Apparent digestibility of phosphorus in barley and corn for growing swine. J. Anim. Sci. 47: 420-426.

Carlson, D. and Poulsen, H. D. 2003. Phytate degradation in soaked and fermented liquid feed - effect of diet, time of soaking, heat treatment, phytase activity, pH and temperature. Anim. Feed Sci. Technol. 103:141-154. Carter, S.D., Cromwell, G.L., Colombo, G. and Fanti, P. 1999. Effects of porcine somatotropin on calcium and phosphorus balance and markers of bone metabolism in finishing pigs. J. Anim. Sci. 77: 2163-2171.

Challa, J., Braithwaite, G. D. and Dhanoa, M. S. 1989. Phosphorus homeostasis in growing calves. J. Agric. Sci. Camb. 112:217-226.

Chase, L. E. 1998. Phosphorus nutrition of dairy cattle. Proceedings of the Mid-South Ruminant Nutrition Conference, April 1998, Dallas/Fort Worth. [Online] Available: http://www.txanc.org/proceedings/1998/Pnutrition.pdf. [28 January 2003]

Chase, L. E. 2000. Phosporus Nutrition of Dairy Cattle. Proc. Mid-Atlantic Environmental Management Conference.

Chastain, J. P., Lucas, W. D., Albrecht, J. E., Pardue, J. C., Adams, J. and Moore, K. P. 1998. Solids and Nutrient Removal From Liquid Swine Manure Using a Screw Press Separator. ASAE Paper No. 98-4110. St. Joseph, MI: ASAE.

Chastain, J. P., Lucas, W. D., Albercht, J. E., Pardue, J. C., Adams, J. and Moore, K. P. 2001a. Removal of solids and major plant nutrients from swine manure using a screw press separator. Appl. Eng. Agric. 17: 355-363.

Chastain, J. P., Vanotti, M. B. and Wingfield, M. M. 2001b. Effectiveness of liquid-soil separation for treatment of flushed dairy manure: a case study. Appl. Eng. Agric. 17: 343-354.

Cheng, J., Landesman, L., Bergmann, B. A., Classen, J. J., Howard, J. W. and Yamamoto, Y. T. 2002. Nutrient removal from swine lagoon liquid by lemna minor 8627. Trans. ASAE 45: 1003-1010.

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http://www.das.psu/edu/dcn/CATNUT/PDF/PHOS.PDF

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2 Impact of Soil Type, P Management, Landscape and Climate on P Retention and Release by Soil

Dr. Brian Wiebe and Dr. Wole Akinremi Department of Soil Science, University of Manitoba

2.1 Chapter Summary The risk of P loss from soil to water is determined by several factors that include soil P retention properties, management factors, and transport factors. The retention of P by soil occurs largely as adsorption and precipitation reactions. Adsorption occurs quickly but a soil’s sorption capacity is limited. Precipitation occurs more slowly but in general a soil’s capacity to precipitate P compounds increases as P concentration increases.

Fe and Al oxides, clay minerals, and calcium carbonate (CaCO3) are all able to adsorb P; however Fe and Al hold P more strongly and are therefore most important when present in sufficient quantities. Adsorption of P by CaCO3 is a function of the mineral’s surface area and becomes increasingly important at higher concentrations of P. As a result, CaCO3 may be more important in retention of inorganic fertilizer-P than manure-P. Manure may also influence soil P retention directly or indirectly. Depending on the soil, organic acids from manure may increase or decrease soil P adsorption. The cation composition of the manure may also affect P retention. Fe, Al, Ca, Mg, and other cations in the manure can directly form precipitates with P or displace cations from soil exchange sites that in turn react with P. The combination of manure P and cation content and soil composition determine which process will occur.

Most of Manitoba’s agricultural soils were formed from calcareous parent material, and therefore, have alkaline pH values. Although CaCO3 can retain P, its capacity is not unlimited and care must be taken not to overestimate its contribution. Also, due to soil development processes, 65% of Manitoba soils no longer have free CaCO3 in their surface horizon and of those that are calcareous at the surface, less than one-half are fine textured (high surface area). Furthermore, about 25% of Manitoba soils have undergone sufficient leaching to reduce the pH to near neutral values (pH of 6.1 to 7.0 measured in CaCl2 solution). This is the pH range where P sorption is weakest and solubility of P precipitates is highest.

Management can play a large role in influencing the forms and quantities of P loss from agricultural fields. In general, P losses are reduced by practices that keep P applications well below the soil P retention capacity and by practices that minimize contact between manure-P and runoff water. Incorporation of manure results in lower losses of P than surface application in most cases. Incorporation may however increase the risk of erosion and loss of P associated with soil particles. Where surface application is unavoidable (e.g. perennial forage), spring application will in general have lower P losses than fall application as the majority of runoff occurs during spring snowmelt. Practices, such as conservation tillage and riparian or buffer strips, that reduce soil erosion losses will reduce loss of particulate phosphorus (PP) but are less effective in reducing losses of dissolved phosphorus (DP) and may actually increase DP losses.

Export of P from the land requires both a source of P, accessible to runoff water, and a transport pathway for that water to reach surface waters. Unlike most of the regions studying P loss from agricultural areas to surface waters, snowmelt is the major source of runoff on the Canadian Prairies and most of the P export to surface waters occurs with snowmelt runoff. Occasional rain

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storms also produce localized runoff and need to be considered in a comprehensive assessment of the risk of P loss.

Several methods for assessing the risk of P loss to surface waters have been developed and tested. A soil test phosphorus (STP) limit has the advantage of simplicity (a single measurement) but does not take into account the differences in P retention or P buffering capacity among soils. Also, a simple STP limit system also does not include a transport risk factor. The balance of P removal and addition method also has the advantage of simplicity but it would allow a field with excess P to be maintained in that high risk state while requiring a P deficient field to be maintained in its state as well.

The degree of phosphorus saturation (DPS) method incorporates measures of STP and the soil P retention capacity and as a result takes into account the fact that P is more tightly bound (and less likely to be exported by runoff) at low levels of P and that P is bound less tightly (and more likely to be exported by runoff) as the soil retention capacity becomes saturated. Although an improvement over a STP method, the DPS still does not account for the risk of transport from the field.

Phosphorus transport models such as FHANTM have been useful in predicting P export for the region where they were developed. However such models require a great number of input variables and although some can be estimated the reliability of the model then depends on the estimate. The large number of climate, soil, and management variables needed, makes this approach impractical for Manitoba at this time.

The phosphorus index (P Index) is a method that attempts to combine source and transport factors into a simple risk index. Different factors are assigned weighting factors and different levels of each factor are assigned risk values. Unlike the transport models, the P index does not attempt to predict how much P will be exported under specific conditions but rather it rates field locations and practices by their potential to result in P export. As a result a P-index requires fewer input variables but still accounts for both source and transport factors. The P Index method is the most robust and widely used method of risk assessment at this time but more research and resources will be required to develop an appropriate P Index for Manitoba.

The development of an effective P Index for Manitoba will require further research in some key areas. Further study is needed on the role of CaCO3 in the retention of P applied as manure and the effect of manure composition (especially the cation and anion contents since it is affected by diet and livestock class) on P retention by soil. The fate of P added in organic amendments is not well understood in comparison to the retention reactions of inorganic P fertilizers. The effects on soil P retention of organic acids, the cation and anion composition of the manure, pH changes upon manure addition, and the forms of P and their concentrations in the manure require further research. The P sorption capacity (PSC) of Manitoba soils must be determined so that the DPS can be calculated. As yet there is no standard method for determining PSC and DPS in neutral and alkaline soils. Research in these areas is fundamental for properly rating the P source risk factors. P transport by rainfall runoff has been the focus of a large amount of research and modelling. P transport by snowmelt runoff has received less attention and research is needs to be conducted to determine what soil and landscape properties are important in determining P export by snowmelt.

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2.2 Introduction Many surface water bodies (rivers and lakes) are phosphorus (P) limited, i.e. the level of available P limits algal growth. P loss from land to surface waters is a natural process required to supply nutrients to the plants and algae which grow in our surface waters and form a crucial part of the aquatic food chain. As the supply of P is increased above what is needed, the extra nutrients cause increased algal growth. When the algae die and decompose, oxygen depletion occurs and if the depletion is great enough, fish kills are the result.

The P that enters surface waters in Manitoba can originate from erosion of stream- and riverbanks (P occurs naturally in our soils), runoff from forest and grasslands, wildlife droppings, industrial activities, urban areas, and agricultural activities. Measurable amounts of P have also been detected in precipitation.

Agricultural sources for P loading to surface waters include fertilizers, livestock manures, and crop residues. P from livestock manure may reach surface waters via runoff from pastures, wintering areas, feedlots, manure storages, and manured forage and cultivated fields. Livestock numbers continue to increase in Manitoba and as a result the quantities of manure that are land applied also increases each year. From 1991 to 2001, the total number of animal units in Manitoba has increased from 1 021 000 to 1 363 000 (Beaulieu and Bédard 2003). Of this increase 228 000 was due to beef cattle, 87 000 was due to hogs, 9 000 was due to poultry, and 41 000 was due to other livestock (includes all other classes of livestock except dairy cattle which decreased by 26 000 animal units over the same time period). Not only has the number of animals increased, but the distribution has also changed. In 1991, 92% of Manitoba livestock was in low density areas (<25 animal units km-2) and the remainder in medium density areas (25-70 animal units km-2). By 2001, 8.4% of livestock was in high density areas (>70 animal units km-2), 9.3% in medium density areas, and 82.3% in low density areas. As livestock density increases, available land for manure application becomes limited within economic spreading distances and over application of manure nutrients is likely to occur. Two Manitoba rural municipalities, La Broquerie and Hanover, were ranked sixth and eleventh for livestock density in Canada in 2001 with 129 and 106 animal units km-2 respectively.

Manitoba soils are generally considered to be dry (moisture deficit for most crops in most years) and to have high calcium carbonate contents. This may have given a false sense of security as most of the research on P dynamics was done with inorganic P sources not manure. More recent research indicates that manure P may react quite differently in soils than P from inorganic P fertilizer. In order to minimize the loss of P from agricultural land, it is important to know how P reacts with soils taking into consideration the unique soil properties in Manitoba. Also the tools that are available for indicating the risk of P loss from soil must be evaluated under Manitoba’s soil and weather conditions if they are to be effective tools for managing phosphorus in Manitoba and to stem potential eutrophication of Lake Winnipeg and Lake Manitoba.

In this chapter we will review the current knowledge on how P reacts with soil constituents and how the predominant reactions vary among soils. The effect of manure on the soil-P reactions will be discussed as well as how management practices can influence the loss of P from fields receiving manure. The pathways by which P can be exported from land to surface waters and several methods of assessing the risk of P export are described. A section summarizing the dominant soil properties across the main agricultural regions of Manitoba will also be included.

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2.3 Reactions of phosphorus in soil Phosphorus (P) invariably occurs naturally in the form of phosphate salts and minerals (Taylor and Kilmer 1980). The particular salts and minerals that dominate in a soil are a function of both the properties of the soil and the form in which the phosphorus is added to the soil (Sample et al. 1980). Due to additions and removals of P from soils, transport of P within soils, and the slow rate of precipitation and dissolution of some P-compounds, the soil P pool is always in a state of flux. These processes can be grouped into several categories with phosphate moving between them via the soil solution (Figure 3.1). The P released into soil solution through weathering of soil minerals and desorption from soil particles is supplemented by P added to the soil as commercial fertilizer, through rainfall and snowfall, and released from animal manures, plant residues, and agricultural, municipal, and industrial wastes or by-products. P is lost from the soil solution through adsorption, precipitation, immobilization, plant uptake, leaching, and runoff. Erosion also removes soil particles (and P) from the soil system.

Figure 3.1 Soil phosphorus pools and processes.

2.3.1 Adsorption and Desorption Reactions The phenomenon of concentration of liquid or gaseous materials on the surface of a solid is known as adsorption (Larsen 1967). Sposito (1989) defines adsorption as the net accumulation of matter at the interface between a solid phase and an aqueous solution phase that differs from

Soil Solution P

PlantUptake

Mineralizationof Plant

Residue P

Runoff P: overland and subsurface lateral flow may contain P

from any or all pools

Groundwater P:sorption and precipitationmay occur

Matrix andPreferential Flow

Sorbed P:Fe & Al oxides

CaCOclay minerals

3

Precipitated P:Ca-PFe-PAl-P

Mg-P

Desorption

Precipitation

Dissolution

Manure P Fertilizer P

Organic P

Mineralization

Immobilization

Vegetative PExport of P:

removal with livestock and harvested produce

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precipitation because it does not include the development of a three-dimensional molecular structure. Adsorption is a surface phenomenon and the matter which accumulates in the two-dimensional molecular arrangements at an interface is the adsorbate. The solid surface on which it accumulates is the adsorbent. In contrast absorption denotes the uptake of matter or energy by a substance (absorption of water or nutrients into a plant root for example) (SSSA 1997). The term sorption includes both adsorption and absorption and is often used when the exact nature of the mechanism of removal is not known (SSSA 1997).

The soil as a whole is made up of particles that generally have a net negative charge on their surfaces. Nevertheless, soil particles are still capable of strongly retaining anions such as P. This process of retention by the soil is referred to as adsorption. Much of the knowledge gained about the adsorption process in soil has been chiefly from studies involving the fate of P fertilizers in the soil (Motts 1981). The adsorption process involves two interacting elements, the adsorbate (e.g. P) and the adsorbent (e.g. the soil) and thus, the properties of these two components govern the adsorption process.

According to Motts (1981), anions can be classified into 2 groups: (1) non-specifically adsorbed ions; and (2) specifically adsorbed ions. The non-specifically adsorbed ions (e.g. Cl- and NO3

-) are those that are retained on positive sites by simple electrostatic attraction following anion exchange. These groups of ions are assumed to obey the diffuse layer equilibrium (Arnold 1978) and are easily displaced or exchanged. The second group consists of anions that have a far greater affinity for soil surfaces than their concentration in solution would suggest. In the adsorption of these ions, simple electrostatic attraction gives way to the formation of chemical ionic and/or covalent bonds with the surface groups. The phosphate ion belongs to this group, the specifically adsorbed ions.

In soils the main adsorbents of phosphates are iron (Fe) and aluminium (Al) oxides and hydrous oxides, alumino-silicate clay minerals, carbonates, and soil organic matter. In the soil solution, H2PO4

- and HPO4-2 are the most common forms of phosphorus that sorb to soil adsorbents with

H2PO4- dominating at low pH.

Fe and Al oxides exist in soil as discrete particles, surface layers/coatings on soil particles, and in clay minerals as gibbsite (Al oxide layer in silicate clay). Fe oxide surfaces are hydroxylated in the presence of water, either single hydroxyls (Fe-OH) or double hydroxyls (OH-Fe-OH). It is this chemisorbed water that controls the reactivity of the Fe and Al oxides. The surface charge is determined by the solution pH that controls the adsorption and desorption of H+ ions in the chemisorbed layer. At low pH, Fe-OH2

+ dominates (positive surface charge). As pH rises, H+ is lost from –OH groups and eventually the net surface charge is zero (Iso-electric point or point of zero charge, PZC). For various Fe-oxides the PZC varies from pH 7.5 to 9.3 therefore Fe oxides will have a net positive surface charge in most soils but some of them will have a net negative charge in alkaline soils. Al-oxides show a similar change in surface charge with pH.

The charged oxide surface can hold HPO4- via “non-specific adsorption” (outer sphere

complexes) which is easily reversible as the ions are held electrostatically by the charged –OH2+

surface groups. Specific adsorption or inner sphere complexes are much stronger and form when ions penetrate the coordination shell of the Fe (or Al) atom and exchange (or displace) –OH, -OH2, or H+ and are bound by covalent bonds. In monodentate bonding, the anion (e.g. HPO4

-2) replaces an aquo group (-OH2) and bonds directly to the Al or Fe atom (aquo groups will be replaced preferentially to hydroxyls). In bidentate bonding, the HPO4

-2 replaces two hydroxyls

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resulting in two bonds to an Fe or Al atom or a bond to two adjacent atoms. In bidentate bonding H2PO4

- will lose a proton during bonding to allow for the formation of two bonds. Soil pH affects specific adsorption by controlling the ionic species of the adsorbate (P) and by affecting the surface charge of the adsorbent. Phosphate will adsorb in the entire pH range of soils (more below pH 5 than above 7 –i.e. more as H2PO4

- than as HPO4-2, more below PZC than above). In

comparison sulphate will not adsorb above pH 6 (only on the acid side of the PZC of the oxide, only replaces aquo groups), while silicate adsorption increases as pH increases and in alkaline soils can decrease phosphate retention due to competition for binding sites. Using anion exchange resin, Taylor and Ellis (1978) concluded that at low concentration, P was bonded by two points of attachment (bidentate bonding) after deprotonation of the H2PO4

- ion. This was followed by one point attachment (monodentate bonding) at high P concentrations during adsorption on resin surface. This resulted in a deviation of the adsorption plot from that predicted by the Langmuir equation (see section 3.3.1.2) due to the increase in potential adsorption sites as P concentration increases.

Electrostatic binding results in anions that are easily exchangeable and will be released if soil solution composition changes, therefore anions held by electrostatic binding are highly bioavailable. Monodentate binding is stronger than electrostatic binding but anions held by monodentate binding are still exchangeable and therefore also bioavailable. Bidentate bonding is the strongest and is largely irreversible under normal soil conditions. The edge faces of alumino-silicate minerals (clay minerals) like kaolinite contain exposed hydroxyl groups attached to Al atoms which have been shown to adsorb P in a similar fashion to that of the Fe and Al oxides (Hingston et al. 1972).

Adsorption isotherms are used to summarize or characterize P sorption in soils. As such, adsorption isotherms are the basic tool for comparing the effects of various amendments, conditions, and treatments on the retention of P by soils. An adsorption isotherm is defined as the relationship between the quantity of the substance adsorbed by an adsorbent and the equilibrium concentration of the adsorbate at constant temperature. The usefulness of adsorption isotherms in describing phosphorus adsorption is indicated by its utilization in one form or another by many researchers dealing with P adsorption (Taylor and Ellis 1978; Mead 1981; Polyzopoulos et al. 1985; Mehadi and Taylor 1988). Examples of adsorption isotherms for six Manitoba soils are shown in Figure 3.2.

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Calcareous Non-calcareous (a) Libau, (a) Portage CL, (b) McCreary, (b) Pipestone, (c) Graysville II (c) Graysville I

Figure 3.2. Phosphorus sorption isotherms for three calcareous and three non calcareous Manitoba soils (Akinremi 1990). The most widely used equations to describe adsorption isotherms are the Freundlich, Langmuir, Gunary, Temkin, BET and the polynomial equations. Although adsorption isotherm equations are useful in characterizing P retention by soils, they cannot be interpreted to indicate a particular adsorption mechanism or even if adsorption rather than precipitation has occurred (Sposito 1989).

2.3.1.1 The Freundlich isotherm The Freundlich adsorption isotherm has been used to characterize phosphorus adsorption in soil (Olsen and Watanabe 1957). It is an empirical formulation, as such, it is applicable to a wide range of equilibrium phosphorus concentrations. The Freundlich equation may be written as

q = kCb [3.1]

where q is the amount of P adsorbed per unit mass of adsorbent (µg/g), C is the equilibrium concentration of the adsorbate (µM), k and b are constants.

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A logarithmic transformation of this equation gives a straight line represented by

log q = log k + blog C [3.2] A plot of log q against log C should give a straight line with log k as the intercept and b as the slope if the sorption data conforms to the Freundlich isotherm. It should be noted that for q = C, a log transformation would similarly give a straight line even though the original equation did not conform to Freundlich isotherm.

2.3.1.2 The Langmuir isotherm In contrast to the Freundlich isotherm, the Langmuir adsorption isotherm was formulated from the kinetic theory of gases and was used extensively to describe gas adsorption on solids. The isotherm has been found to be applicable to phosphorus adsorption under more dilute equilibrium phosphorus concentration. Olsen and Watanabe (1957) reported that the adsorption of phosphorus by soils from dilute solutions showed a closer agreement with the Langmuir isotherm than the Freundlich isotherm.

Apart from the mode of derivation, the popularity of the Langmuir isotherm stems from the ability to generate adsorption parameters which can be of practical significance (Fried and Shapiro 1956; Olsen and Watanabe 1957). According to these earlier workers, the major advantage of the Langmuir equation over the Freundlich is that an adsorption maximum can be calculated. This parameter can be related to various soil properties and supply information about the nature of the reaction between the soil and fertilizer phosphorus.

The Langmuir equation is written as:

q = kbC/(1+kC) [3.3] where q = amount adsorbed per unit mass of adsorbent (µg/g)

C = equilibrium concentration of the adsorbate (µM) b and k are constants b is the capacity factor referred to as the "adsorption maximum" and k is the "affinity"

factor, that reflects the relative rates of adsorption and desorption at equilibrium (Barrow 1978).

Conformity with the Langmuir isotherm is verified by a plot of the experimental data to the linearized form of equation [3.3]. It is worth noting that equation [3.3] can be linearized in several ways each giving slightly different parameters when applied to the same adsorption data (Dowd and Rigs 1965). The most common or universal form is given by

C/q = 1/kb + C/b [3.4] If the data conforms to the Langmuir isotherm, a plot of C/q against C will yield a straight line whose slope is 1/b and intercept is 1/kb. A combination of these regression coefficients allows the calculation of b and k.

In spite of the popularity of the Langmuir isotherm for describing P sorption in the soil, it has repeatedly been criticized on the basis of its limitations. Barrow (1978) was of the opinion that the Langmuir isotherm is seldom applicable to sorption of P in the soil because it assumed constant energy of adsorption which contradicted observable pattern in the soil. Posner and

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Bowden (1980), stated that the assumption of no interaction between the adsorbed species is not realistic when the adsorbing species are charged and adsorption changes the charge on the adsorbing surface. They suggested that except under restricted circumstances neither the Langmuir isotherm nor a combination thereof can give a realistic description of the adsorption process.

Conformity of sorption data with the Langmuir adsorption isotherm is not indicative of an adsorption process as was demonstrated by Veith and Sposito (1977). These authors concluded that the Langmuir equation cannot be used statistically to determine whether adsorption or precipitation is occurring during fixation reaction in soils.

2.3.1.3 Gunary isotherm The Gunary adsorption isotherm is a modification of the Langmuir adsorption equation. It was formulated as a result of the failure of the Langmuir equation to adequately describe the data generated in 24 surface soils (Gunary 1970). The equation is given by:

C/q = A + BC + D/C [3.5] where A, B and D are constant coefficients and C and q are solution P and adsorbed P concentration, respectively.

Though there was no theoretical basis for this modification, the new model gave a better fit in 24 soils accounting for more than 99.8% of the variation in phosphate sorption (Gunary 1970). The presence of the square root term was interpreted to imply that the soil will adsorb a little phosphate firmly, a slightly greater amount of P less firmly, and so on until a limiting value is reached.

Other adsorption isotherms have been used to describe P sorption in the soil. The list includes the Temkin (Mead 1981; Polyzopoulos et al. 1985), the BET (Taylor and Ellis 1978; Mehadi and Taylor 1988); and the polynomial equation (Hater and Foster 1976).

2.3.1.4 Kinetics of adsorption The adsorption isotherms described above were formulated with the implicit assumption that the sorption reactions are instantaneous and the system itself is at equilibrium. Reported sorption studies indicate otherwise, as the influence of equilibration time is commonly observed (Figure 3.3). Phosphorus sorption is typically rapid during the first few hours or days of incubation, this is then followed by a much slower reaction (Rajan and Fox 1972; Ryden et al. 1977; Enfield and Ellis 1983). The initial fast reaction is considered to be primarily due to adsorption while precipitation, to relatively insoluble forms, controls the slow reaction (Rennie and McKercher 1959; Munns and Fox 1976; Sawhney 1977).

Unlike the equilibrium approach, the kinetic aspect of P sorption is not well developed. Several approaches have been utilized in the kinetic study of P sorption, but most of the proposed equations have a limited ability to describe P reaction with the soil (Enfield et al. 1981). The approaches range from empirical descriptions of sorption data (Enfield 1974; Hater and Foster 1976) to theoretical approximation (Griffin and Jurinak 1973a).

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Figure 3.3. Phosphorus adsorption with time in calcareous and non-calcareous soils: (a) initial P = 20 µg P/mL; (b) initial P = 100 µg P/mL; (c) initial P = 500 µg P/mL

(Akinremi 1990).

From the theoretical viewpoint, when the fraction of the surface exposed was used as the only factor governing the rate of adsorption, a first order kinetic equation was obtained (Haque et al. 1968). Ignoring the fraction of the surface exposed and considering concentration as the only rate determining factor (or vice versa), leads to a pseudo first-order reaction (Peterson and Kwei 1961). Kuo and Lotse (1972) utilized a second order kinetic equation which considered the simultaneous change in phosphate concentration and the change in surface unsaturation due to adsorption. Equilibrium condition was described by a Langmuir isotherm. Griffin and Jurinak (1974) reported that the kinetics of phosphate interaction with calcite could be described by two simultaneous reactions. The first reaction was second order ascribed to the adsorption of phosphate on the calcite surface. The second reaction was first-order and was considered to be

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associated with the surface arrangement of phosphate clusters into calcium phosphate heteronuclei (precipitation). The two processes occur simultaneously and were fitted to a two-surface Langmuir isotherm equation. The break in the slope of the Langmuir plot occurred at a P concentration level that indicated hydroxyapatite (HA) was the precipitate being forming. Others have found octacalcium phosphate (OCP) as the predominant precipitate on calcite surfaces (Webber and Mattingly 1970). Griffin and Jurinak (1973b) showed that the specific Ca-P which formed was a function of the Ca/P ratio.

Enfield et al. (1976) made a comparison of 5 kinetic models for describing P sorption in 25 mineral soils. The mean correlation coefficients from the 5 models were 0.81, 0.83, 0.84, 0.86 and 0.86, respectively. Best fit was obtained when a diffusion-limited model was combined with either the Langmuir or Freundlich equation.

2.3.1.5 Effect of organic matter on sorption and desorption In soil, organic matter may increase or decrease the ability of soils to adsorb P. Soil organic matter is normally negatively charged but in association with common cations, such as Fe3+, Al3+, or Ca2+, organic matter can sorb P (Weir and Soper 1963; Appelt et al. 1975; Holford and Mattingly 1975). Moreno et al. (1960) found that organic matter increased P concentration in solution possibly by complexing Ca ions. Organic acids present in the soil may also compete with P for sorption sites thereby reducing P sorption (Nagarajah et al. 1970; Ohno and Crannell 1996) or may weaken the strength of the bond (Weir and Soper 1963). Organic C was released into solution during P sorption suggesting that the organic acids and P were in competition for the same sorption sites (Erich et al. 2002).

Effect of exchangeable cations on sorption

The effect of the exchangeable cation on P sorption was demonstrated by Bar-Yosef et al. (1988). The authors reported that the adsorption of phosphorus by a Ca-clay exceeded that by K-clay. They attributed this to the effect of the different cations on the extent of the diffuse double layer, which determines the accessibility of P to adsorption sites on clay edges.

Sorption versus precipitation

Calcium carbonate (calcite) has been shown to adsorb P from solution (Green et al. 1978; Griffin and Jurinak 1973, 1974) but the P adsorbed on calcium carbonate surfaces is held much less tightly than the P adsorbed on Fe oxides (Holford and Mattingly 1975). Even at low concentrations of P, P reacts with CaCO3 to form Ca-P precipitates (Sample et al. 1980); as such, it is difficult to separate the two processes. This problem arises because of the ability of the phosphate ion to participate in both adsorption and precipitation reactions (Hsu and Rennie 1962; Larsen 1967; Veith and Sposito 1977). Sometimes, both reactions occur simultaneously, either at different points in the soil or on the same surface (Sample et al. 1980; Talibudeen 1981; Motts 1981) as is the case with calcite (Griffin and Jurinak 1974).

Arguments have been made for and against adsorption and precipitation in the literature. The general consensus is that at "low" solution P, adsorption will be the dominant process, while at "high" P levels, precipitation predominates. This distinction has allowed investigators to impose experimental conditions that often guaranteed that the results are best explained as either adsorption or precipitation (Sample et al. 1980). This distinction is, however, only qualitative as no threshold level of P has been established in the literature for either of these processes. Such

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threshold values would be specific to each soil in view of the complexities and differences among soils. The term P retention is therefore a better term than P sorption for grouping the many processes that remove P from the soil solution and will be used in this review wherever a specific process is not clearly identified.

2.3.2 Precipitation and Dissolution Processes Precipitation is a process that denotes the formation of discrete, solid materials and the controlling mineral(s) in the precipitation reactions is highly pH dependent (Campbell and Edwards 2001). In high pH soils, P can combine with Ca or Mg to form Ca or Mg phosphates. At lower pH, P combines instead with Fe and Al. Therefore the amount of P potentially precipitated is a function of the relative abundance of Ca, Mg, Fe, and Al in the soil solution as well as the pH of the solution.

The addition of fertilizer P as granules or droplets produces a zone of soil solution near saturation with respect to the P-salt added (Sample et al. 1980). The saturated zone slowly spreads out from the granule/droplet and depending on the compound that was added, the pH of this zone can range from 1 to 10. This concentrated solution causes the dissolution of soil constituents, such as oxides, silicate clays, soil carbonates and soil organic matter, releasing large quantities of cations into the soil solution. The resultant solution is supersaturated relative to a variety of P compounds that can then precipitate. The soil solution around these new precipitates changes with time and they may go through a further process of dissolution followed by precipitation as other compounds. With time the precipitated products become less and less soluble and are therefore less able to supply P to the soil solution and to the plants growing in the soil.

The precipitates that form are determined by the composition of the soil solution which is a function of the form of the P fertilizer added and the soil composition. Lindsay et al. (1962) added various P fertilizers to both acid (pH 4.9) and calcareous soils (pH 8.3-8.5). Addition of mono-calcium phosphate (MCP) that has a saturated solution pH of 1.48 resulted in the dissolution of oxides and clay minerals and the formation of Al-P and Fe-P minerals in both acid and calcareous soils. The addition of mono-ammonium phosphate (MAP) (saturated solution pH of 3.47) formed taranakite (an Al phosphate) in the acid soil and Ca and Mg phosphates in the calcareous soil. Only trace levels of Al were detected in the soil solution of the calcareous soil indicating that alumino-silicate minerals were not dissolved to the same extent by MAP as by MCP. Similarly the addition of MKP (mono-potassium phosphate) or MAP to Manitoba soils that contained calcium and/or magnesium carbonate or substantial quantities of exchangeable Ca resulted in the production of mainly DCPD (dicalcium phosphate dehydrate) as well as Mg phosphates when adequate Mg was present but no Al- or Fe-P precipitates were detected by X-ray diffraction analysis (Racz and Soper 1967). The addition of P fertilizers which produce alkaline solutions (diammonium phosphate, DAP, and dipotassium phosphate, DKP) followed a similar pattern. Al and Fe were solubilized in the acid soil resulting in the formation of Al phosphates (different compounds than those formed at lower pH) and Ca and Mg phosphates formed in the calcareous soil. The quantity of precipitates was much lower than that seen with the acidic P sources.

Dissolution of soil minerals is not the only source of cations that will form precipitates with P. Akinremi and Cho (1991a, 1991b) demonstrated that the cations added along with P in the fertilizer (K in their case) could displace sufficient Ca from the soil exchange to form Ca-P

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precipitates. Mobility (and hence leaching) of P was reduced due to the formation of Ca-P. The effect was reduced in columns with lower CEC (less Ca available to be displaced into solution) and was not detectable when the exchange was saturated with Na rather than Ca.

In practice, the adsorption isotherm for a particular soil will represent both adsorption and precipitation reactions. Some of the common soil P precipitates such as OCP or HA form slowly and as a result, temperature and reaction time will affect the amount of P sorption (Griffin and Jurinak 1974). The effect of time on P sorption is demonstrated in Figure 3.3. Similarly desorption isotherms will include both desorption and dissolution.

Generally, desorption isotherms do not follow the same shape as adsorption isotherms for the same soil (Figure 3.4) due to irreversible sorption (Campbell and Edwards 2001). Sorption-desorption hysteresis resulted in replacing EPC0 (equilibrium P concentration at zero P sorption and desorption) determined from sorption isotherms with DPC0 (equilibrium P concentration at zero P desorption) calculated from desorption isotherms to estimate the vulnerability of manure amended soils to P loss in Northern Ireland (Anderson and Wu 2001). Although the desorption trend line has been extrapolated to 0 the last data point indicates greater than 200 mg P/kg is still retained in equilibrium with a solution concentration of less than 2 mg P/L and for practical purposes this P is often considered to be irreversibly sorbed (Campbell and Edwards 2001) or irreversibly retained. The term equilibrium concentration is not a true equilibrium as the desorption isotherm includes dissolution which may proceed very slowly. The effect of time on the amount of P released to solution (desorption and dissolution) is shown in Figure 3.5.

Figure 3.4 Adsorption and desorption isotherms for soil which had been treated with cattle manure (Anderson and Wu 2001). HCOWads = adsorption for soil amended with high rate of cow manure, and HCOWdes = desorption isotherm for soil amended with high rate of cow manure.

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Figure 3.5. Effect of time on desorption of P from soil amended with inorganic P fertilizer (FERT) or hog manure (LPIG = low rate, MPIG = medium rate, HPIG = high rate) (Anderson and Wu 2001).

The low rate of hog manure slurry supplied less P than the fertilizer treatment (14 to 32 kg P/ha for hog slurry depending on the year, 32 kg P/ha each year for the fertilizer treatment). Despite the similarity in P application rate, the low rate of hog slurry resulted in P desorption quantities approximately double that of the fertilized treatment at each time (Fig. 3.5). Differences in reactions between inorganic and organic P soil amendments are discussed in more detail in section 3.3.6.

2.3.3 Mineralization and Immobilization Mineralization is defined as the conversion of an element from an organic form to an inorganic state as a result of microbial activity (SSSA 1997). Immobilization is the conversion of an element from the inorganic to the organic form in microbial or plant tissues (SSSA 1997). These are two opposing processes that occur continuously and simultaneously in soils (Campbell and Edwards 2001). The term net mineralization is used for the sum total of the two processes and can be either positive (mineralization is dominant), negative (immobilization dominates), or near zero (the two processes are in equilibrium). Net mineralization of P can vary widely as it is dependent upon both the relative ratios of C, N, and P in the soil (or organic amendment) and upon the form of P in the amendment.

A C:P ratio of less than 200 favours net mineralization of P and ratios above 300 favour net immobilization. The addition of wheat straw (C:P of 1:320 to 1:880, Douglas and Albrecht 2000) will result in net immobilization whereas the addition of manure (C:P about 10) will result in net mineralization. Of course, the addition of inorganic P fertilizer such as MAP along with a C source such as wheat straw will result in some of the added P being immobilized. Moisture and temperature influence the respective processes differently and hence can effect net

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mineralization. Different types of organic amendments (biosolid, hog manure, and cattle manure) also have different net P mineralization rates when incubated with soil under equivalent moisture and temperature conditions (Kashem et al. 2003).

2.3.4 Plant Uptake Crop uptake of P is a function of the crop type, the available P level in the soil, and the growing season conditions. Phosphorus uptake of several crops is shown in Table 3.1 and ranges from 9 to 22.3 kg P/ha with 2 to 7 kg P/ha returned to the soil as crop residue.

Table 3.1. Uptake and harvest removal of phosphorus by several crops

Crop

Yield kg/ha

Total uptake

kg/ha

Harvested portion kg/ha

Reference

Potato 33 300 22.3 18.5 Lorenz and Vittum 1980 Corn 5 000 21 15 Hanway and Olson 1980 soybean 1 800 19 13 Hanway and Olson 1980 wheat 2 400 11 9 Hanway and Olson 1980 Oats 1 600 11 7 Hanway and Olson 1980 barley 1 900 9 7 Hanway and Olson 1980

Wheat grown in Manitoba may have an above ground P accumulation at maturity of 4-19 kg/ha of which 73-91% is contained in the grain (Tomasiewicz 2000). Actual values at harvest ranged from 2700 to 4150 mg P/kg plant material and the harvest index varied from 73.2% for high P soils to 91.4% in a very P deficient situation. The straw also showed a wide range in P contents with a low of 0.72 kg/ha for a P deficient soil to 5 kg/ha for a high P soil.

Plant uptake of P is also affected by crop variety, availability of other nutrients, moisture conditions and temperature (Hanway and Olson 1980). Varietal differences in rooting habit and activity as well as differences in yield potential are suggested as reasons for differences in P uptake. N deficiencies can reduce P uptake and other nutrients such as Ca and Zn can interfere with P uptake. Soil moisture content, especially of the surface horizon, can have a major impact on P uptake. Depletion of soil moisture in the high P surface soil can result in soil P becoming inaccessible to the crop even though it has access to sufficient water from the subsoil to prevent moisture stress. Low soil temperature will also reduce the plants ability to absorb P from the soil and hence will reduce crop P uptake (most common early in the growing season in Manitoba).

2.3.5 Soil Factors Affecting Phosphorus Retention and Release

2.3.5.1 Soil texture The texture of a soil refers to the relative proportions of sand, silt and clay in the mineral fraction of a soil. As the proportions of silt and especially clay increase relative to sand the surface area of the mineral particles expressed per unit mass (or volume) will increase. Fine colloidal clay has about 10 000 times greater surface area than the same mass of medium-sized sand (Brady 1990).

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The colloidal clay surface area ranges from about 10 to 1000 square meters per gram. P adsorption and often precipitation are processes that occur on surfaces therefore it is not surprising that texture has a large effect on P retention characteristics of a soil. In early P sorption studies Olsen and Watanabe (1957) showed that the Langmuir sorption maximum for P correlated closely with surface area (their study soils had clay contents ranging from 3.5 to 37%).

Nine CaCO3 samples with surface areas ranging 250 to 12 780 m2/kg were studied for their ability to sorb P and it was found that 92.8% of the variation in P sorption could be explained on the basis of differences in surface area (Amer et al. 1985). Multiple regression analysis of P sorption and several soil characteristics of 20 calcareous soils from Lebanon with a wide range in clay, CaCO3, Fe-oxide, and organic carbon contents showed that P sorption increased as clay content and oxalate iron content increased and decreased as CEC increased (Ryan et al. 1985). They concluded that when present, Fe oxide effects on P sorption and desorption are dominant and mask the effects of CaCO3 due to their greater affinity for P. This clearly shows that not only is the amount of surface area important for the retention of P but the kind of surface as well. The effect of CEC was attributed to its correlation with degree of soil weathering and hence surface area; CEC decreased and surface area increased in the more weathered soils. Clay content and soil P sorption maxima were highly correlated (r2 = 0.94) in several Deleware soils (Mozaffari and Sims 1994). Leclerc et al. (2001) evaluated soil survey data and found that soil texture was the main factor affecting soil P sorption capacity (PSC) followed by the dominant soil forming process. The PSC increased with increasing clay content (soil surface area) within Gleysolic soils and within Podzolic soils but the relationship was less clear when the two soil orders were grouped together.

2.3.5.2 Carbonate content Griffin and Jurinak (1974) reported that the kinetics of phosphate interaction with calcite could be described by two simultaneous reactions. The first reaction for adsorption of phosphate on the calcite surface was second order. The second reaction was first-order and was considered to be associated with the surface arrangement of phosphate clusters into calcium phosphate heteronuclei (precipitation). The two processes are indistinguishable when describing P interaction with calcite and were fitted to a two-surface Langmuir isotherm equation. The break in the slope of the Langmuir plot occurred at a P concentration level falling between the octocalcium phosphate (OCP) and the hydroxyapatite (HA) line on the solubility diagram. Such a two stage process of P retention in calcareous soils had been suggested much earlier by Olsen and Watanabe (1957). The scanning electron microscope (SEM) and X-ray diffraction analyses of calcite crystals following reaction with solutions covering a range of P concentrations clearly show the formation of coral-like clusters of Ca-P precipitates on the surface of the calcite for the higher P concentrations (Freeman and Rowell 1981). Below a P concentration of 6 µg/mL, P appeared to be adsorbed as a monolayer on the calcite surface. At higher concentrations precipitates were distinguishable and OCP and DCP were identified. The SEM images also showed that large portions of the calcite surface remain exposed and they estimated that only about 3% of the calcite surface was involved in adsorption and heteronuclei formation. Similarly, Griffin and Jurinak (1973b) estimated that only 5% of the surface area of calcite was involved in P retention.

In Manitoba, the P retention of two neutral, non-calcareous soils was compared with two calcareous soils (Weir and Soper 1962). They found that the neutral soils had much greater P

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sorption maxima and that the sorbed P was almost all exchangeable with 32P within 48 hours. In contrast the calcareous soils retained less P of which most was not exchangeable over the same time period. The higher retention in the neutral soils was attributed to their higher clay contents and therefore higher surface areas. The two non-calcareous soils were clay textured whereas one calcareous soil was a sandy loam and the other a silty clay loam. Retention of P in all four soils could be described by two-surface Langmuir equations suggesting that different retention reactions were occurring at low and high P concentrations. The low exchange rates of P in the calcareous soils may be the result of the formation of slowly soluble Ca phosphates while the precipitates that formed in the neutral soils were more soluble. This was corroborated by Soper and El Bagouri (1964), who found that the plant availability of added P was not correlated to CaCO3 content, but rather, most of the added KH2PO4 formed Al phosphates that were largely plant available. Some Ca-P did form and the amount of Ca-P increased during the growing period suggesting that the Ca-P that formed was not available to the oat plants. A comparison of Manitoba soils of similar texture but differing in CaCO3 content found that the P retention was higher for the calcareous soils (Akinremi 1990). Sorption isotherms are shown in Figure 3.2. The adsorption maxima (determined from fitting the Langmuir equation) of the calcareous soils was about 1.5 times greater than that of the non-calcareous soils. Also, calcareous soils tended to have higher affinity constants than non-calcareous soils. The exception was the non-calcareous Portage clay loam soil which had an affinity constant similar to the calcareous McCreary soil. Holford and Mattingly (1975) used a two-surface Langmuir equation to divide the sorption isotherms of 24 calcareous soils into high energy and low energy sorption regions. The high energy sorption capacities were most strongly correlated to dithionite-soluble Fe and the low energy sorption capacities to organic matter content and CaCO3 surface area (not CaCO3 content). Again this shows that in soil, P sorption cannot be described by a single process or estimated based on single soil characteristic.

2.3.5.3 Soil pH Soil acidity or alkalinity is expressed in units of pH, which are based on the concentration of H+ in the soil solution. Soil pH is commonly measured in a suspension of soil and water or soil and a 0.01M CaCl2 solution. Based on a large number of paired measurements, Conyers and Davey (1988) produced the following factor for converting pH measurements in CaCl2 solution to those measured in water:

pHCa = 1.05 pHw - 0.9

where pHCa is pH measured in 0.01M CaCl2 and pHw is pH measured in water at a 2.5:1 solution:soil ratio for both. The pH was measured in the clear supernatant after the shaken soil suspension had settled. This factor was found to be valid for nonsaline, net negatively charged soils. Note that with this conversion factor a pHw of 7.5 is equal to 7.0 pHCa and therefore the range of pHCa of 6.5 to 7.0 in soils is considered to be neutral (pHw of 7.0 to 7.5).

The solubility of the various P compounds common to soil are such that soil solution levels of P and hence plant available P are generally highest in the pH range of 6 to 6.5 (Lindsay 1979). Below pH 6, the solubilities of Al and Fe phosphates decrease therefore reducing the availability of the P from Al-P and Fe-P precipitates. Above pH 6.5 the solubility of Ca phosphates decreases and again reduces the availability of P in Ca-P precipitates.

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However, solution P concentration is not determined by solubility alone. The soil pH also affects P. The form of the phosphate ion changes with pH and the surface charge on Fe and Al oxides and clay edges also changes with pH. These edges are positive at low pH and become negative as pH increases (see section 3.3.1 for pH effects on sorption/desorption reactions). In soils containing Fe- and Al-oxides, Leclerc et al. (2001) showed that for soils of similar clay content, P sorption decreased and water extractable P increased as the pH increased to near neutral.

A third process which needs to be considered is the effect of pH on the cations in the soil solution that may react with P to form precipitates. An Al-montmorillonite subjected to a range of pH values from 4 to 9.0 showed a sorption minimum at pH 4 and a maximum at about pH 7 (Traina et al. 1986) which is the opposite of what one would expect if Al-P solubility was controlling solution concentration. However the solution P appeared to be in equilibrium with an Al-P and the reason for the low solution P concentration at pH 7 was attributed to the replacement of exchangeable Al by other cations and the subsequent reaction of the Al with P in solution to form solid Al-P precipitates. So even though the Al-P was more soluble at pH 7 than at lower pH, the fact that more Al was available to react with P at pH 7 than at lower pH values resulted in much more of the Al-P precipitate at pH 7.

Therefore, due to the heterogeneity of materials present, the effect of pH on P sorption in soils can be very complex. Soils in which aluminosilicates are the predominant P adsorbents may show a decrease or no change in P solubility with increasing pH up to pH 7 (Traina et al. 1986). Erich et al. (2002) reported no change in water soluble P over the range of pH 5 to 7 in soils that contained low amounts of exchangeable Al (< 1.2 mmol Al/kg in their soils). White (1983) predicted that any soil with an exchangeable Al(III) content of less than 2 mmol/kg would not show a change in P solubility as pH was increased to 7 although soils with higher exchangeable Al contents would be expected to show a decrease in P solubility.

2.3.5.4 Fe and Al oxides Fe and Al oxides have been mentioned in many of the previous sections. They can adsorb P onto their surfaces (mono-dentate and bi-dentate bonding) and under certain conditions may dissolve releasing Fe and Al into the soil solution to form precipitates of Fe and Al phosphates (Sample et al. 1980). Oxalate extractable Fe had a dominant effect on P reactions in several Mediterranean calcareous soils (Ryan et al. 1984) and P sorption maxima were best explained by amorphous and organically complexed metals in semi-arid calcareous soils of the Pacific Northwest region of the United States (Leytem and Westerman 2002). In some calcareous soils Fe and Al oxides have been shown to control the soil solution P concentration even if Ca-P is the predominant P form in these soils (Tran and Giroux 1987). Weathering and leaching of soils tend to reduce soil pH and increase the proportion of Fe and Al oxides in the soil. As a result, acid soils tend to have higher levels of Fe and Al oxides than do calcareous soils. It is not surprising therefore that these oxides play a major role in P sorption in acid soils. Dithionite extractable Fe and Al explained 73% of the variability in P sorption in acid soils under various management regimes (Agbenin 2003). The role of Fe and Al oxides is so dominant in acid soils that the P sorption capacity (PSC) and degree of P saturation (DPS) are estimated based upon the quantity of oxalate extractable Fe and Al in the soil (Beauchemin et al. 1996; Breeuwsma et al. 1997; Anderson and Wu 2001, Erich et al. 2002). In the acid and neutral Quebec soils studied by Leclerc et al. (2001), oxalate extractable Fe and Al correlated well with the P sorption index and the water extractable

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P was linearly related to the increase in DPS as estimated based upon oxalate extractable Fe and Al.

2.3.5.5 Redox potential (effect of flooding) Soils from the Red River and Mississippi floodplains were found to release more P into a P-free solution and were also capable of sorbing more P from a high P solution when subjected to anaerobic conditions (Patrick and Khalid 1974). They suggested that the probable cause was that conversion of ferric oxyhydroxide to the more soluble and highly dispersed ferrous form which increased the activity and surface area of the iron compounds that can react with P. This was corroborated by an increase in oxalate-extractable amorphous iron oxide under reducing conditions. Ferric oxyhydroxide binds P more firmly than the ferrous form but would probably have less surface area exposed to P in solution than the gel-like hydrated ferrous oxide or ferrous oxyhydroxide. Therefore, Patrick and Khalid (1974) speculated that the reactions of P with Fe are more important than the influence of pH, Ca, or Mg in anaerobic systems. However, in artificially flooded natural floodplain soils (Florida) the increase in available P with flooding was due to release of microbial P while Al and Fe phosphates remained unchanged (Wright et al. 2001).

2.3.5.6 Soil organic matter Interactions of P with organic matter have long been demonstrated in pure systems. Weir and Soper (1963) showed that humic acid-Fe complexes held P ions more strongly than an anion exchange resin. Synthetic organic acids have been shown to increase or reduce soil P sorption depending on their concentration and affinity for Al (Traina et al. 1986). At low concentrations, and for organic acids with a weak affinity for Al, the acid reacts with adsorbed Al hydroxides to increase the amount exchangeable Al. This exchangeable Al then reacts with P in solution to precipitate as Al-P. Organic acids with a high affinity for Al (if at high enough concentration) will react with Al from the mineral particle and carry it into solution as soluble Al-organic complexes rendering the Al unavailable to react with P. Organic matter has also been shown to interfere with the formation of the more common Ca-P formed in neutral and calcareous soils. Organic acids can slow the formation of OCP (Grossl and Inskeep 1991) and hydroxyapatite (Inskeep and Silvertooth 1988; Grossl and Inskeep 1991), and to a lesser extent DCPD (Grossl and Inskeep 1991).

An influence of organic matter on P sorption has been shown in soil systems as well. Alfalfa residue was shown to increase extractable P of several Quebec acid soils (Mnkeni and Mackenzie 1988) and farm yard manure reduced the P sorption maxima and increased desorption in a highly calcareous soil from Syria (Habib and Hayfa 1994). Sorption of P resulted in an increase in dissolved organic carbon suggesting a competition between P and organic anions for adsorption sites in an acid soil from Maine (Erich et al. 2002). Ohno and Crannell (1996) compared the effects of extracted soluble organic matter from vetch, clover, cattle manure, and poultry manure on P sorption characteristics of an acid soil from Maine (pH 5.0). The green manure extracts and a synthetic citric acid inhibited P sorption but the manure extracts had no effect. Al was released from the soil into the solution only in those cases where P sorption was inhibited indicating that the organic acids inhibited P sorption by making the Al unavailable to react with P as explained above. The greater molecular weight (size) of the organic matter extracted from the manure may have reduced its ability to interact with surface Al and hence did

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not impact P sorption (good correlation between decreasing MW and inhibition). Following addition of the manure, the calcareous soil studied by Habib and Hayfa was incubated for 19 months prior to the determination of P sorption. Decomposition of the manure would be expected during that time and would have resulted in a different suite of organic acids than would be found in a manure extract.

In studies comparing the effects of inorganic fertilizer P (TSP) and organic P sources (manure) on P soil P retention, near neutral soil (pH 6.5) treated with inorganic P, showed a greater ability to sorb and retain solution P than when amended with pig or cow slurry (Anderson and Wu 2001) or dairy manure (McDowell and Sharpley 2001). Calcareous Manitoba soils (pH 7.6 and 8.0) showed similar results up to a P rate equivalent to 440 kg/ha (Ajiboye et al. 2003). Hog and dairy manures both reduced P sorption to a greater extent than did MAP or biosolids amendments. Erich et al. (2002) found that P sorption maxima was not affected by the addition of compost or manure but the sorbed P was held more weakly and as a result the solution P concentration was higher with the manure and compost treatments than in the inorganic fertilizer treatment. Ajiboye et al. (2003) also reported higher P solution concentrations when dairy or hog manure was added than when inorganic P or biosolid was added.

Organic matter and P can interact in many ways due to the wide range of possible reactions of P within the soil and the wide range of organic compounds present in the soil. Depending on the condition organic matter may increase or decrease the soils ability to hold P but in the majority of situations it has been shown to weaken the sorption of P and hence make P more plant available and consequently also more susceptible to loss by leaching or runoff.

2.3.6 Inorganic and Organic P Amendment Reactions in Soil Common inorganic P fertilizers used in Manitoba are highly water-soluble granular forms which contain from 19.6 to 22.7% P (45-52% P2O5) or highly concentrated fluid fertilizers which contain from 14.8 to 23.6% P (34-54% P2O5). In contrast animal manures are a much less concentrated source of P. Akinremi et al. (2003) reported P contents ranging from 0.25 to 4.5% on a dry weight basis or 0.06 to 0.56% P on a fresh weight basis (0.6 to 5.6 g P/kg fresh manure). One-half to two-thirds of the manure P was readily available for plant use (Table 3.2). Barnett (1994) determined the P content of feces from several classes of livestock and from several sources for each class and found wide ranges in P content; for most classes of livestock the manure source with the highest P content contained at least twice as much total P as the lowest P source (Table 3.3). The relatively low concentration of P (and other nutrients) in animal manures requires that large quantities of manure be applied to the soil to supply crop needs. For example, dairy manure containing 1.3 g P/kg of fresh manure would need to be applied at 6700 kg/ha to supply the same amount of P that would be added by 40 kg/ha of MAP. If only the labile fraction of the P is considered almost 10 000 kg/ha of the same manure would need to be applied. Not only is the P concentration much lower in manure than in inorganic P fertilizer, but the P is in a variety of forms and is applied together with large quantities of organic material. Some of the P in manure is therefore unavailable to plants until mineralized through microbial action and some may remain unavailable as it is incorporated into highly recalcitrant soil organic matter, is occluded within soil aggregates, or reacts with cations present in the manure to produce highly insoluble precipitates. The result is that the P in inorganic P fertilizer and in manures reacts very differently after application to the soil. Both inorganic and manure P sources resulted in P leaching to a depth of nearly one meter in neutral to acidic soil horizons (Eghball et al. 1996).

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However, the ability of P to leach deeper into calcareous sub-soils under manure amended soil than under inorganic P amended soil has been attributed to the movement of P in organic form (Campbell and Racz 1975; Eghball et al. 1996; James et al. 1996). James et al. (1996) detected P (Olsen STP) well above background levels as deep as 210 cm in calcareous soils with a history of heavy application of turkey or beef manure in Utah. Both the organic and inorganic P constituents of a manure extract moved more rapidly through calcareous soil than P added as KH2PO4 (Campbell and Racz 1975). The elevated STP levels at depth were likely in part the result of manure interfering with soil retention of inorganic P (Pi) and not only due to Po movement and subsequent mineralization.

Table 3.2. Phosphorus content of animal manures (adapted from Akinremi et al. 2003)

Dry wt basis Fresh wt basis

Percent solids total labilez total labile

--------------- g/kg ----------------- Dairy 23.5 5.5 3.8 1.3 0.9 Beef 30.7 2.5 1.5 0.8 0.5 hog (storage lagoon)

4.2 33.4 18.4 1.4 0.8

hog (sow barn) 12.4 45 28.4 5.6 3.5 hog (nursery barn) 3.0 21.4 13.5 0.6 0.4 z labile P is P that is soluble in water or 0.5N NaHCO3.

Table 3.3. Range of total P content in fresh, uncontaminated livestock feces raised on commercial farms (Barnett 1994)

Dry matter content

Range of dry matter content Total Pz Range of total

Pz

------------------------- g/kg ------------------------- Dairy Feeder cattle Finisher cattle Hogs Broilers Layers

143 156 172 272 675 280

135-162 125-180 136-241 210-365 371-855 236-383

9.3 6.7 6.7 29.1 18.0 24.2

6.0-16.0 4.5-10.7 3.7-10.6 19.7-40.0 13.1-23.3 16.2-30.3

z Dry matter basis Inorganic P fertilizer, even when applied at low rates, results in localized regions with very high P (and other ion) concentrations and for most P fertilizers results in low pH as well. The sequence of events following fertilizer P addition is well described by Sample et al. (1980) and is

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summarized in the section 3.3.2. The types and therefore the solubility of the precipitates that form are a function of the cations that are present (function of the original soil composition), the solution pH, and the P concentration. With continued dilution at the edge of the fertilizer zone eventually dissolution and precipitation cease to be of importance and adsorption of P by cations on the surfaces of soil particles becomes the dominant P reaction. With time and as solution P concentration decreases near the location of the initial granule/droplet (plant uptake or rainfall) some of the initial precipitates will dissolve, release some P to solution and form more insoluble precipitates. Adsorbed P may become more strongly sorbed with time or it can also be released back into solution where it may be taken up by plants or may move to new adsorption sites. If no more P is added, soil P gradually shifts to more stable forms and soil solution P concentration decreases. The solution P concentration and plant available P are therefore controlled largely by the solubility of these precipitates.

Manure P reacts differently from fertilizer P when added to soil. First, due to the lower concentration of P (and other ions) in the manure, application of manure will not result in localized zones of high P concentration and high ionic strength. As a result dissolution of soil minerals will be reduced and precipitation of P compounds will be less likely due to lower solution concentrations of both P and the necessary cations. However, depending on the class of livestock and the feed mineral supplement used, manure can contain substantial quantities of cations (such as Cu, Zn, Mn, K, Na etc.) which may react with P or displace other cations from the soil exchange to react with P. Hog manure may contain up to 800 mg Cu/kg, 3600 mg Zn/kg, 530 mg Mn/kg, and 0.3% Fe (M. A. Kashem, personal communication, postdoctoral fellow, University of Manitoba, Winnipeg, MB). Although cattle manure contains lower concentrations of metals, it still contains significant quantities of Na and K salts which can result in salinity problems in some cases (Hao and Chang 2003) and are also able to displace other more P reactive cations from the soil exchange sites (Akinremi and Cho 1991a, b).

Organic acids have been shown to reduce both P sorption (Traina et al. 1986; Ohno and Crannell 1996) and precipitation of Ca-P (Inskeep and Silvertooth 1988; Grossl and Inskeep 1991), but can also increase the precipitation of Al-P if the soil clay minerals are dominated by exchangeable Al (Traina et al. 1986). Many researchers have reported greater reductions in P sorption and concomitant increases in soluble P in soils amended with manure than when amended with inorganic P. This effect has been reported for calcareous soils (Abbott and Tucker 1973; El-Baruni and Olsen 1979; Habib and Hayfa 1994; Whalen and Chang 2002; Ajiboye et al. 2003), neutral soils (Anderson and Wu 2001; McDowell and Sharpley 2001), and acid soils (Iyamuremye et al. 1996; Zheng et al. 2001). The reduction in P sorption is greater at higher application rates of manure (Figures 3.6 and 3.7). In general this effect has been attributed to organic material (most probably organic acids) either coating potential P sorption sites or interfering with precipitation reactions.

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Figure 3.6. Effect of pig slurry manure rate on P sorption isotherms. Points for the three replicates of each treatment are shown. LPIG is the lowest rate, MPIG is the intermediate rate, and HPIG the highest rate of slurry application. The best fit trend line is shown for each application rate (Anderson and Wu 2001).

Figure 3.7. Effect of cow slurry manure application rate on P sorption isotherms. Points for the three replicates of each treatment are shown. The best fit trend line is shown for each application rate. LCOW is the lowest rate of application, MCOW is the intermediate rate, and HCOW is the highest rate of slurry application (Anderson and Wu 2001).

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Some of the differences between the reaction products of soil and inorganic or manure P (which is a mixture of organic and inorganic P compounds) have been shown by Kashem et al. (submitted 2003). Following 16 weeks incubation, these researchers sequentially extracted soil that had been amended with hog manure, cattle manure, biosolids, and MAP. The soil P extracts following one week of incubation were compared with those taken at 16 weeks. The P transformations with MAP and hog manure were similar over this time period showing a decrease in water extractable P and a corresponding increase in NaHCO3 extractable P with incubation time, whereas the cattle manure treatment showed a decrease in both of these P groups with an increase in less labile forms. The biosolid amended soil showed an increase in both of the more labile P forms and a decrease in the most recalcitrant P. Clearly there can be large differences in soil P reactions among organic P sources as well as between inorganic and organic P sources.

2.3.7 Management Factors Affecting Phosphorus Retention and Release Soil and manure properties clearly have a large impact on the reactions of P with soil and the potential loss of P to surface waters. Because P transport by water requires that the soil/fertilizer/manure P must interact with the water, management can also have a large impact on P loss. Also, in many cases, the majority of runoff and P loss come from only a small portion of a field and from only a few runoff events (Gburek and Sharpley 1998; Sharpley et al. 1999a). Therefore identifying areas within fields that are prone to leaching and runoff is the first step in developing field management plans that will minimize nutrient loss to the environment (Fuller 2001).

2.3.7.1 Application method, placement and timing A worst case scenario in Manitoba would be the application of manure, biosolids, or fertilizer P in January on top of snow on a field next to a river. It would be expected that as the snow melts, the meltwater would carry most of the fertilizer or manure P (and any other soluble constituents) across the frozen soil and into the river. What are the options to minimize such a loss? Application and incorporation of nutrients as close as possible to the time of crop need would be ideal, as this minimizes both the risk of P loss and ammonia volatilization (Steinhilber and Weismiller 2000). Incorporation is important as it mixes the manure with soil which can potentially increase P sorption and also reduce the P concentration at the soil surface. Lower P concentration at the soil surface where the runoff occurs will result in lower P runoff. An obvious exception to incorporation is the application of manure and fertilizer to perennial forages, a practice that is becoming more common in Manitoba as forage acreage increases. The 2001 census reported 877 302 ha (2 167 860 acres) of tame hay and fodder in Manitoba (Statistics Canada 2002). The first rainfall after application of manure or inorganic P fertilizer on perennial forage plots resulted in large DP losses (DeLaune and Moore 2001). The highest runoff losses (103 ppm) were from plots amended with triple super phosphate (TSP; 0-46-0), the inorganic P source, as it had the highest soluble P (SP) content. The loss from the manures (for example 26 ppm for normal poultry litter) was also proportional to their SP content. Heathwaite (1997) reported that runoff from grass plots following surface application of cattle manure, cattle slurry, or calcium phosphate fertilizer was dominated by DP in most cases. The highest losses of P were from the fertilizer treatment (2.5 kg TP/ha, 1.5 kg DP/ha) which was over five times the losses from the manure treatment (0.4 kg TP/ha) and over 20 times the losses from the cattle

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slurry treatment. Edwards and Daniel (1994) reported a similar effect on fescue plots for the first runoff event although for the second and subsequent runoff events P loss was much lower and there was no treatment effect. If several small non-runoff precipitation events occur before the first runoff event, P loss is greatly reduced (DeLaune and Moore 2001). Therefore, whenever manure/fertilizer P is applied, it should be incorporated or injected, if possible. If surface application is unavoidable, manure should be applied when the risk of runoff is lowest to allow for gentle rains to carry the SP into the soil.

2.3.7.2 Application rate Each soil has its own unique properties that determine its P sorption capacity (PSC). As P is added to the soil, the PSC becomes used up and each subsequent increment of sorbed P is held less tightly and therefore solution P concentration increases. Higher solution P concentration means more plant available P but that P is also more susceptible to loss. The degree of P saturation (DPS) is often used as an estimate of how much room is left in the soil for P and is determined by taking a measure of STP and dividing it by the PSC. In the Netherlands, a DPS value (estimated from oxalate extractable P, Al, and Fe) above 25% is considered an unacceptable risk for P loss to shallow groundwater and no further P applications can be made to that soil (Sims 2000b).

Matching soil available P plus P additions to crop needs is the best way to minimize the risk of P loss while maintaining acceptable crop yields (Steinhilber and Weismiller 2000). Under rainfed conditions in Alberta, Whalen and Chang (2001) found that annual beef manure applications up to 30 tonnes/ha did not result in significant increases in available P (Olsen STP) below 100 cm (i.e. leaching losses of P). However, applications of 60 or 90 tonnes manure/ha resulted in a 2.5 fold increase in available P in the 120-150 cm depth. Due to the differences in N:P ratios in manure and in crops (4:1 in manure versus 8:1 crops), manure application based on crop N requirement results in excess P application (Beegle 2000). Manure should be handled to conserve N (give a better N:P ratio) and should be applied to meet P needs with other N sources used to supplement the manure. Steinhilber and Weismiller (2000) recommend a method of infrequent manure application with N supplementation on an annual basis which is operationally easier than trying to apply manure annually at the low P rate. As an example, they used a corn-wheat-bean rotation. Manure would be applied to meet the N requirement for alternate corn crops (i.e. every sixth year) and inorganic fertilizers applied in other years based on soil test information. Within such a system, field regions with a high risk of P loss should be identified and receive less manure or no manure and should also be managed to minimize soil erosion.

2.3.7.3 Crop choice Crops have different P needs and different P concentrations in their harvested portions. For example, canola may contain twice as much P in its seeds as wheat but its lower yield results in similar export values (Steinhilber and Weismiller 2000). Due to their high biomass removal, forage crops would therefore be a better choice for “mining” soils with excessive P. Perennial forages are excellent crops for reducing soil erosion; however, substantial losses of DP from forage fields have been measured during snowmelt (Green and Turner 2002) and are likely due to the large quantity of P which can be leached from forage plants that have been frozen and then thawed (White 1973). The development of special hyperaccumulator crops and varieties for P “mining” has been suggested (Novak and Chan 2002). However, the effect of such a plant on the

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distribution of P in the soil profile could actually exacerbate the problem. Sharpley (1981) found that the majority of runoff P during the growing season originated as leachate from the crop canopy and not from the soil. Even if growing season runoff is minimal, the hyperaccumulator crop could potentially move large amounts of P from the rooting zone to the soil surface through canopy leaching and thereby increase the amount of P loss in runoff.

2.3.7.4 Conservation tillage Conservation tillage reduces the loss of PP due to erosion by water and wind but can increase losses of DP. On averaging the results of 13 studies, Sharpley et al. (2002) reported that conservation tillage decreased total P losses by 75% but increased DP losses by 5%. Snowmelt water extracts P from the soil surface and from plant residues (Rekolainen 1989). In a zero-till or perennial forage system, fertilizer and manure cannot be incorporated as in conventional tillage systems. In combination with vegetative leaching, the result can produce soil P stratification with elevated levels in the uppermost few centimetres of the soil compared to tilled fields (Follet and Peterson 1988). Conservation tillage also tends to increase snow trap and thus the risk of snowmelt runoff (Elliot et al. 2001). As a result, total losses from conservation tilled fields may be at least as great as those from conventionally tilled fields (see section 3.4.1).

2.3.7.5 Vegetative buffer strips and riparian areas Vegetative buffer strips can be effective in reducing soil and P transport to surface waters in some cases. Switchgrass buffer strips reduced total reactive P (TRP) in runoff from manured forage plots by 47-76% in a two year study in Texas (Sanderson et al. 2001). Heathwaite (1997) obtained a 98% reduction in TP (both PP and DP were reduced) load in runoff from plots receiving inorganic P fertilizer, but only a 10% reduction in P loss from the plots receiving cattle slurry. Vegetative buffer strips reduced runoff volume, and total solids and DP in runoff water by 67, 79, and 83% respectively (Young et al. 1980). The removal of DP appears to require the infiltration of the runoff into the soil. There was only a 12% reduction in P due to a vegetative buffer strip when 65% of the runoff from a dairy barnyard passed through the strip as surface flow (27% as subsurface lateral flow) (Schellinger and Clausen 1992). Furthermore, the effectiveness of vegetative filter strips can decrease with time (Gillingham and Thorrold 2000) and could potentially become a source of DP to runoff. Also, the sedimentation of PP in the vegetative strip and the DP that is retained by the surface soil enriches the soil within the strip with P. The increase in available P (and other nutrients) with time would be expected to increase vegetative production and possibly also increase P levels in the vegetation. As 35 to 75% of plant P is rapidly released following tissue death (Richardson 1985), the P contained in the plant residue becomes available for leaching by spring meltwater unless harvested and removed. Under Manitoba conditions, however, where the majority of runoff occurs as a result of snowmelt runoff over frozen soils and when vegetation is dead or dormant, filter strips probably have limited benefit for reducing DP.

2.3.7.6 Constructed wetlands The use of constructed wetlands for the treatment of feedlot runoff and dairy wastewater is being studied in Canada (Manitoba Agriculture and Food 2001; Jamieson et al. 2002) and elsewhere (Healy and Cawley 2002). In wetlands, net microbial and plant P uptake is small as these pools are easily saturated and retention by sedimentation and sorption by the soil are the major

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processes for retaining P (Richardson 1985). Retention of P by freshwater wetlands was correlated to most closely to amorphous (acid oxalate extractable) Al and Fe content (i.e. soil P retention capacity). However, Richardson (1985) found that wastewater filtration by wetlands results in P saturation within a few years that then results in P export from the wetland cells. In Nova Scotia, a 1022 m2 wetland was used to treat the wastewater and feedlot runoff from a 30 cow dairy herd (Jamieson et al. 2002). After four years the P retention capacity of the soil in the wetland had been reduced from 1600 to 925 mg P/kg soil. Based on this, the wetland was estimated to have an effective P removal lifespan of eight years. For a wetland to remain effective, the P enriched vegetation and, more importantly, the P enriched soil and sediment would need to be removed.

2.3.7.7 Grazing intensity and streambank protection Although pasture (and the manure that is “surface applied” by grazing livestock) is not considered a major source of P export to surface water, overgrazing can dramatically increase P losses. Lightly grazed (< 4 dairy cows/ha) and ungrazed grassland plots produced low TP losses (0.076 and 0.020 kg/ha respectively) during a 4 hour simulated rainstorm (Heathwaite 1997). For the same simulated storm, the heavily grazed treatment (> 15 dairy cows/ha) had TP losses of 2.9 kg/ha of which 15% was DPi and the remainder PP and DPo. The increase in P loss due to overgrazing was attributed to removal of most of the vegetative cover, soil surface compaction by trampling, and direct addition of P in excreta. Feeding and watering areas tend to be grazed more heavily and if these areas are located near the drainage network, they can represent a major source of P to surface waters (Heathwaite 1997). This effect would be compounded where cattle on pasture receive supplemental feed (including P) and hence excrete more P onto the pasture than is removed by grazing. The unrestricted access of livestock to streams and lakes can locally add P directly to the water. Not only is P added directly by feces and urine being deposited into the water but also a significant amount of sediment (which contains P) is added due to increased stream bank erosion as a result of livestock trampling. Stream bank sections in pastures that were accessible to cattle had three times the erosion losses of those sections that were protected by fencing (Kauffman et al. 1983). Lyons et al. (2000) found that intensive rotational grazing was just as effective as ungrazed buffer strips in reducing stream bank erosion compared to continuous grazing.

2.3.7.8 Nutrient content of manure Manure is often applied to land to meet plant N requirements. Because the N/P ratio is narrower in manure than that needed by plants, excess P is applied if plant N requirements are to be met. Methods of widening the N/P ratio are discussed in chapter 2. Manure application based on P content with supplemental N added via fertilizer or N fixing crops could be used to prevent P accumulation in soils receiving manure.

Runoff P losses from perennial forage following surface application of manure is directly proportional to the soluble P content of the manure (Delaune and Moore 2001). Several methods of reducing the available P in manure are under study (more details in chapter 2) and include the reduction of inorganic P added to dairy, hog, and poultry diets (reduces total and available P in manure), and the addition of chemical agents such as alum or fly ash to the manure (reduces the availability of the P in the manure). At equivalent rates of manure, the high P dairy diet resulted in 10 times higher DRP concentrations in runoff than did the low P diet (2.84 and 0.30 mg/L

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respectively)(Ebeling et al. 2002). The high P diet resulted in runoff DRP concentrations four times that of the low P diet when the manure was applied at equivalent rates of P. The high P diet had much higher contents of total P, water soluble P, and bioavailable P. Addition of phytase has been shown to reduce total P in manure but if dietary supplementation with inorganic P is not reduced, soluble P in the manure and risk of loss to water may not be reduced (Waldroup 2002). The addition of alum or fly ash did not change the ratio of total N to total P of the manure but did greatly increase the ratio of N to water extractable P (Dao 1999).

Manure contains variable amounts of cations such as Ca, Mg, Al, and Fe which can react directly with manure P to form precipitates under the right conditions. Other cations which are present may not react directly with P but may displace Ca, Mg, Al, and Fe from the soil exchange is sufficient quantities to affect P retention.

2.4 Mode of Phosphorus Transport to Surface Waters Water on the surface of soil will either infiltrate into the soil, be lost by evaporation, or leave the field as surface runoff. The water that infiltrates into the soil and that which is lost in surface runoff is the water of importance in considering P transport pathways.

Surface runoff, also called overland flow, is divided into two types based upon the cause of the flow (Scherrer and Naff 2003). Hortonian overland flow occurs when precipitation exceeds the infiltration rate of the soil and the excess is lost as runoff (infiltration and surface runoff occur simultaneously). Saturation overland flow occurs when precipitation falls on a saturated soil and the water is lost as surface runoff because infiltration cannot occur into the saturated soil. Water that infiltrates into the soil may also be lost by deep percolation, or by subsurface lateral flow. Subsurface lateral flow may emerge at the soil surface at another point and become overland flow. Soil texture, structure (porosity, compaction, continuity of macropores), slope, presence of impervious layers, depth to water table, and antecedent moisture are the major soil factors that determine the pathways by which water is lost (Table 3.4). On a field scale, topography can have a major influence on saturation overland flow due to localized and temporary regions of saturation. This can result from the convergence of subsurface lateral flows above less permeable layers, high water tables caused by convergent flow into hillslope hollows, or high water tables along streams (Heathwaite et al. 2000). The four main source areas for saturated flow in a field with hilly topography are: (1) areas adjacent to perennial streams (base of hillslopes), (2) areas of concave upward-slope profile (slope profile concavities), (3) hollows (areas of concave outward contours, and (4) areas with thin or impermeable soils (Heathwaite et al. 2000).

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Table 3.4. Process criteria to determine the different flow processes that will occur in a soil (adapted from Scherrer and Naff 2003).

Type Intensity of runoff process Process criteria: characteristics and conditions

Overland flow processes

Hortonian (infiltration excess flow)

Immediate Hortonian overland flow due to infiltration hindrance

Soil or surfaces with serious infiltration hindrances: soils with extremely high clay content, compacted soils (machines or cattle), bedrock surfaces with low permeability.

Delayed Hortonian overland flow due to infiltration hindrance

Hydrophobic soils (soils with extremely dense root network near the surface), compacted soils with low macropore density, sealed and crusted soils, macroporous soils with small water exchange between macropores and soil matrix.

Saturated overland flow

Immediate saturation overland flow due to antecedent soil saturation

Soil water level near surface combined with good permeability of soil layers (macroporous, permeable matrix), which enable infiltration and saturation after short rainfall, absence of lateral flow structures.

Saturation overland flow due to slowly saturating soils

Permeable, shallow soils with a low permeable subsoil (e.g. bedrock), soils with a water level in the subsoil, absence of lateral flow structures

Delayed overland flow due to very slowly saturated soils

Thick macroporous soils with permeable soil matrix, which can only be saturated after extensive rainfall

Subsurface flow processes

Lateral flow Subsurface flow Lateral flow in steep and shallow hill-slope soils due to effective lateral flow paths (macropores, pipes, highly permeable layers) in combination with low permeability subsoil or shallow bedrock and minimal interaction between macropore flow and the soil matrix.

Delayed subsurface flow Lateral flow in the soil due to lateral flow paths (macropores, permeable layers) with medium water exchange to the surrounding soil matrix and low permeable subsoil/bedrock,

Strongly delayed subsurface flow

Delayed lateral flow controlled by lateral flow paths (macropores, highly permeable layers) in deep soils

Vertical flow Deep percolation Permeable and deep soil or permeable soil with a permeable geological underground

Snowmelt runoff is a special case due to the influence that frost has on soil porosity and the complex nature of the interaction between soil water and soil particles at low temperatures. Snowmelt forms the majority of the runoff on the Canadian Prairies (Eilers and Buckley 2002) and also in west central Minnesota (Burwell et al. 1975; Timmons and Holt 1977). Runoff from

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snowmelt accounted for over 85% of total runoff from several agricultural watersheds in southern Saskatchewan (Nicholaichuk 1967). Infiltration into frozen prairie soils can be categorized as: (1) Unlimited: the soil contains enough large, surface connected, air-filled macropores (biopores and cracks) to allow the infiltration of most of the meltwater; (2) Limited: infiltration is a function of the snow cover water equivalent and the water/ice content of the surface 30 cm of the soil at the time of snowmelt; (3) Restricted: an ice lens on the soil surface or at a shallow depth impedes infiltration resulting in most of the melt water being lost as surface runoff or evaporation (Gray et al. 1985). Ten years of data on snowmelt infiltration into frozen soil from the prairie region of Saskatchewan show that soil texture has only a small effect on infiltration (Zhao et al. 2002). Clay, silty clay loam, and silt loam soils showed only small differences in infiltration over a 24 hour period. The snowmelt infiltration for the sandy loam soil was only 23% greater over the same time period. In contrast, for the unfrozen soils the hydraulic conductivities were 0.045, 0.146, 0.610, and 3.81 cm/h for the clay, silty clay loam, silt loam, and sandy loam soil respectively (almost a 100 fold difference between the clay and sandy loam soils). Zhao and Gray (1999) proposed a model for estimating snowmelt infiltration into frozen soils based on total soil moisture saturation (water + ice), soil temperature at the start of snowmelt, the soil surface saturation during melting, and the infiltration opportunity time (the time the meltwater is available for infiltration at the soil surface). The model gave good predictions over a wide range of soil textures and demonstrates that traditional runoff and soil erosion models which rely heavily on soil texture to estimate infiltration and hence runoff would not be applicable in the case of frozen soil. It is important to note that it is the soil water and ice content at the time of snowmelt not at the time of initial soil freezing that is the determining factor. Throughout the winter and during snowmelt, temperature gradients can cause soil water to move to the freezing front and form frozen water-confining beds that will restrict infiltration of meltwater (Demidov et al. 1995). The refreezing of infiltrating meltwater will also limit infiltration (Gray and Granger 1985).

Loss of P from land can occur both by surface runoff and subsurface flow and the forms of P transported vary with the hydrologic pathway and the source of P. P can leave the field associated with soil and residue particles carried by water (entrainment) or as dissolved phosphorus in water. The water and associated P can reach surface water bodies by surface flow or by subsurface flow that is discharged to surface water bodies. Subsurface flow can reach surface waters via tile drains, lateral flow which resurfaces, or by percolation to groundwater and subsequent groundwater discharge to surface waters. A conceptual model of P transport (Figure 3.8) illustrates how both source and transport factors play important roles in P transport. Water flows can only move P if P is available to be carried in particulate or dissolved form and large amounts of readily available P will not be transported by water from a field if there is no water leaving the field.

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Energy + Carrier P Sources

↓↓↓↓ ↓↓↓↓ ↓↓↓↓

Hydrology Soil Characteristics Agronomic Management Spatial controls -catchment topography -surface and subsurface hydrological pathways

Temporal controls -precipitation and snowmelt --duration --intensity --magnitude --return period

Soil P Erosion risk Texture/structure Fe and Al oxides Carbonates pH

Nutrient balance (crop export vs. nutrient addition) Inputs -fertilizer -grazing -manures -tillage

↓↓↓↓ ↓↓↓↓ ↓↓↓↓

P forms (organic/inorganic; dissolved/colloidal/particulate

↓↓↓↓

Potentially Mobile P ↓↓↓↓

P Transport -P dissolution -P physical entrainment -incidental P loss

Figure 3.8. Conceptual model of phosphorus transport (adapted from Heathwaite et al. 2000) P is transported primarily as ions of inorganic orthophosphate or associated with organic and inorganic particles that are being carried by water (Heathwaite et al. 2000). The separation of dissolved phosphorus (DP) and particulate phosphorus (PP) in practice is difficult and the arbitrary division of 0.45 µm filtration is generally accepted although some researchers use 0.2 µm or even finer filters. P that passes through the 0.45 µm filter is therefore called DP and the portion of the DP which is detected by the molybdate colorimetric test (Murphy and Riley 1962) is called dissolved reactive phosphorus (DRP). The P that is attached to fine colloidal particles may also pass through such a filter and will then be included in this measure of DP. To reduce confusion the term “P(<0.45)” has been suggested to designate this fraction (Haygarth and Sharpley 2000) but the term DP is still the one most commonly used and will be used in this review as well.

The availability of P for transport is controlled by many soil characteristics and is affected by management practices (Figure 3.8). The influences of soil properties on P retention by soils, and the effects of microbial activity (mineralization/immobilization), crop uptake, and agronomic management are discussed in previous sections.

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2.4.1 Surface Pathways of P Loss The surface layer of a soil normally has a higher P content than deeper soil layers. This surface layer is in contact with surface runoff and, therefore, supplies most of P that is lost to surface water. Sharpley et al. (1981) found that for a range of soil textures, the interaction depth between runoff water and surface soil during a 30 minute simulated rainfall was only 1.5 to 3.0 mm. Under grassland, STP is substantially higher for the 0-1 cm depth than for a sample taken from the 0-10 cm depth for the same treatment (Humphreys et al. 1999). Tomasiewicz (2000) found similar results for several cultivated Manitoba soils. The most extreme variation was in an Osborne clay soil which had 77.0 ppm P in the 2-3 cm depth, 10.6 ppm in the 7-8 cm depth, and 4.5 ppm in the 12-13 cm depth. The other Manitoba soils showed a similar pattern of variation but the range in values was smaller.

The P can be transported as particulate P (PP) if soil erosion occurs and can also be dissolved from the surface soil and carried away by the runoff water. Soil erosion is affected by the intensity, duration, and magnitude of the precipitation as well as the soil condition. The antecedent moisture content, presence of frost, inherent hydraulic conductivity, structure, and vegetative cover will all influence the erodibility of the soil. The concentration of P in the soil which is removed by erosion and the amount of soil lost will then determine the PP lost. Several models, USLE, RUSLE, WEPP and several modifications of these, have been developed to predict erosion. In many agricultural areas the majority of P is lost as PP from tilled fields (Heathwaite et al. 2000). Adsorption and surface precipitation of P onto clay and silt sized particles results in P loss with the eroded soil. A special case of surface erosion loss is termed incidental loss and refers to the erosion loss of applied fertilizer, manure, or dung particles (Haygarth and Sharpley 2000), which in most cases are high in P. There is a high risk of incidental loss of P whenever mineral fertilizer or manure is not incorporated and runoff occurs shortly after application. The incidental loss in surface runoff was 6-8% of the total P surface applied as inorganic P fertilizer or dairy slurry to grassland soils in the UK (Preedy et al. 2001).

Prediction of DP loss is more difficult as the factors that control P dissolution are highly soil specific and complex (see previous discussion on reactions of P in soils). Vegetation and crop residue can also have a major impact on DP content of surface runoff. Sharpley et al. (1992) measured DP and PP losses from several fields under different management (Table 3.5). Total P (TP) losses were reduced dramatically by no-till and even more so under native grass. However the fraction of TP loss that was DP increased as TP was reduced (6% of TP for conventional till wheat, 51% for no-till wheat, and 61% for grassland) and in the case of wheat the actual quantity of DP lost under no-till was higher than under conventional tillage. Conventional tillage, in general, is less intensive on the Canadian prairies than in the studies reported by Sharpley et al. (1992). The use of moldboard plows has been largely replaced by other tillage practices, such as chisel plowing, which retain more crop residue at the soil surface and produce less vertical soil mixing. This results in less soil erosion and potentially greater nutrient stratification in the “plow” layer (surface 15 cm). The DP component dominated P loss in runoff from conventionally tilled fields receiving beef (77% DP) or hog (89% DP) manure in Alberta (Wright et al. 2002a). In contrast, DP made up only 9% of TP losses from a field receiving inorganic P fertilizer (Wright et al. 2002a). However, total runoff and PP loss were much higher from the inorganic P fertilized field resulting in TP losses of 2.14 kg/ha per year compared to 1.21 kg/ha for the beef manured field and 0.07 kg/ha for the hog manured field.

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Table 3.5 Loss of soil and phosphorus from fields under different management (adapted from Sharpley et al. 1992)

Management Soil loss DP PP TP

kg/ha/yr ------------ g/ha/yr ----------

Native grass (6 fields) 43 122 78 200

No-till wheat (2 fields) 1147 1178 1132 2310

Conventional till wheatz

(4 fields) 15315 391 6461 6852

z conventional till wheat included a moldboard plow operation

Runoff studies conducted at the Twin Watersheds study area in the South Tobacco Creek region of Manitoba (Green and Turner 2002) demonstrate the effect of conventional and no-till tillage practices on P losses. Unlike the results reported by Sharpley et al. (1992), DP was the dominant form of P loss from both conventional and no-till watersheds at South Tobacco Creek (85% and 92% of TP losses respectively). From 1998 through 2001, the no-till watershed lost twice as much TP per hectare as the conventional tilled watershed (and over twice as much DP). The concentration of TP in runoff from both watersheds was similar in 1998 and 1999 and ranged from 0.5 to 1.0 mg P/L however the TP loss per hectare was higher for no-till due to higher runoff quantities. Runoff P concentration in 2000 and 2001 was much higher for the no-till watershed (about 1.5 and 2.5 mg/L in 2000 and 2001 respectively) than the conventional tilled watershed (no runoff in 2000 and 1.5 mg P/L in 2001). Preplant broadcast P fertilizer in the spring of 1999 is the likely cause for the large increase in P concentration in the runoff from the no-till watershed and its relatively high proportion of DP (94%). As spring runoff accounted for the majority of runoff from both watersheds, the fertilizer applied in spring 1999 would not be detected in runoff until the following spring. Green and Turner (2002) also reported substantial losses of DP from the alfalfa field likely due to leaching of P from the vegetation as well as from deer droppings (higher fecal coliform counts in the spring runoff from the alfalfa than the other fields indicate a higher population of animals over winter). Other researchers have reported that leaching of P from decaying crop residue may be a major source of DP in the case of no-till or perennial forage (White 1973; White and Williamson 1973; Sharpley et al. 1992; Eghball and Gilley 1999). Under Manitoba agricultural practices, DP is likely to make up the majority of P losses from agricultural fields.

In situations where subsurface lateral flow is contributing to surface runoff, the runoff water has been moving through the soil and will have extracted P from pools not accessible to water moving across the soil surface or may have lost P due to retention by the soil matrix. Whether there is a net gain or loss in P is a function of the P retention capacity of the soil, the degree of saturation of the soil P retention capacity, the concentration of P in the solution, and the degree of contact between the soil and solution (if flow is mainly through macropores the water is not in contact with the bulk solution and there will be less opportunity to gain or lose P; more information on subsurface pathways is presented in section 3.4.2).

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Several workers have also reported a soil test P (STP) change point or breakpoint below which they found minimal DP in runoff or leachate and above which the increase in P loss was linearly related to STP (Heckrath et al. 1995; Hesketh and Brookes 2000; Jordan et al. 2000; Anderson and Xia 2001; McDowell and Sharpley 2001). The value of the change point did vary with soil type and soil extractant. The largest study was that of Jordan et al. (2000) which included 5600 Irish soil samples. They found that soils with less than 22 mg/kg Olsen-P in the surface 7.5 cm of soil lost negligible amounts of P to surface waters. Above 22 mg P/kg the data was much more scattered but with a general trend for increasing P loss with increasing STP. In an attempt to quantify the risk of surface transport P loss from soils several risk indices have been developed (Lemunyon and Gilbert 1993; Sharpley et al. 1999b; Bolinder et al. 2000; Hilborn and Stone 2000; Simard et al. 2001a) and are discussed in a later section.

2.4.2 Subsurface Pathways of P Loss Subsurface flow can occur as preferential flow (flow through cracks, old root channels, earthworm borrows, etc. which is much faster than flow through the bulk soil) or as matrix flow (flow through the bulk soil that results is fairly consistent flow rates at all points within the soil) (Heathwaite et al. 2000). Preferential flow is the dominant subsurface pathway in soils with low matrix flow such as in cracking clays with large quantity of macropores (cracks or biopores). In soils with high conductivity, such as coarse textured soils, matrix flow will dominate. Suburface flow may result in percolation to groundwater or may flow laterally if the conditions of slope and soil permeability are right.

Matrix flow is unlikely to result in significant P transport (Heathwaite et al. 2000) due to the intimate contact of P with soil constituents and the slow movement of the water allowing adequate time for reaction between P and soil. However if groundwater is near the soil surface and the soil has a high rate of matrix flow (sandy soil) the potential for P movement into the groundwater exists where high loading rates of P are applied. DP would be the dominant P form to move by this method. In a study of a shallow unconfined sand aquifer in the Battersea drainage Basin near Lethbridge, Rodvang et al. (2001) found TP levels of 0.39 mg/L in the groundwater beneath an irrigated manured field. Although natural levels were variable in the aquifer, a contribution from the agricultural field was suspected. Leaching of P has been detected under feedlots in Manitoba (Campbell and Racz 1975). The leaching of P was partially attributed to the movement of Po followed by mineralization at depth and partially due to the enhanced movement of inorganic P if in the presence of manure extracts. Lateral subsurface flow has the potential to extract P from the high P surface soils and transport that DP into surface pathways.

Preferential flow (also know as bypass flow) allows for rapid transport of water and P through soil macropores. The reduced time for interaction between P and the soil as well as the limited volume of soil which is in contact with the P suggests that preferential flow could be an important pathway especially where tile drains are present. Lateral preferential flow (also called pipe flow) may also occur in some soils. High P levels were detected in tile drains in a soil with high P sorption capacity at Rothamsted and were attributed to preferential flow (Heckrath et al. 1995). Rapid macropore flow was able to transfer P through P deficient subsoils under grassland in the UK (Preedy et al. 2001). Both DP and PP were detected in about equal proportions in tile drains under clay soils in Sweden (Ulén 1995). TP loss ranged from 0.064 kg P/ha under pasture to 0.233 kg P/ha under annual crops with most of the P loss occurring from the soil surface with heavy rain on newly frozen soil. In an attempt to take preferential P loss into account, soil texture

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and distance between tile drains were included in the province of Quebec’s P index (Simard et al. 2001b).

2.4.3 Snowmelt Runoff and P Loss Snowmelt is the source of most of the runoff on the Canadian Prairies (Eilers and Buckley 2002; Green and Turner 2002) and also in west central Minnesota (Burwell et al. 1975; Timmons and Holt 1977). It is not surprising then that most of the soluble nutrient losses from standard erosion plots in Minnesota occurred during snowmelt runoff (Burwell et al. 1975). Winter or fall application (on frozen soil) of fertilizer (Nicholaichuk and Read 1978) or manure (Uhlen 1989) resulted in very high P losses. Almost 10% (2.3 kg/ha) of the fertilizer was lost in runoff and DP in the runoff reached concentrations of 4.6 mg/L. The manure that was applied in winter resulted in snowmelt runoff DP concentrations of 1.73 mg/L and 0.7 mg/L for row cropped and perennial forage plots, respectively. Interestingly, forage plots which received annual applications of 50 kg P/ha surface applied in spring had P losses as high as those from the plots receiving winter applications of manure (Uhlen 1989). Adjacent grain plots that were fall plowed had much lower runoff DP concentrations (0.07 to 0.08 mg/L) regardless of manure or fertilizer treatment. Under equivalent fertilizer or manure treatments DP in the runoff from the perennial forage plots was seven times higher than the grain plots (with fall tillage) and almost four times greater than the row crop plots. Due to greater erosion and therefore higher PP loss from the tilled plots the difference in TP losses was smaller (0.19 mg/L for grain, 0.22 mg/L for row crops, and 0.67 mg/L for forage). After accounting for differences in runoff volume, the TP loss from forage plots was still about double that from tilled plots (9.0 kg/ha for forage, 4.7 kg/ha for grain, and 5.5 kg/ha for row crops). DP concentrations in snowmelt runoff from unfertilized stubble and summerfallow fields in Saskatchewan were similar (0.2 mg/L and 0.3 mg/L respectively) but total P loss was much lower for stubble fields at least partly due to lower runoff volumes (Nicholaichuk and Read 1978). The effect of tillage was also evident in the Saskatchewan study as tilled fallow fields had higher TP losses than stubble despite similar DP losses.

In a comparison of forage and fall tilled grain plots, snowmelt P losses were most closely correlated to the amount of residue left on the ground surface in autumn, the P content of the residue, and fertilizer P history but not to STP (Uhlen 1988). Perhaps, P stratification can explain the lack of correlation with STP. For agronomic purposes, STP is usually determined on a plow layer sample (0-15 or 0-20 cm) which in most cases will include the majority of the plant available soil P pool. On the forage plots, fertilizer was surface applied each spring and the P would move into the soil with rainfall resulting in a much higher P concentration in the surface few centimetres than at depth. Fertilization history would therefore be a better predictor of soil P near the soil surface of forage plots than STP. The effect of residue P content on spring runoff P losses was demonstrated by Uhlen (1988). Plots mulched with vegetable residue (three-fold higher P content than the hay) had higher P losses than plots mulched with timothy-clover hay when both were applied at equivalents rates based on dry matter content. Timmons and Holt (1977) attributed P loss from native prairie (0.10 mg/L DRP and 0.3 mg/L DPo) to P in the precipitation and to leaching of dormant vegetation and decomposing residue. In some cases, a significant portion of the DP in runoff is supplied by the precipitation. An annual contribution of 0.13 kg P/ha was measured in Minnesota precipitation (Burwell et al. 1975) and 0.5 kg P/ha was reported in France (Pommel and Dorioz 1997). Fine dust and other air pollutants are the likely sources for this P. Soil erosion test plots in Minnesota have shown that the leaching of winter dormant alfalfa by snowmelt runoff can result in significant losses of DP (Burwell et al. 1975).

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Although the alfalfa plots had the lowest annual TP losses they had the highest losses of DP and DRP (Table 3.6) most of which occurred during snowmelt runoff. In contrast, 95% of the P lost from the other crops (tillage included fall plowing and spring disking) was lost as PP during rainfall events.

Table 3.6 Estimated annual phosphorus losses from soil erosion plots in Minnesota (Burwell et al. 1975) TP PP DP DRP

-------------------------------- kg/ha -------------------------------- Fallow Continuous Corn Oats Forage

33.3 18.6 5.25 0.68

33.1 18.2 5.01 0.02

0.19 0.41 0.24 0.66

0.10 0.25 0.14 0.39

Demidov et al. (1995) described four potential hydrological processes during snowmelt: (1) Snow melts on non-frozen soil (common in forests soils). Infiltration capacity may be several times greater than snowmelt water and the meltwater infiltrates throughout the winter with the result that there is no runoff. (2) Soil thaws before or at the onset of snowmelt (shallow frost due to mild winter). Most meltwater infiltrates and runoff is minimal but may occur on fine textured soils. (3) Complete thaw of soil occurs during the snowmelt period (thawing takes place from below and above). The soil above the frozen confining bed is saturated by meltwater and is therefore unstable and is easily eroded. Some runoff will occur, especially in the earlier snowmelt period. (4) Frozen soil layer thaws exclusively from below and the thaw is not complete until after the snow cover is melted. Meltwater flows along the surface of the frozen soil. Some destruction of soil by meltwater occurs and erosion potential is highest. Most of the meltwater may be lost as runoff. Of these processes, 3 and 4 produce runoff and are therefore likely to remove P from a field during snowmelt.

Uhlen (1988) described the change in snowmelt runoff P concentration with time as having a characteristic high-low-high pattern during snow and ice melt. The P concentration in runoff during the initial “high” concentration period was well correlated to the grass residue P content and the P concentration during the second “high” period was well correlated to the P fertilizer history of the plot. If we compare the processes proposed by Demidov et al. (1995) to this pattern it appears that the initial period of high concentration was due to water flowing over frozen soil and extracting P from the dormant forage and residue on the soil surface with minimal interaction with the soil beneath. During the “low” phase the soil is still frozen and most of the water extractable P has been removed from the residue. The second “high” phase occurs as the soil begins to thaw from above and the meltwater is able to saturate the uppermost layer of soil, extract P, and then by lateral flow return to the soil surface and be lost as runoff. This would suggest that different pools of P can be involved in snowmelt P runoff at different times during snowmelt and in different years.

Climatic indices have been developed for the Canadian Prairies for both snowmelt and rainfall runoff (Eilers and Buckley 2002). Because snowmelt runoff varies from year to year, their snowmelt index includes both magnitude and variability factors. Maps of the relative risk of

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runoff from snowmelt and rain (Figures 3.9 and 3.10, respectively) show that most of the agricultural regions of Manitoba have a high risk of snowmelt runoff and all of the areas have a high risk of rainfall runoff. As a result any method for assessing risk of P loss to surface waters in Manitoba must consider both rainfall and snowmelt runoff processes.

Figure 3.9. Relative risk of snowmelt runoff for the main agricultural regions of the Canadian Prairies (Eilers and Buckley 2002).

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Figure 3.10. Relative risk of rainfall runoff from June 1 to August 15 for the main agricultural regions of the Canadian Prairies (Eilers and Buckley 2002).

2.5 Soils of Manitoba The province of Manitoba covers 647 797 km2 and makes up 6.49% of the area of Canada. Of this area, 553 556 km2 is land and 94 241 km2 is water. The Canadian Shield extends over about 70% of the province and the Interior Plain covers the remaining 30%. Only 72 000 km2 of the Interior Plains region is considered to be arable or potentially arable. Another 5700 km2 of potentially arable land can be found in the “clay belt” of the Canadian Shield region of the province.

Past glaciations and the formation and drainage of glacial lakes, such as Lake Agassiz, have resulted in a wide range in soil parent material and surface topography across Manitoba. Combined with differences in climate, the result is the highly variable group we know as the soils of Manitoba. This group includes clayey soils such as the alkaline and calcareous Osborne clay of the Red River valley and the acidic Keld clay loam of the Parkland region. Sandy soils also range from the alkaline and calcareous (Glenella soil of the Eastern District) to the acidic (Glenboro soil of the Western District). Manitoba also has large areas of organic soils, such as the very acidic Rat River soil of the Eastern District, however these are of much less importance agriculturally than the mineral soils.

Within each of the six rural municipality (RM) districts in the Association of Manitoba Municipalities (AMM), there are some general trends. Despite the fact that many of the soils were formed from calcareous parent material (high pH or alkaline), at least one-quarter of the soil area within each district is categorized as neutral to acidic (pH of 7 or less measured in 0.01 M CaCl2) (Table 3.7). Many of the alkaline soils (pH greater than 7) still contain carbonates in their surface horizons and such soils make up from 23 to 50% of the soil area within each of the RM districts (Table 3.8). The interaction of phosphorus and soil components such as carbonates is affected by particle surface area. Soils with calcareous surface horizons and a texture of clay loam or finer make up a much smaller portion of the districts’ areas ranging from a low of 9% of

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the Western District to 27% of the Central District (Table 3.8). Between municipalities within a district there are also wide differences in dominant soils.

2.5.1 Central District The municipalities of Cartier, MacDonald and Dufferin have the most alkaline soils (over 90% with a pH above 7) and the municipalities of Lorne, Louise, Pembina, South Norfolk, and Victoria have the most neutral to acidic soils (over 40%) (Table 3.9). The RM of Victoria has about 70% of its soil area with a pH of 7 or less (21% below 6). Greater than 50% of the soils in the RMs of Cartier, Dufferin, Morris, MacDonald, Portage and Stanley have calcareous surface horizons (Table 3.10). Only Cartier, MacDonald and Morris have over 50% of their soils which are both calcareous and of a texture of clay loam or finer at the surface. In contrast several RMs have less than one-third of their soil area which is calcareous at the surface and a few, such as North Norfolk, South Norfolk, and Victoria, have less than 15% of their soil area which is both calcareous and of fine texture.

Table 3.7. Distribution of surface soil pH in Manitoba rural municipality districtsz <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 total area percent of soil area ha Central District 0.0 0.0 3.1 26.3 46.0 24.6 0.0 1515545 Eastern District 1.8 17.4 10.0 29.7 26.6 14.5 0.0 1763816 Interlake District 0.2 1.1 0.4 27.8 51.9 18.6 0.0 1489973 Mid Western District 0.0 0.7 4.3 20.8 47.6 26.7 0.0 1473868 Parkland 1.7 1.5 3.3 22.6 43.5 27.3 0.0 1798054 Western District 0.0 0.0 6.8 23.4 39.2 30.0 0.6 2087069 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water Table 3.8. Extent of calcareous surface horizons in Manitoba soils in each rural municipality district Area with soils with a calcareous surface horizon total area any texture texture of clay loam or finer classified District ha % ha % ha Central 733110 47.8 420536 27.4 1534076 Eastern 427305 23.2 214961 11.7 1844273 Interlake 772311 49.8 327208 21.1 1551392 Midwestern 395907 25.0 208089 13.1 1584394 Parkland 839063 44.8 232210 12.4 1874035 Western 508366 23.9 200110 9.4 2130218

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Table 3.9. Distribution of surface soil pH in the Central District of Manitobaz

Rural <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 Municipality percent of RM soil area Cartier 8.4 82.0 8.6 Dufferin 1.2 17.1 49.8 31.4 Grey 0.0 33.5 43.8 22.5 Lorne 2.3 38.7 34.2 22.3 Louise 9.1 33.2 32.9 24.1 MacDonald 8.8 90.3 0.8 Morris 0.0 2.9 94.9 1.7 NorthNorfolk 5.3 24.6 37.0 33.0 Pembina 1.0 44.7 25.2 28.8 Portage 0.2 19.3 35.9 40.3 Rhineland 0.3 31.6 31.1 35.4 Roland 9.9 27.6 52.4 9.5 SouthNorfolk < 0.1 9.3 36.6 28.2 24.8 Stanley 0.1 26.2 37.2 36.0 Thompson 1.0 24.2 48.3 26.6 Victoria 21.3 49.5 15.4 12.6 < 0.1 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water Table 3.10. Extent of calcareous surface horizons in soils of the Central District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finerRural Municipality ha % ha % Cartier 32863 58.8 32144 57.5 Dufferin 52706 57.1 21767 23.6 Grey 47139 48.6 19668 20.3 Lorne 27853 28.7 17439 18.0 Louise 27746 28.5 20775 21.3 MacDonald 62125 53.7 62044 53.7 Morris 66065 63.3 66065 63.3 NorthNorfolk 57601 49.5 2711 2.3 Pembina 42151 36.9 25910 22.7 Portage 139668 66.1 74937 35.5 Rhineland 45593 47.0 22161 22.9 Roland 16417 33.8 12189 25.1 SouthNorfolk 25490 34.1 10136 13.6 Stanley 47228 54.1 14066 16.1 Thompson 24245 45.3 9780 18.3 Victoria 18219 25.7 8743 12.3

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2.5.2 Eastern District The soils of the Eastern District are even more variable. Five RMs have neutral to acidic soil surface horizons in over 70% of their area and nine RMs have over 70% of their soil area with an alkaline surface (Table 3.11). There is also a large difference between RMs when carbonate content of surface soil horizons is considered (Table 3.12). Although many RMs have calcareous surface horizons on about 50% of their soil area some are much lower. Piney and Reynolds, for example, have less than 10% of their area covered by soils with calcareous surface horizons. Several RMs with large areas of soils with calcareous surface horizons have only very small areas which are both calcareous and fine textured at the surface (La Broquerie and Stuartburn, for example). La Broquerie has 41.1% of its area with a calcareous surface but only 0.3% of the area is both calcareous and fine textured.

Table 3.11. Distribution of surface soil pH in Eastern District of Manitoba z <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 percent of RM soil area Alexander 5.5 11.9 28.0 30.8 5.4 1.2 Brokenhead < 0.1 2.9 1.6 17.9 28.2 49.4 DeSalaberry 0.5 15.1 71.7 12.6 EastStPaul 34.6 58.3 0.0 Franklin 0.1 13.8 53.1 32.3 Hanover 0.7 17.2 51.4 29.3 LaBroquerie 0.3 1.2 18.7 21.1 25.2 33.5 LacDuBonnet 0.4 22.6 12.6 42.8 8.6 3.2 Montcalm 4.0 62.7 32.0 Piney 4.1 27.4 23.9 33.8 8.2 1.6 Reynolds 1.6 45.2 7.1 34.5 3.4 2.0 Ritchot 27.4 66.6 4.4 Springfield 1.7 0.1 1.6 17.5 54.6 22.5 StClement < 0.1 3.8 3.8 38.8 19.2 25.2 SteAnne 1.1 3.2 26.8 42.0 26.8 Stuartburn 2.3 2.7 4.3 35.8 27.1 27.5 Tache 0.5 1.0 11.6 72.6 14.1 VictoriaBeach 2.5 68.7 13.6 5.2 8.9 Whitemouth 0.9 31.0 5.8 36.4 19.9 3.8 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water

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Table 3.12. Extent of calcareous surface horizons in soils of the Eastern District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finerRural Municipality ha % ha % Alexander 14897 9.1 7732 4.7 Brokenhead 50052 66.3 37019 49.1 DeSalaberry 39844 59.2 34457 51.2 EastStPaul 1822 41.7 1776 40.7 Franklin 60798 61.2 34578 34.8 Hanover 36367 47.2 16237 21.1 LaBroquerie 23873 41.1 150 0.3 LacDuBonnet 16114 13.4 9155 7.6 Montcalm 24211 49.6 24200 49.6 Piney 21541 8.7 585 0.2 Reynolds 18641 5.2 2522 0.7 Ritchot 15529 45.2 15271 44.4 Springfield 61581 55.8 47413 43.0 StClement 33923 39.7 24686 28.9 SteAnne 21517 44.4 7784 16.1 Stuartburn 47337 40.5 730 0.6 Tache 34213 58.6 26672 45.7 VictoriaBeach 399 18.8 95 4.5 Whitemouth 9438 13.5 3108 4.5

2.5.3 Interlake District The Interlake District is generally thought of as a region of high pH, calcareous soils. Ten of 16 RMs have alkaline surface horizons in over 75% of their soil area and eight RMs have calcareous surface horizons in over 50% of their area (Table 3.13). However, there are also considerable areas with neutral and acidic surface soil horizons with the largest areas being found in the RMs of Bifrost, Fisher, Gimli, and St. Andrews. Fine textured calcareous surface horizons dominate in the RMs of West St. Paul, St. Francois Xavier, and Rosser (Table 3.14). In contrast, fine textured calcareous surface horizons make up less than 5% of the area of the RMs of Armstrong, Eriksdale, Grahamdale, and Siglunes.

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Table 3.13. Distribution of surface soil pH in the Interlake District of Manitobaz <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 percent of RM soil area

Armstrong < 0.1 15.9 74.8 7.0 Bifrost 1.3 7.5 2.6 59.9 24.2 3.6 Coldwell 5.5 36.3 55.2 Eriksdale 19.5 73.6 4.3 Fisher 0.2 0.9 0.7 47.3 42.7 4.3 Gimli 0.3 40.3 48.0 9.5 Grahamdale 0.2 0.9 33.9 58.6 3.2 Headingley 12.3 85.2 1.6 Rockwood 0.2 20.7 40.0 38.7 Rosser 1.2 59.1 39.7 Siglunes < 0.1 24.3 52.4 5.2 St Andrews < 0.1 44.1 40.0 9.0 St Francois Xavier 4.3 84.1 10.8 St Laurent 4.6 33.1 50.7 West St Paul 6.1 69.1 23.9 Woodlands 2.6 33.6 60.0 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water Table 3.14. Extent of calcareous surface horizons in soils of the Interlake District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finerRural Municipality ha % ha %

Armstrong 97323 50.5 7088 3.7 Bifrost 44862 27.1 24806 15.0 Coldwell 80701 87.9 27471 29.9 Eriksdale 38083 46.8 2464 3.0 Fisher 54086 36.1 24626 16.5 Gimli 12264 37.9 4224 13.0 Grahamdale 65438 26.6 7390 3.0 Headingley 5057 46.7 5051 46.6 Rockwood 77120 63.6 46376 38.2 Rosser 30954 69.8 30589 68.9 Siglunes 36198 35.1 689 0.7 St Andrews 31072 37.0 28483 33.9 St Francois Xavier 35220 84.8 35078 84.5 St Laurent 44460 82.3 16878 31.2 West St Paul 6996 79.1 6905 78.0 Woodlands 112477 90.6 59091 47.6

2.5.4 Mid-Western District High pH soils also dominate the Mid-Western District (Table 3.15). Fifteen of 24 RMs have surface pH values above 7 in over 70% of their area and of the remainder several are above 50%.

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The exceptions are the RMs of Clanwilliam, Langford, and Park which have over 50% their soils with neutral to acid surface pH. Despite the relatively high pH values over much of the Mid-Western District, only the RMs of Glenella, McCreary, and Westbourne have over 50% of their soil area with calcareous surface horizons (Table 3.16). Only the RM of McCreary has calcareous and fine textured surface horizons on over 50% of its soil area. Many of the RMs have less than 10% of their area which is both fine textured and calcareous at the surface.

Table 3.15. Distribution of surface soil pH in the Mid-Western District of Manitobaz <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 percent of RM soil area Birtle 2.5 17.0 51.2 29.1 Blanshard 22.6 64.1 13.4 Clanwilliam 0.7 0.1 4.9 47.5 36.2 0.2 Ellice 6.4 42.1 35.3 15.3 Glenella 1.5 43.2 55.1 Hamiota 27.6 66.0 6.1 Harrison 2.8 4.2 31.4 48.4 5.6 Lakeview 4.0 20.8 23.8 Langford 36.6 17.0 31.7 13.5 Lansdowne 3.4 8.3 42.3 45.9 McCreary 0.6 7.4 41.5 49.8 Miniota 12.9 25.1 42.8 18.1 Minto 0.3 22.2 63.6 12.6 Odanah 68.3 31.7 Park-rm 10.2 5.7 49.6 27.6 0.5 Rosedale 0.3 12.0 20.4 28.2 17.5 Rossburn 2.9 3.7 33.2 47.2 10.6 Russell 0.7 25.7 52.1 20.0 Saskatchewan 16.3 63.2 17.8 Shellmouth-Boulton 0.1 23.9 50.0 20.9 Shoal Lake 5.9 49.0 43.2 Silver Creek 0.4 18.4 56.4 22.9 Strathclair 21.4 65.1 10.8 Westbourne < 0.1 7.2 29.9 62.2 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water

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Table 3.16. Extent of calcareous surface horizons in soils of the Mid-Western District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finerRural Municipality ha % ha % Birtle 26711 30.7 20902 24.0 Blanshard 7763 13.4 6574 11.3 Clanwilliam 2290 5.9 1157 3.0 Ellice 15275 26.4 6891 11.9 Glenella 44601 88.7 19350 38.5 Hamiota 3527 6.1 3000 5.2 Harrison 6653 11.5 2791 4.8 Lakeview 54245 43.2 23813 19.0 Langford 12256 21.1 5538 9.5 Lansdowne 58802 76.0 15363 19.9 McCreary 45247 85.7 29136 55.2 Miniota 22345 25.6 11772 13.5 Minto 6278 16.2 2520 6.5 Odanah 12293 31.8 12201 31.6 Park-rm 2916 5.4 1284 2.4 Rosedale 22517 25.9 12034 13.8 Rossburn 10839 14.0 6673 8.6 Russell 15322 26.3 6146 10.6 Saskatchewan 11142 19.2 9736 16.8 Shellmouth-Boulton 17864 30.7 4563 7.9 Shoal Lake 25069 43.2 25069 43.2 Silver Creek 13380 23.0 13123 22.5 Strathclair 7295 12.6 5909 10.2 Westbourne 112343 86.0 47021 36.0

2.5.5 Parkland District All municipalities in the Parkland District have a surface pH above 7 in over half of their soil area (Table 3.17). Alkaline soils make up over 75% of the area in the RMs of Alonsa, Lawrence, Mossey River, and Ste. Rose. Neutral to alkaline soils make up one-third to one-half the area in the RMs of Ethelbert, Minitonas, Shell River, and Swan River. Calcareous surface horizons are found on over 80% of the soil areas of the RMs of Alonsa, Lawrence, and Ste. Rose, but make up less than 10% of the areas of the RMs of Hillsburg and Shell River (Table 3.18). Surface horizons that are both calcareous and fine textured make up less than 20% of the area of any RM in this district with the exception of the RM of Ste. Rose in which fine textured calcareous surface horizons make up 59% of the soil area.

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Table 3.17. Distribution of surface soil pH in the Parkland District of Manitobaz <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 percent of RM soil area Alonsa 11.5 54.4 32.6 Dauphin 2.0 0.1 21.4 32.4 31.6 Ethelbert 1.4 0.8 2.7 34.2 39.5 21.0 Gilbert Plains < 0.1 < 0.1 0.4 19.0 29.5 37.6 Grandview 2.3 0.8 6.4 11.4 26.2 23.5 Hillsburg 2.1 1.1 8.8 29.1 14.9 5.6 Lawrence 7.2 51.2 38.7 Minitonas 3.5 1.8 10.2 22.7 38.7 22.7 Mossey River 0.2 32.4 40.6 25.4 Mountain North 4.1 6.4 1.6 8.2 70.7 7.2 Mountain South 3.8 1.2 1.4 17.7 18.0 7.7 Shell River 0.4 0.2 1.9 35.4 51.1 7.7 Ste Rose < 0.1 0.1 7.5 20.9 67.6 Swan River < 0.1 1.5 11.6 36.7 26.5 22.8 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water Table 3.18. Extent of calcareous surface horizons in soils of the Parkland District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finer Rural Municipality ha % ha % Alonsa 248873 79.1 60016 19.1 Dauphin 86966 56.9 24104 15.8 Ethelbert 50804 44.6 8129 7.1 Gilbert Plains 48006 45.5 12037 11.4 Grandview 23719 19.7 8315 6.9 Hillsburg 4189 5.9 670 0.9 Lawrence 64630 82.7 10475 13.4 Minitonas 47716 39.6 18980 15.7 Mossey River 58772 51.9 5879 5.2 Mountain South 62689 19.3 7318 2.3 Shell River 7012 9.1 881 1.1 Ste Rose 87415 81.2 63900 59.3 Swan River 48272 27.7 11505 6.6

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2.5.6 Western District The Western District is also dominated by alkaline soils with the exception of the RMs of North Cypress and South Cypress, which have neutral to acid surface horizons covering 87 and 77% of their areas respectively (Table 3.19). Calcareous surface horizons are found on about 30 to 45% of the area of most RMs (Table 3.20). Notable exceptions are North Cypress at 16%, South Cypress at 20%, Daly at 68%, Glenwood at 71%, Sifton at 75%, and Whitehead at 71%. Due to the low clay content of many soils in this district, most RMs have less than 20% of their area which is both calcareous and fine textured at the surface.

Table 3.19. Distribution of surface soil pH in the Western District of Manitobaz Rural <4 4.1 to 5.0 5.1 to 6.0 6.1 to 7.0 7.1 to 7.5 7.6 to 8.0 > 8 Municipality percent of RM soil area Albert 0.4 10.3 62.3 24.5 2.0 Archie 0.3 15.3 57.4 26.9 Argyle 0.1 3.0 17.9 37.5 40.2 Arthur 3.0 31.3 38.4 24.2 2.3 Brenda 0.1 24.3 44.6 30.0 0.8 Cameron 12.0 14.5 27.6 42.4 1.9 Cornwallis 18.6 38.4 20.3 16.2 Daly 13.1 16.3 29.0 39.9 Edward 0.2 15.3 53.0 28.6 2.8 Elton 2.1 8.4 53.4 36.2 Glenwood 1.8 11.8 16.7 68.6 0.1 Morton 3.9 23.9 33.1 31.0 0.4 North Cypress 0.3 39.6 46.9 7.6 5.1 0.1 Oakland 20.8 8.7 32.8 37.1 0.3 Pipestone 11.5 71.6 16.7 Riverside 0.5 28.2 52.6 16.9 < 0.1 Roblin 0.1 36.6 32.0 29.5 Sifton 4.4 8.3 5.7 69.2 4.8 South Cypress 18.5 58.9 16.9 4.2 Strathcona 0.6 20.5 37.6 34.2 < 0.1 Turtle Mountain 14.8 75.1 9.7 0.1 Wallace 1.3 10.3 60.0 28.3 Whitehead 9.8 15.0 28.3 45.8 Whitewater 18.4 47.6 32.5 0.8 Winchester 2.1 30.5 25.5 34.0 0.3 Woodworth 5.9 36.7 29.8 25.8 z pH measured in 0.01 M CaCl2 which is typically 0.5 units lower than pH measured in water

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Table 3.20. Extent of calcareous surface horizons in soils of the Western District of Manitoba Area with soils with a calcareous surface horizon any texture texture of clay loam or finerRural Municipality ha % ha % Albert 33194 42.7 6297 8.1 Archie 17585 30.3 13563 23.4 Argyle 68883 42.5 60729 37.5 Arthur 26136 33.6 4278 5.5 Brenda 26029 33.4 7488 9.6 Cameron 35549 45.7 6987 9.0 Cornwallis 19972 34.3 12762 21.9 Daly 39219 67.5 25861 44.5 Edward 35317 45.4 1189 1.5 Elton 26070 44.8 25356 43.6 Glenwood 41238 70.7 18615 31.9 Morton 47707 41.0 22940 19.7 North Cypress 18845 15.6 2961 2.4 Oakland 20641 35.7 15839 27.4 Pipestone 30107 25.9 10861 9.3 Riverside 17918 30.4 642 1.1 Roblin 31271 42.2 5081 6.9 Sifton 65742 75.4 3584 4.1 South Cypress 22018 19.8 13316 12.0 Strathcona 17128 32.0 9413 17.6 Turtle Mountain 30769 32.6 1082 1.1 Wallace 34259 29.4 26940 23.1 Whitehead 41064 70.7 20157 34.7 Whitewater 22791 39.2 7881 13.6 Winchester 30474 39.3 13998 18.0 Woodworth 36394 41.7 26800 30.7

2.6 Methods Of Assessing Risk Of Phosphorus Transfer From Soil To Water The management of P in Manitoba needs to minimize the transfer of P from soil to our surface waters while allowing for adequate fertilization of agricultural fields. The majority of Manitoba soils cannot supply adequate phosphorus (P) for optimum yields and hence some level of P fertilization must be maintained. During the past 25 years, more than 50% of the Manitoba fields tested for plant available P have been rated consistently as very low and low in P (Manitoba Agriculture and Food 1999). The difficulty arises when we try to predict what concentration of soil P poses an unacceptable risk to surface waters.

The amount of total P in soils can vary from 100 to 2500 mg/kg depending upon the soil parent material, texture, and management including fertilizer application and cultivation history (Daniel

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et al. 1994). Of this P, only a small fraction, generally less than 10%, is available at any one time for uptake by plants. It is this plant available P which is also at risk of loss during runoff events or through leaching.

Several methods of assessing risk of P transfer from soil to water have been studied in Europe, the USA, and Canada.

2.6.1 Soil Test Phosphorus Methods of testing soil for P are generally divided into three categories: agronomic P tests, environmental P tests, and total P tests (Table 3.21). The designed purpose of each test is quite different. Environmental soil P tests have been developed to extract that portion of soil P which is easily lost through surface runoff or subsurface flow and therefore use very mild extractants. Agronomic tests have been developed to estimate the amount of P that will be available to a crop throughout the growing season. As a result agronomic P tests use stronger extractants than the environmental P tests. Agronomic P test results are grouped into categories from low to excessive based on the probability of crop yield response due to added P fertilizer in field tests. As a result the P values for the different categories will different for the different agronomic tests and will change with crop and soil type for the same test. Total soil P analysis requires the use of very strong extractants and heating to extract the most recalcitrant forms of soil P. Environmental test methods extract the lowest amounts of P and for most soils both environmental and agronomic test methods extract only a small fraction of total P. For example, a Lakeland clay loam soil from near Winnipeg had extractable P levels of 1400, 120, 105, 60, and 40 mg P/kg soil for total P, Mehlich-3, Kelowna, Olsen, and NH4Cl extractants respectively. An Osborne clay soil had extractable P levels of 790, 22, 20, 12, and 4 mg P/kg soil for the same extractants (M. A. Kashem, personal communications, postdoctoral fellow, University of Manitoba).

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Table 3.21. Common methods of determining environmental, agronomic, and total soil phosphorus levels

Method Extractant Soil to solution ratio (w/v) Extraction time

Environmental

Distilled waterz Sissingh water-Py

CaCl2

z NH4Clx

Fe-oxide stripw

Anion exchange stripsv

Distilled/deionized water Distilled/deionized water

0.01M CaCl2 1.0M NH4Cl

FeO coated paper in 0.01M CaCl2 Bicarbonate saturated anion

exchange membrane in distilled water

1:10 1:4 initially, then 1:60

1:25 1:50 1:40 1:60

1 hr 22 hr

60 min 1 hr

30 min 16 hrs 16 hrs

Agronomic

Olsenx

Bray and Kurtz P-1u Kelownax

Mehlich-3x

0.5M NaHCO3; pH 8.5 0.025M HCl + 0.03M NH4F

0.26N CH3COOH + 0.015N NH4F 0.2N CH3COOH + 0.013N HNO3 + 0.015N NH4F + 0.25N NH4NO3 +

0.001M EDTA

1:20 1:10 1:10 1:10

30 min 5 min

15 min 5 min

Total

Perchloric acid digestiont Sulfuric acid/Hydrogen peroxide/Hydrofluoric acid digestiont

Concentrated HNO3 followed by 60% HClO4

concentrated H2SO4 followed by 30% H2O2 followed by concentrated

HF followed by distilled water

1:10 initially 1:25 in final

solution

1:10 initially, then 1:16, 1:18,

and finally 1:100

OM digested then 30-35 min with

HClO4

Until reaction subsides at each

step z Self-Davis et al. 2000 y Indiati and Rossi 2002 x Kashem et al. 2003 w Chardon 2000 v Tiessen and Moir 1993 u Sims 2000a t O’Halloran 1993 Several studies have shown that the loss of dissolved P in runoff is correlated with soil test P. Using a rainfall simulator on grass plots in Arkansas, Pote et al. (1996) evaluated several soil P tests and found that DRP in runoff correlated best with soil P (0-2 cm depth) extracted by distilled water, NH4-oxalate, and Fe2O3-paper (r2= 0.82, 0.85, and 0.82 respectively). Pote et al. (1999a) also found better correlation between DP losses and STP determined by distilled water or NH4-oxalate than by other STP methods. In Alberta, Wright et al. (2002b) also used rainfall simulators and found that when the manure has had sufficient time to come into equilibrium with the soil, STP alone was an excellent predictor of DRP in runoff for a wide range of soils exhibiting a wide range of runoff and infiltration rates (soils were compared under the same slope and rainfall conditions). STP correlated with P concentration in runoff, but not total P load. Therefore, if it is possible to accurately predict runoff volumes from a field, STP analyses could be valuable for predicting total DRP or BAP loads” (Pote et al. 1996). Predicting the specific

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sources of the runoff within the field is especially critical for manured fields as STP values can range by greater than an order of magnitude (Wright et al. 2002a). A field receiving regular applications of beef manure had modified Kelowna STP values ranging from 150-1659 ppm and the field receiving hog manure had values ranging from 19-299 ppm. In contrast, the field receiving inorganic P fertilizer was much more uniform (37-66 ppm modified Kelowna STP).

P leaching studies have also shown good correlations between STP and P content of leachate (Maguire and Sims 2002b). For many of the soils in this study there was a change point, a STP value below which P loss increased slowly per unit increase in STP and above which leachate DRP increased rapidly. The change point was well above the agronomic optimum STP level. In a review of soil P testing and environmentally based agricultural management practices, Sims et al. (2000a) summarized the results of several researchers to show the consistent pattern that soluble and desorbable P in soil extracts and P in runoff water increase as STP values increase beyond the agronomically optimum range.

However, the relationship between P loss to water and STP varies with soil type, crops grown, and between runoff episodes (Sibbesen and Sharpley 1997). Sharpley (1995) compared runoff DP for several soil types over a range of STP levels (three are shown in Figure 3.11). For example, at a surface soil (0-1 cm) with Mehlich-3 STP threshold of 200 mg/kg, a runoff DP concentration of 280 µg/L would be expected from the San Saba soil but at the same STP level 1360 µg/L would be expected from the Stigler soil. The reasons for the effect of soil type on P-buffering capacity include varying concentrations of Fe and Al hydroxides, clay, carbonates, and organic matter content (Sharpley et al. 1996). The relationship between STP and runoff DP also changes with season/soil moisture status (Pote et al. 1999b). Relationships between DP in runoff and the STP of the surface 2 cm of soils (measured using both water and Mehlich III extracts) were better in May than in August in Arkansas. For soils with the same STP value, the DP in runoff in August was almost double that found in May. Prior to rainfall simulations, the soils were much drier in August and they suspected that P immobilized by soil microbes in May was available for extraction by runoff water in August. Another possible reason was that the fescue vegetation was mostly dry and wilted by August and therefore the vegetative P was more susceptible to leaching by the rainfall. Wright et al. (2002b) reported the opposite seasonal effect with higher flow weighted mean concentrations of DP during spring runoff compared to summer runoff events for manure treatments but not for inorganic P fertilizer treatments (Table 3.22).

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Table 3.22. Effect of manure and season on the concentration of DRP in runoff water for three Alberta soils in 2000 (Wright et al. 2002b) Flow weighted mean concentration (ppm) Treatment Spring runoff Summer runoff Beef manure 10.3 7.3 Hog manure 5.13 2.01 Inorganic fertilizer 0.21 0.27

The depth of soil sampled can also affect the relationship between STP and runoff P loss. Generally, soil samples for agronomic P are taken from the 0-15 cm depth where the majority of plant available phosphorus is found. However, such samples are of limited usefulness for predicting P loss in runoff. Loss of P in runoff begins with desorption, dissolution, and extraction of P from soil, crop residue, fertilizer, and manure as water interacts with a thin layer (often < 2 cm thick) of surface soil (Daniel et al. 1994). The P content measured in the surface 2 cm of a soil can be different from that in a sample which includes the entire 0 to 15 cm depth (Tomasiewicz 2000). Accordingly, it has been recommended that soil samples for environmental STP purposes be taken from the 0 to 5 cm layer as this better correlates to the region that is directly affected by water during a rainfall runoff event (Sibbesen and Sharpley 1997).

Due to the nature of the product, manure application tends to produce greater spatial variability of nutrients than does the application of granular or liquid inorganic fertilizers. This complicates the process of field scale assessment of a STP-P loss correlation in manured soils. Spatial

0

500

1000

1500

2000

2500

3000

0 100 200 300 400

Mehlich-3 P (mg kg-1)

Dis

solv

ed P

( µµ µµg

L-1)

San Saba r2 = 0.90

Shermore r2 = 0.94

Stigler r2 = 0.96

Figure 3.11. Relationship between dissolved P in runoff water and Mehlich-3 STP of the 0-1 cm soil depth (adapted from Sharpley 1995).

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variability of STP for fields treated with manure was much greater than that of a field receiving inorganic P fertilizer (Wright et al. 2002b). Both beef and hog manure treated fields had STP values which differed by an order of magnitude (150-1659 ppm and 19-299 ppm for beef and hog respectively) whereas the STP values of the inorganic P fertilized field ranged from 37-66 ppm.

The measurement of STP for the surface 5 cm of a soil is an important part of estimating the risk of P loss due to runoff but alone cannot give an accurate assessment of the risk of P loss.

2.6.2 Balance of P Removal and Addition Risk assessment based on the method of balancing P removal and addition assumes that any P added in excess of crop export from the field is at risk of loss to surface waters. It is basically a nutrient budget and states that only that P which is removed in a harvested crop or by livestock from that field can be replaced by the addition of manure or fertilizer. It is a very simple method as the only measurements required are yield and P content of the removed produce (grain, straw, hay, cattle, etc.) although regulation and enforcement can be challenging. The Dutch MINAS (Minerals Accounting System) program introduced in 1998 is based on nutrient balance (N and P) with a slight net excess allowed to compensate for unavoidable losses.

There are two main drawbacks to the method. The P budget does not take into account the availability of added P in its original from or after the P has reacted with soil. Therefore, this approach does not consider the portion of added P that is unavailable for plant uptake or loss with runoff or leaching. This method also does not take into account the P fertility level of the soil. Difficulty arises when a field is either deficient in P or has a P content in excess of crop needs. In the former case, crops will be permanently P deficient as only the amount of P in the previously low yielding crop can be replaced by fertilizer or manure P. In the latter case, a field which has an excess of plant available P may be a major source of P to surface and ground water but continue to receive additional P to replace crop removal thereby maintaining high and potentially polluting P levels.

Studying regional P balances can be helpful in flagging regions with large P build-ups. Johnston and Roberts (2001) have prepared such a balance for Manitoba based on census regions (see Chapter 1 for details). Such areas can then be targeted for water protection programs and more field specific methods of risk assessment.

2.6.3 Degree of Phosphorus Saturation The Degree of Phosphorus Saturation (DPS) is an index of the proportion of a soils P retention capacity that is already filled by P. This index was developed in order to account for the effect of soil type on the relationship between STP and water soluble P. The DPS is also known as the degree of P sorption saturation (DPSS) but DPS is the more common term and will be the term used in this review. The basic DPS calculation is given by:

DPS = (soil P / P retention capacity) × 100 For the acid soils of the Netherlands, P retention capacity (also know as phosphate sorption capacity or PSC) can be estimated from the total acid oxalate extractable Fe and Al content of each soil (a pedotransfer function) and oxalate extractable P is well correlated to soil P (van der Zee and van Riemsdijk 1988). For non-calcareous light textured soils, a concentration of 0.1 mg

156

RP/L in groundwater was associated with a DPS of 24% (Brookes et al. 1997). Based on this, over 70% of the soils in the manure surplus regions of the Netherlands are classified as P-saturated (DPS>25%) (Breeuwsma et al. 1997). Leclerc et al. (2001) also reported a very good relationship between water extractable P and the DPS as estimated by oxalate extractable P, Al, and Fe for several Quebec soils (pH 5.4 to 6.8). In Arkansas, for a three year old fescue stand which had received different manure treatments, Pote et al. (1996) found that DPS (again based on oxalate extractable P, Fe, and Al) ranged from 16 to 80% (mean 39%) and correlated to runoff DRP and BAP with r2 of 0.769 and 0.756 respectively. The relationship between runoff DRP and DPS was poorer than between runoff DRP and water, acidified NH4-oxalate or Fe2O3 paper extractable P, but was better than the relationship between DRP and Mehlich-3, Bray-Kurtz P1, or Olsen STP. DPS calculated from Mehlich-3 extractable P, Fe, and Al correlate well with DPS based on oxalate extracts for acid Delaware soils and also exhibited a change point or threshold above which P loss in leachate increased rapidly (Maguire and Sims 2002a).

Other methods of determining the DPS of soils have also been used. Soil DPS values estimated by dividing the oxalate extractable P by estimates of P sorption maximum determined from Langmuir isotherms were highly correlated with runoff DP for a range of soils in the USA (Sharpley 1995; Sims et al. 1998). The acid oxalate extractable P procedure is inappropriate for calcareous soils and a variation of the DPS based on the Olsen P soil test and the soil phosphate sorption index (PSI) was suggested by Hughes et al. (2000). They found that the simple index of Olsen P/PSI determined for the 0-2 cm soil depth was a good predictor of the equilibrium P concentration (EPC) and readily desorbable P for a wide range of arable soils having organic matter contents below 10%. Mozaffari and Sims (1994) also found good correlation between the PSI and the soil P sorption maxima for several Deleware soils. Soil PSI has been suggested as a quick and simple single-point estimate of the P retention maxima for soils (Bache and Williams 1971). Soil PSI was determined from the P retention resulting from the addition of 150 mg P per 100 g soil at a soil:solution ratio of 1:20 after shaking overnight (Hughes et al. 2000). The PSI was calculated as:

PSI = x / log c where:

x = the mg of P adsorbed per 100 g air dried soil, c = final solution P in µmol P/L.

Another variation of the DPS has been used in Florida for calcareous soils (Zhou and Li 2001). The P loss from several calcareous soils was well correlated with the DPS calculated by Olsen-P divided by the amount of P sorbed from a solution initially at 400 µg/mL and added to soil in a 10:1 ratio (10 mL solution to 1 g soil). The relationship was just as good as that obtained using the Langmuir sorption maxima but this estimate of DPS was much easier to determine.

The DPS is an improvement over an STP method in that it accounts for differences in P holding capacities of different soil types and therefore gives a better correlation with runoff DP concentration across soil types. However, it does not assess the risk of water leaving the field and reaching surface waters or groundwater and as a result, on its own, cannot predict DP load losses (Sibbesen and Sharpley 1997).

The Langmuir sorption maxima and the PSI have been determined for a few Manitoba soils as part of P sorption studies, but the majority of Manitoba soils have not been analyzed for PSC by any method. An effective method of assessing risk of P loss from land to water will require the

157

determination of the DPS for the soils present. An effective measure of DPS requires not only a STP level for the surface 5 cm of the soil but also an estimate of the PSC. The Langmuir sorption maxima has been recommended as the best estimate of PSC for a wide range of soils, but is too expensive to be practical (Sharpley 1995). Estimation of PSC from oxalate extractable Fe and Al is effective for acid soils but not for many calcareous soils (Hughes et al. 2000). The PSI, like the Langmuir sorption maxima, has been shown to be an effective estimate of PSC for both acid and calcareous soils (Bache and Williams 1971) and can be used effectively for estimating PSC and runoff DP concentrations (Hughes et al. 2000). For Manitoba soils, DPS based on STP and PSI for the surface 5 cm will be critical to assessing the risk of P loss to surface waters. Research is required to determine the best STP and PSC methods for Manitoba soils.

2.6.4 P Index A national committee of soil scientists in the USA developed the P index in the early 1990’s (Lemunyon and Gilbert 1993). The index includes and ranks transport and source factors controlling P loss in surface runoff (Sharpley et al. 1999b). The original P index used a simple weighted matrix to integrate agronomic STP with other criteria that quantify soil erosion and runoff, as well as P fertilizer/manure rate, method and timing of application. The result was an index to identify soils, landforms, and management practices with the potential to negatively impact water bodies because of P loss from agricultural land (Sims et al. 2000). The original version used eight characteristics and each characteristic was assigned a weighting factor which indicates its relative importance to P loss. Within each characteristic, levels of Low, Medium, High or Very High were set and each assigned a numerical value (1, 2, 4, and 8 respectively). The level and weighting factor for each characteristic were multiplied and then the weighted values for all eight characteristics were summed together to produce a P index for a location (see equation below).

where: NVi is the numerical value of the ith factor and wi is the weight factor for the i factors from 1 through 8. The higher the P index the greater the risk of P loss. The weighting factors were based on the professional judgment of scientists involved and not based directly on empirical evidence. The intent was that individual regions will modify the ratings, values, and weighting factors to suit local conditions. A slightly modified version of the P index (Sharpley et al. 1999b) is shown in Table 3.23. In this version a P index of < 8 has a low potential for P loss under current management practices. A value of 8 to 14 has a medium potential for P loss and some changes should be made to protect surface waters. A value of 15 to 32 has a high potential for P loss and adverse impacts on surface waters. Soil and water conservation measures and P management plans are needed in such situations. Finally a P index of > 32 indicates a very high potential of P loss and adverse impacts on surface waters. In this case, all necessary soil and water conservation measures and a P management plan must be implemented.

Several modifications of the P index have been suggested by other researchers: • Gburek et al. (2000), Heathwaite et al. (2000), and Simard et al. (2001a) recommended

that site factors should not be summed but rather multiplied. This avoids a very high P

∑=

×=8

1iii wNVPI

158

index for an area with a very high STP but little or no erosion or runoff risk and allows for sites with low STP and extremely high erosion, runoff, or leaching potential and possibly very bad management to be identified as high risk.

• Gburek et al. (2000) included the return-period concept into the transport portion of their P index. The return-period factor estimates the risk of runoff from a particular part of the field (i.e. every storm, once a year, once every 10 years, etc).

• Sims et al. (1998) suggested that for regions where P leaching and subsurface lateral flow to streams or drainage ditches are important, the P index should include factors like the soil drainage class, soil texture and cracking potential. Depth to water tables and tile drains could also be important in these soils.

• Soil sampling depth should be reduced from the traditional 15 or 20 cm depth used for agronomic STP determination to the surface 5 cm to properly assess the risk of P loss in surface runoff (Pote et al. 1996; Sibbesen and Sharpley 1997, Torbert et al. 2002). However, if leaching to tile drains or groundwater is a major concern, soil sampling to a depth of 1 meter may be required to accurately assess the risk of P loss (Moore et al. 1998).

• Sims (2000b) recommended the use of environmental rather than agronomic soil test methods. He advocated the use of methods which measure water soluble P, use a P sink (ion exchange membrane or Al- or Fe-oxide paper strip methods), or use a DPS determination to better estimate the amount of P which is at risk of loss to runoff or leaching water.

• Consideration of both pedogenesis and soil pH could improve prediction of soil P sorption capacity based on oxalate extractable Fe and Al (Beauchemin and Simard 1999).

• Clay content and pH were important factors in categorizing Quebec soils by P retention and desorption properties (LeClerc et al. 2001).

An assessment of the risk of P transfer for soils in the Province of Quebec was prepared as an adaptation of the P index and included some of the suggestions listed above (Table 3.24). It includes both STP and an estimate of the DPS as well as rates P addition based upon crop P export, weighting mineral and organic P sources differently. Soil texture and the distance between tile drains are included to account for P loss due to subsurface and lateral flow.

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Table 3.23. The P index (Sharpley et al. 1999b)

Loss rating (value) Site characteristic (weighting factor) None (0) Low (1) Medium (2) High (4) Very high

(8)

Transport factors Soil erosion (1.5) N/A < 5 tons/acre 5 to 10

tons/acre 10 to 15 tons/acre

> 15 tons/acre

Irrigation erosion (1.5)

N/A Infrequent irrigation on well-drained

soils

Moderate irrigation on

soil with slopes < 5%

Frequent irrigation on

soils with slopes of 2 to

5%

Frequent irrigation on

soil with slopes > 5%

Soil runoff class (0.5)

N/A Very low or low

Medium Optimum Excessive

Distance from watercourse (1.0)

> 1000 ft 1000 to 500 ft

500 to 200 ft 200 to 30 ft < 30 ft

Source factors Soil test P (1.0) N/A Low Medium Optimum Excessive P fertilizer application rate, lb P/acre (0.75)

None applied

Placed with planter

deeper than 2 inches

16 to 40 41 to 65 > 65

P fertilizer application method (0.5)

None applied

<15 Incorporated immediately before crop

Incorporated > 3 months

before crop or surface applied

< 3 months before crop

Surface applied to pasture or

applied > 3 months

before crop Organic P source application method (0.5)

None applied

Injected deeper than 2

inches

16 to 40 41 to 65 > 65

Incorporated immediately

before planting

Incorporated > 3 months

before crop or surface applied

< 3 months before crop

Surface applied to pasture or

applied > 3 months

before crop

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Table 3.24. P index modified for the Province of Quebec (Ministere de l’Environnement du Quebec 1998 as summarized by Simard et al. 2001a) Classes (value) Very

low Low Moderate High Very high

Parameter Weight factor

(1) (2) (4) (8) (16)

Site characteristics Erosion (T/ha/yr) (4) 0-3 3-6 6-12 12-18 > 18 Surface runoff (4) More detailed hydrology and management tables are used to determine

this parameter Risk of preferential flow Texture (1.5) Sandy

loam Loam,

silt loam Clay loam,

silty clay loam Medium

sandy loam, clay

Coarse sands, heavy

clay Distance between tile drains (m)

(1.5) No tile drains

> 35 25-35 15-25 < 15

Soil P status Degree of P Saturationz (6) 0-2.5 2.5-5 5-10 10-20 > 20 Mehlich-3 extractable P (kg/ha)

(6) 0-60 60-150 150-250 250-500 > 500

P inputs management Net P added (kg P2O5/ha/yr )y

(3) < -20 -20-0 0-20 20-40 > 40

P added in manure or organic form (% of crop exports)

(2) < 50 50-100 100-150 150-200 > 200

Mineral fertilizer P (% of crop exports)

(1) < 50 50-100 100-150 150-200 > 200

Manure/fertilizer type and incorporation mode

(7) Detailed assessment of risk varies with combination of time of application, incorporation and type of manure

Cumulative Index 36-54 55-108 109-221 222-432 433-576 z as calculated by the ratio of Mehlich-3 P (mg/kg) over Mehlich-3 Al (mg/kg) y balance at the soil surface after accounting for off-site crop exports

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The Province of Ontario also uses a modified P index which is described by Hilborn and Stone (2000). In Ontario, agricultural nutrient management standards require that a P index must be determined if the Olsen STP for a particular field exceeds 30 ppm. The P index was introduced into Ontario with four objectives:

• to rank the relative risk of surface water contamination resulting from phosphorus application on crop land,

• to select management strategies that can be used to reduce this risk, • to determine the distance that phosphorus applications must be set back from

watercourses, and • to set restrictions on rates of phosphorus applied to a field.

The classification and weighting of the various parameters included in the P index are shown in Table 3.25. Soil erosion is estimated by the Universal Soil Loss Equation (USLE) which is based on rainfall, soil, landform, vegetation, and farming practices. The values for the soil erosion component, the method of P application, and P application rate and therefore the STP can all be affected by management. Best management practices (BMPs) can therefore be used to lower the P index for a field over time. The water runoff class is determined by a combination of soil texture and slope and therefore cannot be modified by management practices. Like the previously described P indices, this index uses weighting factors which are multiplied by class ratings to get a value for each parameter and the parameter values are then summed to get a P index value for the field.

The interpretation of the P index includes a general grouping for risk of P loss to surface water and also gives minimum setback distances for P application depending on the rate of P to be applied (Table 3.26). Application method of P (manure or commercial fertilizer) has a larger impact on the Ontario P index than does the rate of P. For example, reducing manure application from 67 to 35 kg P2O5/ha will reduce the P index by 3 points whereas changing the method of application from non-incorporated on bare soil to injection causes a reduction of 10.5 points. The Ontario P index has fewer parameters than the Quebec P index and as a result does not account for subsurface pathways of P transport either by leaching to groundwater or transfer to tile drains. The Quebec model includes an estimate of the DPS which should give it a better capability to assess the risk of both leaching losses and dissolved P losses in surface runoff. Both models try to account for incidental transfer of P directly from manure or mineral fertilizer to water in their runoff and P application method and timing parameters.

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Table 3.25. The Ontario method of determining the P index for a field (Hilborn and Stone 2000)

Classes (rating) Low Moderate High Very high Extreme Parameter

Weight factor (1) (2) (4) (8) (16)

Soil erosion (T/ha/yr)

2.0 <12 12-25 26-37 > 37

Water runoff class within 150 m of a watercourse by texture (slope %)

1.0 Sand (0-5) Loam (<0.5)

Sand (>5) Loam (0.5-5)

Clay loam (<0.5)

Loam (>5) Clay loam (0.5-5)

Clay (<0.5)

Clay loam (>5) Clay (>0.5)

Olsen STP (ppm) 2.0 <15 15-30 31-60 61-100 >100

Commercial fertilizer application rate (kg P2O5/ha)

0.5 <25 25-50 51-75 >75

Commercial fertilizer application method

1.5 Placed with

planter

Incorporated < 2 weeks after application

Incorporated > 2 weeks after application

Not incorporated

Manure/biosolid (organic P) application rate (kg P2O5/ha)

0.5 <12 12-36 37-60 >60

Manure/biosolid (organic P) application method

1.5 Injected in

season

Incorporated within 5 days of spreading

Not incorporated, spread on pre-tilled soil, crop residue or standing crop

Not incorporated,

spread on bare soil

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Table 3.26. Interpretation of the Ontario Phosphorus Index (Hilborn and Stone 2000)

P index for site

Generalized interpretation of P index for site

Minimum setback from watercourse if P is applied up to

crop removal m (ft)

Minimum setback from watercourse if

P is applied over crop removal

m (ft) < 15

LOW potential for P movement from the site. If farming practices are maintained at the current level there is a small chance that P losses from this site will have an adverse impact on surface waters.

3 (10)

30 (100)

15 - 29 MEDIUM potential for P movement from the site. The chance for an adverse impact to surface waters exists. Some remedial action should be taken to lessen the potential for P loss if application is close to a watercourse.

3 (10) 30 (100)

30 - 50 HIGH potential for P movement from the site and for an adverse impact on surface waters to occur unless remedial action is taken. In areas close to a watercourse, soil and water conservation along with P management practices are needed in order to reduce the risk of P movement and water quality degradation.

3 (10) 60 (200)

> 50 VERY HIGH potential for P movement from site and for an adverse impact on surface waters. Remedial action is required to reduce the risk of P movement. All necessary soil and water conservation practices plus a P management plan must be put in place to avoid the potential for water quality degradation.

30 (100) DO NOT apply over crop removal

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The majority of P loss to surface waters comes from only a small portion of the watershed and as a result of relatively few runoff events (Sharpley et al. 1999b). If water or soil does not move from a field or below the root zone, then P will not be lost. Therefore management practices will be most effective if focused on the hydrologically active P source areas in a watershed that operate during the few major runoff events. A P index can help identify these active source areas so that appropriate management practices can be implemented and P transfer from soil to water minimized.

Birr and Mulla (2001) tested a modified P index for use on a regional scale at six river basins in Minnesota. Their P index included many of the parameters used in field scale P indices, however, they were determined for entire watersheds or based upon county mean values. The transport factors used were soil erosion, runoff, and the percentage of cropland and pastureland within 91.4 m of a watercourse. Soil erosion was estimated as the area weighted mean based on the USLE. The runoff value used was the watershed average based on annual discharge. Their source factors included STP (county averages of Bray-1 extractable P), mineral and organic P application rates (estimated from cropland area, fertilizer purchases, and livestock numbers in each county), and application method. The highest risk rating (i.e. worst case scenario) was assumed for application method because it could not be accurately depicted on a regional scale. A strong relationship (r2 = 0.70) between stream monitoring data and P index rating was found in five of the six river basins namely: the Upper Mississippi River, Lower Mississippi River, Minnesota River, St. Croix River, and Cedar River Basins. The relationship between P index values and stream water quality of watersheds in the Red River Basin was not as strong (r2 = 0.51). In comparison to the other basins, the P index values for the Red River Basin watersheds were consistently lower as was the cutoff value (the point at which stream P concentration exceeded 0.25 mg/L). The cutoff P index value was 32 for the other basins and 24-25 for the Red River Basin watersheds. The authors attributed these differences to the highly erodible fine textured lacustrine deposits, high stream bank erosion, and relatively low STP and P application rates of the Red River Basin watersheds which were not accounted for adequately in the current P index model. They conclude that a regional P index can be useful for detecting watersheds that must receive priority in the implementation of BMPs.

Timmons and Holt (1977) found that 80% of the runoff in west central Minnesota (i.e. Red River Basin) was due to snowmelt. As discussed in section 3.4, the interaction between snowmelt runoff and soil can be quite different than that of rainfall runoff. Also due to frozen soils, the runoff source areas can be greater during snowmelt than during rainfall events. Therefore it is not surprising that a P index initially developed to assess risk of P loss due to soil erosion during rainfall events would be less accurate in the Red River Basin than in Basins further south and east. Also, Birr and Mulla (2001) were unable to estimate tillage practices across watersheds and therefore assumed that all fields were under similar practices. As a result they could not account for the increased P stratification that occurs with reduced tillage in comparison to plowing. The main effect of the greater mixing associated with plowing is that a standard agronomic soil test (0-15 cm) is quite representative of the surface 5 cm whereas in reduced tillage systems a 0-15 cm STP value can greatly underestimate the STP value near the soil surface (Tomasiewicz 2000). The STP factor in Birr and Mulla’s P index was based on records of standard agronomic STP values provided by the University of Minnesota’s soil test laboratory. Therefore the STP factor was quite likely too low for any region where reduced tillage was practiced. Moncrief et

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al. (2002) have recently developed a Minnesota P index that includes snowmelt runoff as a separate transport mechanism. Validation studies are not available at this time.

Simard et al. (2001a) proposed a modified P index for the Prairies based on the Quebec model, including a specific parameter for incidental transfer, adding risk of wind erosion transfer, and including grazing intensity as one of the management components (Table 3.27).

The various P index models assign a level of risk of P loss to surface waters from a specific area of land with specific properties and management practices. Their purpose is not to predict the amount of P that will be lost but rather to identify low risk areas and flag high risk areas so that the high risk areas can be managed to minimize P loss. As several regions in Manitoba are faced with net agricultural P surpluses, the development of a P index for Manitoba is essential for the environmentally sound management of our agricultural lands.

Table 3.27. Components for a modified P index for the Prairies (Simard et al. 2001a)

Modes of transport: Risk of water erosion Risk of surface runoff Risk of wind erosion Risk of incidental transfer Surface transfer of manure/fertilizer particles Preferential flow Charge components: STP (soil test phosphorus) DPS (degree of P saturation) Management components: Fertilizer P added (kg/ha) Manure P added (kg/ha) Manure and fertilizer application method Grazing intensity

2.6.5 Phosphorus Transport Models Transport models are designed to be predictive. The major processes and pathways of phosphorus transport are described by equations which attempt to include all pertinent variables and produce an estimate of P loss to surface or groundwater for a selected time period. Several such models have been developed.

2.6.5.1 FHANTM The Field Hydrologic and Nutrient Transport Model (FHANTM) was developed from the hydrologic model DRAINMOD for use in the Lake Okeechobee Basin, Florida (Fraisse and Campbell 1997). It relied on a highly simplified representation of P reactions within the soil and

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has since been modified. FHANTM 2.0 retained the hydrologic components of FHANTM but replaced the P components with those of the GLEAMS model (Groundwater Loading Effects of Agricultural Management Systems) and added the nitrogen components of GLEAMS as well. The nutrient components were further expanded to include additions of P and N by precipitation and to allow for grazed pasture as a management system. This model requires input of weather and climate data, drainage system design, soil properties, soil nutrient content and additions (both natural and applied, organic and inorganic), cropping systems (crop rotation is included and the model also allows for multiple cuts of forage or grazing of pasture), detailed information on fertilizer application methods and tillage operations (type of implement, depth of operation, and timing in relation to fertilizer application). A sampling of input variables is shown in Table 3.28.

Parameters were included based upon an understanding of the underlying mechanisms or as a result of statistical analyses which showed good predictive relationships. For example, the prediction of soil P sorption is based upon clay content, extractable magnesium, organic carbon, and oxalate-extractable aluminum for each soil horizon. However for good predictions the degree of soil weathering must also be accounted for. Calcareous soils required the inclusion of calcium carbonate content and slightly weathered soils required the inclusion of pH and base saturation for each horizon.

Vadas et al. (2002) evaluated a modified version of FHANTM 2.0 for soils from the Mid-Atlantic Coastal Plain. The equations based on extractable soil Mg were replaced by equations using ammonium-oxalate-extractable Fe and Al and organic carbon content for the prediction of P sorption and release and the resultant concentration of P in runoff water. For tilled field conditions under simulated rainfall an additional factor based on eroded sediment load was required to correct for sorption of dissolved P onto sediment particles during runoff. This revised model has not yet been evaluated under natural runoff conditions nor has it been evaluated on soils other than those used to develop it.

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Table 3.28. Partial list of variables used in FHANTM 2.0 for the estimation of phosphorus transport from land to water

Climate Section

Daily potential evapotranspiration (PET) Hourly rainfall Daily maximum and minimum temperature

Drainage System

Depth to tile drain Drain spacing Depth of surface depressional storage Depth to water table Slope percent Slope length Vertical and horizontal hydraulic conductivity Ditch dimensions and weir settings

Soil

Depth from ground surface to the impermeable layer Thickness of each soil horizon Porosity of each horizon or bulk density of each horizon Moisture holding capacities (field capacity and wilting point) Organic matter content of each horizon Double-acid-extractable Mg content of each soil horizon Oxalate-extractable Al content of each horizon Clay content of each horizon Silt content of each horizon Soil pH for each horizon (only required for slightly weathered soils) Base saturation in each horizon (required for slightly weathered soils) Calcium carbonate content in each horizon ( required for calcareous soils)

Nutrient content, application, and incorporation

Initial crop residue on the soil surface N concentration in rainfall P concentration in rainfall Total N and P in the surface horizon Nitrate-N in the surface horizon Potentially mineralizable N Organic N from animal waste in plow layer Labile P in surface horizon (NH4Cl extractable P or estimate calculated from other extraction method) Organic P from animal waste in plow layer Number of fertilizer and animal waste applications Number of tillage operations Tillage mixing efficiency Efficiency of incorporation of surface residue Potential yield C:N ratio of crop N:P ratio for crop Date and type of nutrient applications Method of nutrient application Nutrient content of animal waste applied

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2.6.5.2 NLM The Nutrient Loading Model (NLM) is a relatively simple model developed by ECOMatters Inc. (2002) to assist in setting P application limits. The NLM is based on the Universal Soil Loss Equation to simulate the loss of particle related P and the Baes and Sharp model to simulate leaching of soluble P and nitrates to the groundwater. This model requires the input of several field variables such as soil hydraulic conductivity, percent slope, length of slope, field size, soil bulk density, amounts of manure/fertilizer nutrients added, depth of incorporation, and pre-application nutrient contents. Also required are crop removal estimates and several stream parameters necessary to calculate stream concentrations of P. The NLM requires fewer inputs than FHANTM 2.0; however, it focuses on loss of P by leaching to groundwater and loss as PP in runoff and has only a limited ability to estimate DP losses in surface runoff. The model relies on the USLE and STP to predict PP loss and uses STP and a single partition coefficient to estimate transfer of DP to runoff water. Validation of the NLM on Manitoba data was restricted by the limited availability of P loss data. Using data from the twin watersheds of the South Tobacco Creek study area, the model adequately predicted the loss of total P from both the conventionally- and zero-tilled watersheds. However, it was unable to predict that the P loss from the conventionally tilled field would be mostly PP and that from the zero-tilled field would be mostly DP. The PP is less bioavailable than DP, therefore the ability to assess the risk of DP loss is critical to protecting Manitoba waters.

2.6.5.3 EFPEM The Edge of Field DP Export Model (EFPEM) has been developed out of P mobility studies in Alberta (Wright et al. 2002a). Laboratory rainfall simulations showed good correlation between STP and DPi (dissolved inorganic P) concentration in runoff. Export of DPi from the rainfall simulations could be estimated by:

DPiexp = Q × (0.0031 × STP – 0.0437) where: DPiexp = Total DPi exported in runoff (mg) Q = runoff volume (L) STP = modified Kelowna STP in the 0-5 cm depth (mg/kg) The values obtained from rainfall simulations was compared to runoff data from small catchments and it was found that for equivalent STP values, DPi concentrations were higher in field runoff than in simulations at the same STP level. The differences were attributed to longer runoff path lengths in the field and the contribution to DPi due to interflow. The catchment scaling factor (CSF) was obtained from three catchments under different management. One received beef manure, one hog manure, and the third inorganic fertilizer amendments. The average scaling factor was 5.9 and results in the EFPEM equation for edge of field export:

DPiexp = Q × (0.0183 × STP – 0.258) where: DPiexp = Total DPi exported in runoff (mg) Q = runoff volume (L) STP = modified Kelowna STP in the 0-5 cm depth (mg/kg)

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In the EFPEM, Q is estimated from the WEPP soil erosion model. Successful use of the WEPP model requires digital terrain data, land management data, soils data, and climate data. Although estimates can be used for some parameters, site specific values result in more accurate estimates.

The model has yet to be validated with data from other soils and catchments.

Applying this model and assuming a uniform STP value for the surface 15 cm of soil, to obtain a DPi concentration of <0.05 mg P/L in runoff water, a modified Kelowna STP of <17 mg P/kg soil (78 kg P2O5/ha) would be required. In contrast, an STP level of 78 kg P2O5/ha is at or near the point where winter wheat will no longer respond to P fertilization. Furthermore, fertilizer P is recommended at modified Kelowna STP levels up to 90 kg P2O5/ha for thin Black soils, 80 kg/ha for Dark Brown soils, and 70 kg/ha for Brown soils in Alberta (Alberta Agriculture Food and Rural Development 2000). Therefore, the impact of the EFPEM on predicting the consequences of normal agronomic practices is very significant. In conjunction with the EFPEM, a phosphorus sorption capacity model was developed to give farmers the ability to estimate the amount of P amendment that can be safely added without raising STP levels above a predetermined limit (e.g. 78 kg P2O5/ha). The model predicts final STP levels by multiplying the amount of P added by a soil specific sorption factor (determined experimentally) and adding this value to the initial STP.

2.7 Conclusions “What is now required is to ensure that P, a finite earth resource, is not wasted and that soils do not become so enriched with P that there is an unnecessary risk of too much P being carried to water from agricultural soils” (Higgs et al. 2000).

The capacity of a soil to retain P and the risk of P transfer from soil to water is influenced dramatically by soil type. Most of Manitoba’s agricultural soils have formed on mineral parent material which was produced directly or indirectly by processes associated with glaciation. The result is a wide range of textures and topography. Although much of the parent material was high in carbonates, soil forming processes have resulted in the removal of carbonates from the surface horizon or horizons and a subsequent decrease in soil pH of many Manitoba soils. Many of these soils are now neutral or acidic at the surface. A few of the more acid mineral soils have formed on acid shale parent material. The result is a wide range of pH values in the surface soils with considerable areas having properties associated with reduced phosphorus retention and hence an increased potential for P loss to surface waters. A recent Winnipeg Free Press article (Feb 21, 2003 “Hogs a threat to rural water”) listed the eastern Manitoba RMs of Hanover, La Broquerie and Stuartburn as being first, second, and sixth in the province in terms of manure production per hectare. The RMs of La Broquerie and Stuartburn each have large areas of alkaline and calcareous soils (over 55% alkaline and over 40% calcareous) however, most of the soils are also coarse textured and low P retention capacity would be expected in these soils. Only 0.3% of the surface soils in the RM of La Broquerie and 0.6% of the surface soils in the RM of Stuartburn are both calcareous and fine textured. In contrast, the RM of Hanover has calcareous fine textured surface soils on 21% of its area. The Fe and Al oxide content of Manitoba soils is largely unknown and hence the P retention capacity that they might contribute cannot be estimated at this time.

In addition to soil type, landscape and climate have a large impact on the risk of P transfer to water because of their influence on transport processes. Various hydrological pathways are

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responsible for losses of P to surface waters from agricultural fields. Loss of particulate P by erosion of soil particles during severe rain events is the dominant form of P loss from annual crops under conventional tillage in many areas and has been studied in great detail. In most cases the runoff originates from only a small portion of the field and from only a few storms (Sharpley et al. 1999b). Reducing particulate P loss is possible by adopting conservation tillage, residue management, buffer strips, riparian zones, terracing, contour tillage, cover crops, and settling basins/impoundments but these methods are less effective for reducing dissolved P (Sharpley et al. 1992; Sharpley et al. 1999b). Dissolved P constitutes the majority of the P in runoff from perennial forages or reduced tillage fields and in snowmelt runoff, therefore, reducing losses of dissolved P is critical to reducing P losses to water in Manitoba. Snowmelt runoff P losses have been studied to a much lesser degree than rainfall runoff, however understanding the processes by which snowmelt water extracts P from soils are critical for assessing the risk of P loss from Manitoba soils.

Several methods for assessing the risk of P loss to surface waters have been developed. A soil test phosphorus (STP) limit has the advantage of being very simple (a single measurement) but does not take into account the differences in P retention or P buffering capacity among soils. A simple STP limit system also does not include a transport risk factor. Therefore although an important part of assessing risk, STP alone is not adequate.

The balance of P removal and addition method also has the advantage of simplicity but it would allow a field with excess P to be maintained in that high risk state while requiring a P deficient field to be maintained in its state as well. This method also ignores the irreversible retention of a portion of added P by soil and would, with time, result in ever decreasing available P levels and decrease yields in P limited soils. The P balance is, however, a useful method for identifying regions that could build substantial P surpluses over the long term.

The degree of phosphorus saturation (DPS) method incorporates a measure of STP and a measure of the soil P retention capacity and as a result takes into account the fact that P is more tightly bound (and less likely to be exported by runoff) at low levels of P and that P is bound less tightly as the soil retention capacity becomes saturated. Although an improvement over a strict STP method, the DPS still does not account for the risk of transport from the field and needs to be used in conjuction with an estimate of runoff or leaching.

Phosphorus transport models such as FHANTM and the EFPEM have been useful in predicting P export for the region where they were developed. However, such models require a great number of input variables and although some can be estimated the reliability of the model then depends on the estimate. The large number of climate, soil, and management variables needed make this approach impractical for Manitoba at this time.

The phosphorus index (P Index) is a method that combines source and transport factors into a simple risk index. Different levels of each factor are assigned a risk value and different factors are also assigned a weighting. Unlike the transport models, the P index does not attempt to predict how much P will be exported under specific conditions but rather it rates field locations and practices by their potential to result in P export. As a result, a P-index requires fewer input variables but still accounts for both source and transport factors. The P Index was originally developed as an extension tool but because it is the most robust method of risk assessment it has become the most widely used method. More research and resources will be required to develop an appropriate P Index for Manitoba.

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The development of an effective P Index for Manitoba will require further research in some key areas. Further study is needed on the role of CaCO3 in retention of P applied as manure and the effect of manure composition (especially the cation and anion content since it is affected by diet and livestock class) on P retention by soil. The fate of P added in organic amendments is not as well understood as the retention reactions of inorganic P fertilizers. Research is needed on the effects on soil P retention of organic acids, the cation and ion composition of the manure, pH changes upon manure addition, and the forms of P and their concentrations in the manure.

The P sorption capacity (PSC) of Manitoba soils must be determined so that the DPS can be calculated. As yet there is no recognized method for easily determining PSC and DPS in neutral and alkaline soils. Research in these areas is fundamental for properly rating the P source risk factors. P transport by rainfall runoff has been the focus of a large amount of research and modelling. P transport by snowmelt runoff has received less attention and work needs to be done to determine what soil and landscape properties are important in determining P export by snowmelt.

2.8 Abbreviations and Definitions DP dissolved phosphorus - generally includes colloidal P < 0.45 µm in diameter DPi dissolved inorganic phosphorus DPo dissolved organic phosphorus DPS degree of phosphorus saturation – a measure of the degree of saturation of a soil’s

phosphorus sorption capacity DRP dissolved reactive phosphorus – DP that is measureable by the ascorbic acid-

molybdate method EFPEM Edge of Field dissolved Phosphorus Export Model EPC Equilibrium P concentration HA Hydroxyapatite MCP Monocalcium phosphate MKP Monopotassium phosphate OCP Octacalcium phosphate PP particulate phosphorus – includes all P in the fraction that will not pass a 0.45 µm

filter PSI Phosphorus sorption index (a single point isotherm for estimation of P sorption

capacity of a soil) PSC P sorption capacity RUSLE Revised Universal Soil Loss Equation STP Soil Test Phosphorus TP total phosphorus USLE Universal Soil Loss Equation WEPP Water Erosion Prediction Project

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Critical source areas Areas within a landscape which are the major sources of both runoff and

phosphorus losses to surface waters Hortonian flow Overland flow that occurs when precipitation exceeds the infiltration rate

of the soil and the excess is lost as runoff (infiltration and surface runoff occur simultaneously).

Interflow Flow that travels laterally through the soil Matrix flow Flow through the bulk soil Overland flow Flow that travels over the soil surface Preferential flow flow through soil macropores such as cracks, old root channels, worm

burrows, etc. rather than through the bulk soil. Preferential flow is faster than matrix flow in most soils.

Saturation flow Overland flow that occurs when precipitation falls on a saturated soil and the water is lost as surface runoff because infiltration cannot occur into the saturated soil.

2.9 References Abbott, J. L. and Tucker, T. C. 1973. Persistence of manure phosphorus availability in calcareous soil. Proc. Soil Sci. Soc. Am. 37: 60-63.

Agbenin, J. O. 2003. Extractable iron and aluminum effects on phosphate sorption in a savanna alfisol. Soil Sci. Soc. Am. J. 67: 589-595.

Ajiboye, B., Akinremi, O. O. and Racz, G. J. 2003. Retention characteristics of phosphorus (P) in manured and fertilized soils. Proc. Manitoba Soil Science Society Annual Meetings, Winnipeg, MB.

Akinremi, O. O. 1990. The diffusive transport of phosphate and associated cation in soil and soil-like systems. Ph.D. Thesis, University of Manitoba, Winnipeg, MB.

Akinremi, O. O. and Cho, C. M. 1991a. Phosphate and accompanying cation transport in a calcareous cation-exchange resin system. Soil Sci. Soc. Am. J. 55: 959-964.

Akinremi, O. O. and Cho, C. M. 1991b. Phosphate transport in calcium-saturated systems: II. Experimental results in a model system. Soil Sci. Soc. Am. J. 55: 1282-1287.

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Sharpley, A. N., Kleinman, P. J. A., McDowell, R.W., Gitau, M. and Bryant, R. B. 2002. Modelling phosphorus transport in agricultural watersheds: Processes and possibilities. J. Soil Water Conserv. 57: 425-439.

Sharpley, A. N., Smith, S. J., Jones, O. R., Berg, W. A. and Coleman, G. A. 1992. The transport of bioavailable phosphorus in agricultural runoff. J. Environ. Qual. 21: 30-35.

Sharpley, A., Daniel, T. C., Sims, J. T. and Pote, D. H. 1996. Determining environmentally sound soil phosphorus levels. J. Soil Water Conserv. 51: 160-166.

Sibbesen, E. and Sharpley, A. N. 1997. Setting and justifying upper critical limits for phosphorus in soils. Pages 151-176 in H. Tunney, O. T. Carton, P. C. Brookes and A. E. Johnston, eds. Phosphorus loss from soil to water. CAB International, Wallingford, UK.

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Sims, J. T. 2000b. The role of soil testing in environmental risk assessment for phosphorus. Pages in 57-81. A. N. Sharpley, ed. Agriculture and phosphorus management: The Chesapeake Bay. Lewis Publishers, Boca Raton, FL.

Sims, J. T., Edwards, A. C., Schoumans, O. F. and Simard, R. R. 2000. Integrating soil phosphorus testing into environmentally based agricultural management practices. J. Environ. Qual. 29: 60-71.

Sims, J. T., Simard, R. R. and Joern, B. C. 1998. Phosphorus loss in agricultural drainage: Historical perspective and current research. J. Environ. Qual. 27:277-293.

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3 Legislation and Regulations Regarding Phosphorus Management in Other Jurisdictions

Dr. Ed Tyrchniewicz Department of Agribusiness and Agricultural Economics, University of Manitoba

3.1 Chapter Summary Concerns about phosphorus loadings in Lake Winnipeg have intensified political pressures to bring in phosphorus-based regulations in Manitoba. As has been indicated earlier in this report, the amount of phosphorus discharged into the environment from agriculture appears to be increasing, but it is difficult to separate out agricultural discharges from other sources because of lack of adequate data and monitoring. The purpose of this Chapter is to review legislation and regulations regarding agricultural phosphorus management in other jurisdictions with a view to drawing some conclusions as to appropriate phosphorus management strategies for Manitoba.

This Chapter is presented in three sections. It begins with a step back into why and how regulations are used in society, and specifically the rationale and challenges for environmental regulation in agriculture. This is followed by a review of legislation and regulations relating to phosphorus management in selected Canadian, US, and European jurisdictions. A concluding section summarizes the findings and draws some conclusions and implications.

In the US, federal rules define large "concentrated animal feeding operations" (CAFOs) as point sources of pollution. The USDA and USEPA established the Unified National Strategy for Animal Feeding Operations which sets forth a framework of actions that USDA and EPA can take under existing and legal regulatory authority to reduce water quality and public health impacts from improperly managed animal wastes. Regulations controlling non-point source pollution control is largely left up to individual states.

Regulations were reviewed for four US states that have implemented phosphorus-based regulations - North Carolina, Iowa, Minnesota, and Maryland. North Carolina was reviewed as it is viewed in some circles as the “classic” situation for inadequate environmental regulations. No scientific analysis was found as to the effectiveness of its regulations. Minnesota, Iowa and Maryland regulations are more recent, and all have an emphasis on phosphorus-based regulations.

Iowa and Maryland have both placed considerable emphasis on the use of a phosphorus index. In both case, people involved in the development and application of the P index in those states have expressed caution about the applicability of such an index for developing individual farm nutrient management plans or as an evaluation scale for determining whether land users are complying with water quality or nutrient management standards established by local, state or federal agencies.

In Europe, largely driven by the EU Nitrate Directive of 1991, legislation has been introduced in a number of countries to control application of animal manure. In some of them, e.g. Denmark and England, the primary aim seems to be to control nitrogen losses. In others, e.g. the Netherlands and Ireland, regulations have been broadened to include phosphorus application. There appear to be some questions as to the effectiveness of phosphorus regulations in that farmers are sometimes prepared to pay penalties rather than incur the costs of compliance. In other case, some farmers simply stopped raising livestock.

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Unlike the US and Europe, there are no national level regulations relating to intensive livestock operations, or water quality for that matter, in Canada. In Canada, only Quebec and Ontario have moved towards phosphorus based regulations. Quebec has moved aggressively towards regulating manure application based on phosphorus in regions where the phosphorus buildup has been deemed to be excessive. Ontario is in the process of introducing legislation to do the same. At this point, it is too soon to be able to judge the effectiveness of the Quebec and Ontario approaches.

The link between technical phosphorus risk assessment tools and regulations based on these assessment tools is critical. Choosing an inappropriate risk assessment tool to regulate phosphorus transfer from soil to water will not only add to producers’ production costs, but also may unnecessarily constrain agricultural activities or even be ineffective in ameliorating environmental concerns. The development of an effective P Index for Manitoba will require further research in some key areas before one could be comfortable that it would be an effective regulatory tool.

Before reaching a final conclusion on the applicability of regulating manure management based on phosphorus loadings in Manitoba, it is necessary to investigate the feasibility and effectiveness of several science based options for regulating phosphorus management in Manitoba. At that juncture, it would be useful to apply the criteria mentioned in this chapter to a number of case studies to evaluate the impact of various regulatory options as to their suitability to the Manitoba situation. It would also be useful to explore in more depth and monitor the developments in Quebec, Ontario, Minnesota, Maryland, and Iowa, as well as the Netherlands and Ireland, to assess the ongoing effectiveness of their regulations.

Some regulatory recommendations and regulatory cautions follow.

Regulatory recommendations include: • Given that about 60% of phosphorus loadings in the Red River originate from US

sources, efforts need to be expended by the Government of Manitoba to work with US jurisdictions to reduce phosphorus loadings before they enter Manitoba.

• There is a need for a more collaborative approach among government departments and agencies in the development of a phosphorus management strategy for Manitoba, especially among Manitoba Agriculture and Food, Manitoba Conservation, Environment Canada’s National Water Research Centre, and Fisheries and Oceans Canada’s Freshwater Institute.


• Invest in research to reduce P in manure (e.g. phytase management and other feed additives), before regulating P.





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Regulatory cautions include: • Tighter environmental regulations will impact small-scale farm operators more

negatively than large-scale farm operators. Larger farm operations are in a better position to have the financial resources, technical knowledge, and human resources to know and follow increasingly complex regulations.


3.2 Introduction Concerns about phosphorus loadings in Lake Winnipeg have intensified political pressures to bring in phosphorus-based regulations in Manitoba. As has been indicated earlier in this report, the amount of phosphorus discharged into the environment from agriculture appears to be increasing, but it is difficult to separate out agricultural discharges from other sources because of lack of adequate data and monitoring. The purpose of this Chapter is to review legislation and regulations regarding agricultural phosphorus management in other jurisdictions with a view to drawing some conclusions as to appropriate phosphorus management strategies for Manitoba.

This Chapter is presented in three sections. It begins with a step back into why and how regulations are used in society, and specifically the rationale and challenges for environmental regulation in agriculture. This is followed by a review of legislation and regulations relating to phosphorus management in selected Canadian, US, and European jurisdictions. A concluding section summarizes the findings and draws some conclusions and implications.

3.3 Background to Environmental Regulation in Agriculture In essence, governments introduce policies and regulations into agriculture for one or more of the following reasons:

• To ensure human food safety • To ensure the safety of animals and the environment • To protect consumers from fraud • To help consumers determine the quality characteristics of products • To create trade barriers.

Focusing specifically on environmental issues, there are a number of complications in developing policies and regulations in agriculture (Brinkman 1998). These include:

• Market failures through externalities • Decisions often affect future as well as present generations • Property rights are often poorly defined • Causes, impact and benefits are often difficult to measure

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Externalities are the impacts of actions of producers (or consumers) that do not accrue to the producer (or consumer) but instead spill over onto others, including society in general. An example of this would be the case where a producer is polluting water, but is not charged the cost of cleanup. Thus he has no disincentive to maximize his returns by polluting, even though his actions may have serious negative effects on others.

Environmental and resource issues typically span long time periods. Often these issues involve benefits and costs not only to present producers and consumers, but to future generations as well. This raises difficult questions as to the value to future generations of these benefits and costs. Policy decisions therefore must involve choices for future generations without fully knowing the full implications for resource use.

A significant conceptual issue associated with environmental concerns is that property rights regarding resource “ownership” and use are poorly defined, or may not be defined at all. Property rights represent society’s rules as to what producers and consumers can do with these resources. Environmental property rights are poorly defined as they relate to air and water where private and public uses often conflict.

Environmental issues in agriculture are plagued by a lack of information and difficulties in identifying precise sources of problems. Further complications arise from a number of factors: emissions are often not really monitored, the relationship between production activity and environmental quality are uncertain, and the cause and effect of environmental degradation are often separated by space and time. Polluters often are family-oriented operations with modest returns and a minimal “ability to pay”. A key complication is that the science necessary to formulate regional specific policies and regulations either does not exist or is not relevant.

There is a diversity of policy instruments that have been developed in industrialized countries to deal with environmental matters. Policies and regulations relating to agriculture and the environment can be grouped into three broad categories. These include:

• Regulatory instruments, often referred to as “command and control” • Economic instruments, such as taxes and subsidies • Extension and participatory instruments, emphasizing a voluntary approach

Regulatory and economic instruments can be further classified according to where they are intervening in the production and resource management process. These subcategories would be: designed-based or preventative, and results or performance based. Design based instruments impact the production process directly; they regulate practices, distances, application rates, environmental certification, etc. Results based instruments tend to control emissions, erosion rates, impose penalties in situations of contamination problems, etc. They normally require a monitoring system to control the output and evaluate the impact of agricultural activities.

Historically, policymakers have generally opted for the more traditional ”command and control’ instruments involving explicit limitations on allowable levels of emissions and the use of specified abatement techniques, instead of taxes or “effluent fees” on polluting activities. Pricing measures for the regulation of pollution have been rare. With the decline in publicly supported agricultural extension and research, there has been an increase in producer involvement and leadership in environmental stewardship initiatives.

While a thorough evaluation of all of these categories is beyond the scope of this report, some key challenges associated with these options are presented in the concluding section. For the

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reader interested in a classical treatise on the theory of environmental policies and regulations, Baumol and Oates (1988) provide a theoretically rigorous treatment. A general description and evaluation of policy instruments for environmental protection in agriculture was prepared by the Eastern Canada Soil and Water Conservation Centre (1997).

The most common approaches to regulating phosphorus applications to land are summarized in Landwise (2001, pp 8-9) and are quoted below. These are:

• Water monitoring – limit phosphorus application rates to maintain the phosphorus load or concentration below a critical level in water being discharged from a property. This is the best method for minimizing phosphorus losses from any system. However, it is very difficult and expensive to monitor due to the point and nonpoint nature of phosphorus losses from the land. It is more often used to document performance on a watershed basis.

• Phosphorus balance – phosphorus application rates are limited to levels that meet crop requirements plus an allowable surplus. The Netherlands and Denmark use this approach to ensure that phosphorus does not build up in the soil.

• Soil monitoring – phosphorus application rates are limited when soil phosphorus levels exceed a critical level. Extracting soil phosphorus in order to estimate the upper limit for safe application is complicated because the actual loss of phosphorus from a specific soil will vary with soil characteristics and landscape position. Several approaches have been utilized:

o Single limit approach – Rates of phosphorus applications are reduced or eliminated when soils exceed a single critical value of extractable soil phosphorus. This approach is used in a number of jurisdictions but is not considered to have a strong scientific basis because of the importance of soil characteristics and landscape position on phosphorus losses. However, in regions with relatively uniform soils, this approach could provide a simple mechanism to ensure safe phosphorus levels in the soil profile.

o Percent saturation or threshold approach – The potential for phosphorus to stay in and be leached from most soils increases rapidly after a soil capacity of 25% has been exceeded. Hesketh and Brookes (2000) suggest a simple method of determining the critical value of soil test phosphorus based on the relationship between 0.01 M CaCl2 extractable phosphorus and soil test phosphorus. This approach is most appropriate for phosphorus losses through drainage. It may also be useful for regions with high levels of soluble phosphorus loss in runoff.

o Soil index approach – This approach integrates the risk of phosphorus transport with soil phosphorus test information in order to determine the appropriate areas and application rates. The index approach provides the most flexibility to assess and manage phosphorus losses through surface runoff.

3.4 Review of Regulations relating to Phosphorus Management in Agriculture This section focuses on environmental regulations that impact on animal agriculture in selected jurisdictions including Canadian provinces, the US, and Europe. Jurisdictions were chosen where phosphorus was already included in regulations, or where its inclusion was being contemplated, as well as jurisdictions that are or could be significant competitors with Manitoba’s hog industry. Looking at only phosphorus regulations, however, is not particularly instructive without some

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background to the regulatory approach to nutrients in agriculture in the particular jurisdiction. Accordingly, each review begins with an overview of environmental regulations relating to intensive livestock operations.

3.4.1 Environmental Regulation Affecting Intensive Livestock Operations in Canada There are no regulations or policies at the national level that directly regulate intensive livestock operations in Canada. Accordingly, the focus is only on provincial regulations.

Four provinces were chosen: Alberta, because of its recent new legislation for intensive livestock operations and its position as a significant livestock-producing province; Saskatchewan, because of its potential for livestock production; Ontario, because it is currently revising its regulations, including a focus on phosphorus; and Quebec, because it has recently severely tightened up its regulations on livestock operations, including a focus on phosphorus.

3.4.1.1 Alberta Amendments to the Alberta Agricultural Operation Practices Act were passed in the 2001 fall session of the Legislative Assembly of Alberta. The amendments enhance the province’s ability to deal with nuisance, such as odour, noise, dust, smoke or other disturbances resulting from an agricultural operation. They also provide producers and other stakeholders with a one-window process for the siting of new and expanding confined feeding operations (CFOs). The Act also lays out a set of clear standards for manure storage and application for all farming and ranching operations.

Under the amendments, the Natural Resources Conservation Board (NRCB) is responsible for monitoring compliance and enforcement of province-wide standards. The NRCB will continue to be responsible to the Minister of Sustainable Resource Development. Alberta Agriculture, Food and Rural Development (AAFRD) will be responsible for updating the regulations to ensure they meet the needs of the livestock industry and the public. AAFRD will also take the lead role in providing extension services and technology transfer of applied research to the livestock industry.

Municipalities and counties will not issue development permits for CFOs, but will automatically be notified by the NRCB, which will seek their input on applications for new and expanding operations. They will be encouraged to develop land-use plans that identify where CFOs would not be compatible with current or future land uses, and to provide the reasons why that is so.

Minimum distance separation (MDS) provides an area of separation between CFOs and neighbours. MDS is measured from the outside walls of neighbouring residences (not property line) to the point closest to the applicant’s livestock facility, manure storage facility, catch basin, feeding pen or barn, milking facility or compost area. Setback distances from water bodies and control of run-on and runoff are required. The legislation encourages operators to include odour suppression technology.

Agricultural operations must manage manure in accordance with the nutrient management requirements in the Standards and Administration Regulation (Sections 23 and 25). The rules also apply for composted manure but do not apply to manure to which the Fertilizers Act (Canada) applies The NRCB may authorize a person to apply manure to land in accordance with a nutrient management plan (NMP) proposed by the person if the Board is satisfied that

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following the NMP will provide the equivalent or greater protection to the water and the soil (Standards and Administration Regulation, Section 26).

Operations will have until the end of 2004 to comply with the new manure management standards. Operators must apply manure only to arable land and must maintain a sufficient land base for nutrient management as defined by the standards. Those who apply over 300 tonnes of manure annually are required to perform soil tests and keep records. Records of application and nutrient levels must be kept for five years. Regulations permit the application of manure on frozen or snow-covered ground, as long as the requirements are met for minimum setback distances and slope limitations. Incorporation of manure must occur within 48 hours of application. Operators will have three years to comply with the new manure management standards. More information on environmental regulation of livestock-based agriculture in Alberta can be found at http://www.agric.gov.ab.ca/navigation/livestock/cfo/index.html.

Although there are no phosphorus-based regulations in Alberta currently, the Alberta Soil Phosphorus Limits Project (initiated in 1999) had the objectives of developing soil phosphorus limits for all agricultural lands in Alberta, and identifying phosphorus management options for producers (Olson et al. 2003). Based on a series of case studies, the first phase recommendations of this project are summarized in Landwise (2001):

• Implementing soil phosphorus regulations should include a voluntary education program within a regulatory framework.

• Phosphorus regulations should be variable, depending on soil, climate and landscape conditions.

• Implementation of phosphorus standards should be staged. • Monitoring should be required to ensure phosphorus standards are met. • Implementation of soil phosphorus standards should be combined with a coordinated

nutrient management strategy.

3.4.1.2 Saskatchewan A close working relationship exists between Saskatchewan Agriculture, Food and Rural Revitalization (SAFRR) and Saskatchewan Environment in administering the Intensive Livestock Provisions of the Agricultural Operations Act and the Environmental Assessment Act. SAFRR has included the Environmental Assessment Branch of Saskatchewan Environment in the ILO referral process since 1989. Applications for approval of large livestock operations, or where environmental sensitivities may exist, are provided to Saskatchewan Environment by SAFRR for review.

An intensive livestock operation is any confining of animals where space per animal unit is less than 370 square metres. Approval is required for any intensive livestock operation that has an earthen manure storage area or lagoon, involves rearing, confining or feed of 300 or more animal units, or confines more than 20 animal units but less than 300 for more than ten days in any thirty day period within 300 metres of surface water or 30 metres of a domestic water well

As part of the process to SAFRR approval, applicants must complete a workbook requiring a description of the animals, manure production, storage and utilization, nitrogen, phosphate and potassium production, nitrogen utilization, areas available for manure spreading and management of dead animals. Besides completing the workbook, the farmer must undertake a

http://www.agric.gov.ab.ca/navigation/livestock/cfo/index.html

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geo-technical investigation at the proposed site to ensure the soil and proposed crops are suitable for the amount and types of nutrients in the manure.

Manure must be applied according to an approved waste management plan at rates that supply crop nutrients equal to plant nutrient use. Manure information must specify the form of manure (liquid/solid/semisolid), annual volume or mass of manure, adjusted annual manure N, N utilized annually, and N, P and K yields and concentrations. This will maximize the fertilizer value of the manure and minimize the risk of pollution. The intent is to help the farmer balance nutrient production with crop usage and determine the size of the manure handling facility required.

There appear to be no phosphorus-based regulations in Saskatchewan. More information on environmental regulation of Saskatchewan livestock-based agriculture can be found at http://www.agr.gov.sk.ca/docs/livestock/pork/intensive_hog_operations/ILOreview2002.pdf. It should be noted that phosphorus concentrations and loadings on the Saskatchewan and Assiniboine Rivers near the Saskatchewan border are not as high as in the Red River basin (see Figure 2, Chapter 1). This would suggest that the lack of phosphorus regulations for agriculture in Alberta and Saskatchewan may not be as great a concern for those areas as for the U.S. portion of the Red River basin.

3.4.1.3 Ontario The Ontario Ministry of Agriculture and Food (OMAF) and the Ministry of the Environment are responsible for the regulation for the Nutrient Management Act 2002 that was enacted in June 2002. The Act is a comprehensive, province-wide approach to nutrient management that protects water, the environment and the well being of communities in rural Ontario, while ensuring that farmers can invest in and operate their farms with confidence. As part of the Ontario government's Clean Water Strategy, the Nutrient Management Act provides for province-wide standards to address the effects of agricultural practices on the environment, especially as they relate to land-applied materials containing nutrients.

The legislation provides authority to establish province-wide standards for the management of materials containing nutrients and sets out requirements and responsibilities for farmers, municipalities and others in the business of managing nutrients. The sources of these nutrients include manure and other materials generated through agricultural operations, commercial fertilizers, biosolids generated by municipal sewage treatment and pulp and paper sludge. The land application of these materials is governed by an array of legislative and regulatory provisions, guidelines, voluntary best management practices and a patchwork of municipal by-laws.

This is enabling legislation that supports the implementation of a comprehensive regulatory framework regarding nutrient management and other related farm practices in Ontario. The key to this framework is the Nutrient Management Plan (NMP), which is a science-based tool identifying how manure, commercial fertilizers, other nutrients and existing soil fertility are effectively managed in an environmentally responsible manner. Different types of operations will have different requirements and eventually all land-applied materials containing nutrients will be managed according to NMPs. Generators of materials such as municipal biosolids and pulp and paper sludge, will be required to complete a Nutrient Management Strategy (NMS), which outlines how they are managing materials.

http://www.agr.gov.sk.ca/docs/livestock/pork/intensive_hog_operations/ILOreview2002.pdf

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The legislation also provides authority for clear, strong enforcement. In line with other environmental legislation, provincial government officers who are knowledgeable in agriculture and the environment will have the authority to inspect and issue compliance and preventive orders. The legislation also establishes the right to appeal to the Environmental Review Tribunal. Municipal responsibilities will be clarified under the Act. New standards will replace the patchwork of municipal by-laws regarding nutrient management. Municipalities will have the Act as support for their continued responsibility for land use planning and building code approvals. The Act also allows for the creation of local advisory committees to promote awareness of the new rules, and mediate local nutrient management issues that are not related to enforcement.

Administratively, the legislation provides for alternate delivery of the review and approval of NMPs and for the establishment of a registry for NMPs. It also provides the authority to establish fees for any activity undertaken. Initially, the province will review and approve nutrient management plans and other requirements for large livestock operations. The legislation requires the delivery of enforcement by the Ontario government. The Act re-affirms the ultimate authority of the Environmental Protection Act, the Ontario Water Resources Act and the Pesticides Act. It effects complementary amendments to these Acts, and the Farming and Food Production Protection Act. Different categories of operations will be regulated in different ways, focusing a greater level of attention and resources where the risk to the environment is greatest. The Act provides for a framework to phase in standards over time, depending on the size of the operations and the kinds of practices that are carried out. Any number of sub-categories could also be defined to ensure that different types of operations would be regulated in the most effective way. All farms will eventually be governed by new regulations that incorporate best management practices and standards for the management of materials containing nutrients.

The Act establishes authority for a range of new approval and review requirements designed to minimize environmental risks. These will be most stringent for large livestock operations, which will need provincial certification, including approval for their NMPs. A team of provincial government staff who are knowledgeable in agriculture and the environment will inspect these operations. Mid-size livestock operations wanting to build or expand will be subject to provincial review. These and other agricultural and smaller livestock operations will be responsible for having up-to-date NMPs available for inspection and review. The Act provides authority for several functions including the review and approval of NMPs, education, training and certification.

The second-stage round of consultations took place between August 2002 and February 2003 and focused on proposed requirements regarding:

• Categories of non-livestock, municipal and industrial generators of materials containing nutrients

• Content requirements of nutrient management strategies for municipal and industrial generators; construction and siting of barns and manure storages

• Setbacks and buffers from watercourses for land application • Training and certification for anyone who prepares nutrient management plans and

strategies, as well as haulers and applicators • Quality standards for land-applied nutrients

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• Nutrient management at feedlot operations • Roles and responsibilities of local advisory committees • Winter spreading • Land application near municipal wells • Enhancements to the Ministry of the Environment's land application program.

On March 21 2003, the Government of Ontario announced that as a result of input from the consultations, several changes are being proposed to the government's approach to nutrient management in the province (OMAF News Release March 21 2003). The changes would provide farmers with greater flexibility to comply with standards and maintain the government's key water and environmental protection objectives

The government is proposing to make the following changes regarding the implementation of the regulations under the Nutrient Management Act:

• Making July 1, 2003 the implementation date of the proposed regulations for all new livestock farms and those expanding into and within the large category (more than 300 Nutrient Units). A nutrient unit is the amount of manure that gives the fertilizer replacement value of the lower of 43 kg (95 pounds) of nitrogen or 55 kg (121 pounds) of phosphate.

• Making 2005 the implementation date for existing large livestock farms (more than 300 Nutrient Units).

• Setting up a provincial advisory committee. It would provide recommendations to the government regarding nutrient management issues. The committee would include farmers, environmental scientists, municipal representatives and others.

• Some of the issues that would be referred to this committee for further examination and recommendations are:

o When the proposed regulations would apply to all types of farms except new livestock farms, large livestock farms and those expanding into the large livestock category

o Restrictions regarding the siting and construction of nutrient storage, as well as manure handling and application near municipal wells

o Seasonal outdoor feeding area standards o Manure storage issues for existing operations o Decommissioning of manure storages o Nutrient application on tile-drained land o Nutrient application on shallow soils o Odour-related setbacks and standards o Winter spreading restrictions for nutrients from the pulp and paper sector

• Tying the implementation dates of any future regulations, other than for new and expanding livestock farms, to the availability of cost-shared funding.

• A protocol would be established whereby the Ministry of the Environment would have the ultimate authority to ensure compliance with the regulations through investigations and enforcement.

• The Ministry of Agriculture and Food would be the first point of contact for on-farm nutrient management issues, including monitoring.

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Looking specifically at phosphorus, Ontario Nutrient Management protocols under the Nutrient Management Act specify limits of phosphorus applications based on the Agronomic Balance Calculation for Phosphorus. Agronomic Balance is the total available phosphorus from all applied sources minus crop production requirements. To determine the application limits for phosphorus to a field the farmer must calculate the agronomic balance and, if applicable, the crop removal balance to determine the maximum allowable application rate of phosphorus to a field. If the soil test for phosphorus is greater than 30 mg Olsen-P per kg soil, the P-Index must be calculated to determine required separation distances from water sources. This tool is to be used in the context of nutrient management planning. Ontario's Phosphorus Index was adapted by the University of Guelph from the 1993 U.S. index of Lemunyon & Gilbert, and modified to suit local conditions (Hilborn and Stone 2000). The P-Index can be completed using the Nutrient Management Workbook, the NMAN computer program developed by OMAF, or OMAF Fact sheet 98-079. See section 3.6.4 for more description on the application of the P index in Ontario.

More information on environmental regulation of Ontario livestock-based agriculture can be found in the 2002 Nutrient Management Act and Regulations on-line at http://www.gov.on.ca/OMAFRA/english/agops/index.html. More detailed information on phosphorus and nutrient management can be found in the following sources:

• Ministry Protocols for Ontario Regulation Made under the Nutrient Management Act, 2002, http://www.gov.on.ca/OMAFRA/english/agops/nut_units.htm

• Hilborn and Stone, Determining the Phosphorus Index for a Field, 98-079 http://www.gov.on.ca/OMAFRA/english/engineer/facts/98-079.htm

• The Poop on Phosphorus, November 2000 http://www.gov.on.ca/OMAFRA/english/livestock/dairy/facts/info_poop.htm.

3.4.1.4 Quebec The Regulation for the Reduction of Pollution from Agricultural Sources (new regulations in 1998) was amended with more stringent requirements in 2002. It was accompanied by a moratorium on pig production in certain areas of the province where phosphorus concentration was greater than 20 pounds/hectare. The Regulation respecting agricultural operations replaces the Regulation respecting the reduction of pollution from agricultural sources. Its application also refers to the Environment Quality Act. The purpose of the Regulation respecting agricultural operations is to ensure increased protection of the environment, particularly water and soil, from pollution caused by certain agricultural activities. It focuses on the management standards for manure regarding its storage, spreading and treatment, and on nutrient management, the standards for the location of facilities for raising livestock and for storing manure, as well as livestock accessibility to bodies of water. The new regulation also targets the soil support capacity and the actual fertilizing value of manure with the goal being to reach a nutrient balance between the soil support capacity in phosphorus and the amount of nutrient, especially manure before 2010. In addition to the adoption of a number of new standards, the 2002 regulations embodied a significant shift in approach and philosophy. Specifically, the regulation set fixed environmental standards rather than focusing on specific agro-environmental practices (Bertrand and Beaulieu 2003).

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Some of the provisions of the new regulations that specifically mention phosphorus include: • Operations with solid manure management, whose annual production of phosphorus

exceeds 1600 kg, must have access to watertight storage facilities for all livestock waste produced in them, or to any other equipment or building intended to prevent the contamination of surface and ground waters. This obligation applies as of April 1, 2010 for operations in existence on June 15, 2002, and as of April 1, 2005 for new operations.

• The spreading of manure and other fertilizers is only permitted on already cultivated parcels of land. It must comply with the provisions of an “agro-environmental fertilization” plan established in accordance with the regulation for each parcel of land to be fertilized. The plan must be signed by an agronomist and must list the amount of fertilizers to be used for each of the parcels, the mode of spreading, and the duration and dates of spreading.

• All new facilities, whose annual phosphorus production will exceed 3200 kg, will require a certificate of authorization from the Ministry of Environment.

• An operator must establish an agro-environmental fertilization plan in 2002 if the raising facility has a liquid manure system. Operators of solid manure systems with an annual phosphorus production above 1600 kg have until April 1, 2004 to produce this plan, provided that their phosphorus production does not exceed 3200 kg. The plan must consider the annual phosphorus count by establishing the annual volume of phosphorus production of the herd combined with all other fertilizers, as well as the amount of waste that can be spread on available land, while staying within the maximum levels listed in the regulations.

• An operator must, at least once, arrange for the analysis of the nutrient content of the manure which is produced at his/her facility and which is intended to be spread onto cultivated land; this analysis is not required if the annual phosphorus production of solid manure does not exceed 1600 kg.

• An operator must submit a phosphorus status report no later than June 15, 2003. • An operator whose current raising facility (without an increase in the herd) has above-

limit phosphorus levels, must take steps to reduce phosphorus levels within the following time frames: 50% or more of the phosphorus load by April 1, 2005, 75% or more by April 1, 2008, and 100% by April 1, 2010.

Quebec considered the use of the US based P index, but rejected it because of the following challenges (Bertrand and Beaulieu 2003): the need to calibrate the models as a function of certain soil properties and climate considerations such as precipitation and snow melt, the need to develop computer based tools to facilitate the elaboration of the fertilization plan by agronomists, and the effort needed to train agronomists in the use of the P index. An assessment of the risk of P transfer for soils in the Province of Quebec was prepared as an adaptation of the P index; see section 3.6.6 for more details..

Because of the significant increase in pig farming operations over the past few years in some regions of Quebec and the environmental impacts stemming from these operations, particularly the degradation of the quality of several watercourses and the over fertilization of soils, the government has decided to impose strict limitations on pig farming for 18 months in 281 municipalities. This period will make it possible to carry out complete phosphorus balance checks on all Quebec farms, to compile the results, and, using the data thus obtained, to exercise better management and control over the growth of the pig industry.

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Within these 281 municipalities, there are two zones: limited activity zones and outside the limited activity zones. Some of the regulations that apply within the limited activity zone include:

• No new pig raising facilities will be authorized • In pig raising facilities in existence on June 15, 2002, sow or boar stocks of more than

250 pigs may be produced, provided that the livestock waste undergoes complete treatment and that the resulting products are used outside the limited activity zone

• In pig raising facilities in existence on June 15, 2002, an increase of stocks of up to 250 pigs is permissible, if one of the following conditions is met: complete treatment of livestock waste and utilization of resulting products outside of the limited activity zone, or the cultivated parcels of land are less than 20 km away from the raising location. This authorization is a one-time certificate that must be issued before June 15, 2004, and is valid for one raising facility belonging to one owner.

• Facilities for raising livestock other than pigs can only be authorized if one of the following two conditions is met: the livestock waste undergoes complete treatment and the resulting products are used outside the limited activity zone, or, the livestock waste can be spread on the cultivated parcels of land owned by the farm operator of the raising facility.

Outside of the limited activity zones, the following regulations apply: • Authorizations for new pig raising facilities will only be granted if the waste undergoes

complete treatment and if the resulting products are used outside a limited activity zone • In pig raising facilities in existence on June 15, 2002, sow or boar stocks of more than

250 pigs may be raised, if one of the following two conditions are met: the livestock waste undergoes complete treatment and the resulting products are used outside the limited activity zone, or, the livestock waste can be spread on the cultivated parcels of land owned by the farm operator of the raising facility.

More information on environmental regulation of Quebec animal-based agriculture can be found in Bertrand and Beaulieu (2003), and Regulation Respecting Agricultural Operations – Highlights, online at http://www.menv.gouv.qc.ca/sol/agricole-en/explagriANG6.pdf.

3.4.2 US Jurisdictions From a national perspective, federal rules in the US specifically define large "concentrated animal feeding operations" (CAFOs) as point sources of pollution. This implies that they should be regulated under the same National Pollution Discharge Elimination System (NPDES) that issues permits for industrial and municipal wastewater discharges. A large CAFO under these rules is defined as having more than 1000 cattle, 2500 swine or 10 000 sheep.

From the federal perspective, farms smaller than the limits described above are also assumed to be prohibited from directly discharging pollutants to "navigable waters" (i.e.: acting as point discharges) unless they have a permit to do so. The exception is if the discharge is the consequence of a rainfall greater than the 25-year, 24-hour storm. Federal Law gives individual states the authority to administer a state program related to the permitting of industries susceptible to generating point source pollution. Typically, pollution from agricultural activity is often more appropriately classified as non-point source pollution (i.e., polluted runoff from

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fields, feedlots etc.). Regulations controlling non-point source pollution control is largely left up to individual states.

The USDA and USEPA released the Unified National Strategy for Animal Feeding Operations on March 9, 1999. (http://cfpub.epa.gov/npdes/afo/ustrategy.cfm.). The Unified Strategy sets forth a framework of actions that USDA and EPA plan to take under existing and legal regulatory authority to reduce water quality and public health impacts from improperly managed animal wastes. The strategy still appears to focus on the use of permits in combination with a voluntary approach to pollution control on farms not regarded as "large" (1000 AUs as per NPDES definitions). Comprehensive nutrient management planning (CNMP) is a cornerstone of this voluntary approach. Another key piece is Natural Resources Conservation Service publication 590 entitled National Conservation Practice Standard - Nutrient Management. It can be found at the following website: http://www.nrcs.usda.gov/technical/ECS/nutrient/590.html.

Several states that have a large confined intensive livestock sector and where phosphorus is considered to be a concern have been selected for review. These include North Carolina, Iowa, Minnesota, and Maryland. For general information on state manure management regulations, see http://extension.agron.iastate.edu/immag/ppstothreg.html.

3.4.2.1 North Carolina The livestock industry is regulated largely at the state level in North Carolina. Most environmental regulations are administered by the North Carolina Department of Environment and Natural Resources. In North Carolina, up until 1992, under the state’s non-discharge program, farm operations were "deemed permitted" as long as they were not (knowingly) discharging wastewater to surface waters. In December 1992 a more formal compliance procedure was enacted and these rules (popularly known in as the .0200 rules) became effective February 1993. Under .0200 rules∗ , in order to be "deemed permitted" as per NPDES requirements, some larger facilities had to meet additional criteria including registering their system and completing waste management plans.

Technical specialists through the state soil and water conservation service, cooperative extension service or private professional engineers, provided technical assistance to producers and could certify waste management plans submitted. The deadline to meet the .0200 rules for existing operations was December 31, 1997. Under .0200 rules, an ILO operation (i.e. one that required this extra information) was defined as having greater than 100 cattle, 250 swine, 75 horses, 1000 sheep or 30 000 poultry.

Rapid expansion of the swine industry in the state occurred in the late 80’s and early 90’s. Similarly, large lagoon/sprayfield based manure management systems on large hog farms were established. Serious lagoon spills in 1995 prompted the establishment of the Blue Ribbon Study Commission on Agricultural Waste The Commission sent strong recommendations to the state which were largely adopted and resulted in Senate Bill 1217. In this new bill ILOs were still defined as they were in .0200. State minimum setback distances were established in 1995. Swine barns had to be at least 1500 feet from a residence or 2500 feet from a church, school or hospital and 100 feet from a property line. Liquid hog manure was becoming the focus by this point in North Carolina. Also, the North Carolina Division of Environmental Management, in

∗ .0200 rules refer to the number of the legislation and not to any particular standards.

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Cooperation with the Cooperative Extension Service, was legislated to administer a training and certification program for animal waste managers on swine farms. By January 1998, only a certified operator could apply animal waste to land in North Carolina.

Prior to moratoriums set at the state level, some counties imposed moratoriums on their own. County public health departments may have also enacted ordinances if there was a public health basis to support it. The moratorium on the construction or expansion of swine farms and lagoon/sprayfield style animal waste management systems, established in 1997 and extended until July 2001, was enacted in large part to give counties time to adopt zoning ordinances under G.S. 153A-340. This applies to swine farms having a manure management system having a design capacity of 600 000 lbs Steady State Live Weight (SSLW) or more.

There do not appear to be any specific phosphorus based regulations in North Carolina. For more information on environmental regulations relating to animal-based agriculture in North Carolina, see State Plan for Managing Animal Manures and Animal Derived Nutrients in North Carolina, September 2001. http://www.enr.state.nc.us/DSWC/pages/state%20nutrient%20plan.pdf.

3.4.2.2 Iowa The Iowa Legislature passed a bill in 2002 regulating animal production in the state, with an eye towards increasing environmental protection. The bill addressed a number of issues, including air quality and manure nutrient planning. Implementing the legislation will be a staged process, directed by the Iowa Department of Natural Resources (IDNR) via the rule-making process. It will be July 1, 2007 before all provisions are in place.

Regulation is based on facility size. In the past, there have been essentially two important size thresholds:

• Manure management plans have been required of facilities with 200 000 pounds of bodyweight for swine and poultry, one-time capacity or more, 400 000 pounds for bovine

• Construction permits have been required of facilities with 625 000 pounds of bodyweight (swine and poultry) capacity or more, and 1.25 million pounds for bovine. The new legislation switches from regulating on pounds of bodyweight to animal units (AU). Although AUs are not as precise as bodyweights, they are the method used by the U.S. Environmental Protection Agency.

Two very significant changes in manure management plans (MMPs) under the new legislation include the requirement to submit plans to IDNR annually, and the switch to phosphorus planning. In the past, manure management plans were only required to be submitted to IDNR once - when the facility was constructed. After that any changes were kept in the producers' files, but were not submitted to IDNR. Starting March 1, 2003, plans must be submitted to IDNR annually. Confinement facilities with 500 AUs or more must submit MMPs to IDNR annually. New confinements with 1000 AU or more also must obtain construction permits before building. MMPs must be submitted to all counties in which manure will be applied, as well as the county in which the facility resides.

A significant change in regulations is that IDNR must implement a phosphorus index based on the current Iowa Natural Resources Conservation Service phosphorus index (PI). Information on this index can be found at the following website: http://www.ia.nrcs.usda.gov/Technical/Phosphorus/phosphorusstandard.htm. MMPs will then be based on the PI and associated rules. P planning will be phased in over several years, depending

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on when the first MMP was submitted. If an original plan was submitted before April 1, 2002, the PI will be required by July 1, 2007. If an original plan is submitted April 1, 2002, or after, but before September 1, 2003, the PI will be required by July 1, 2005. If an original plan is submitted after September 1, 2003, the PI will be required in the original plan. It's important to note that the current PI as used by the Natural Resources Conservation Service (NRCS) does not specify P application rates; IDNR will have to develop application regulations based on the PI via the rule-making process.

There is an interesting cautionary note attached to the Iowa P index. The P index is not intended to be an evaluation scale for determining whether land users are complying with water quality or nutrient management standards established by local, state or federal agencies. Use of this P index as a regulatory tool would be beyond the concept and philosophy of the working group that developed it. This P Index has been adapted to local conditions from appropriate regional and available in-state research. This version of the Index should be tested and modified periodically as new research data become available.

More information about the new Iowa legislation, the IDNR rule-making process, and environmental regulations relating to Iowa animal-based agriculture, can be found at the Iowa Manure Management Action Group website http://extension.agron.iastate.edu/immag/.

3.4.2.3 Minnesota Minnesota’s definition of intensive livestock operations proposed under the new rules of a concentrated animal feeding operation (CAFO) is a modification of the federal (US EPA) definition. This affects what facilities are required to apply for and obtain a NPDES permit in Minnesota. Federal regulations basically define a CAFO as having more than 1000 AUs or more than 300 AUs and meeting at least one of two discharge criteria. The proposed rule in Minnesota requires all facilities with 1000 or more animal units comply with the standards and permit application requirements as CAFOs. The proposed new rule would also establish an animal unit threshold at 300 AUs or more to distinguish facilities for purposes of the permitting program and technical standards.

With new rules, additional technical requirements are being placed on farms with lower AUs. For example manure management plans are required for all feedlots with 100 or more AUs. Animals, regardless of numbers present, must be restricted from lakes by October 2001. Animals on pastures (where vegetation is maintained) are prohibited from entering lakes unless an NRCS approved restricted access point is in place.

These regulations are covered in Minnesota Rules 7001, 7002, 7020. (Rules Relating to Animal Feedlots, Storage, Transportation and Utilization of Manure) and are administered by the Minnesota Pollution Control Agency (MPCA). Information about the new rules can be found on the MPCA website at: http://www.pca.state.mn.us/hot/feedlots.html

Some of the regulations adopted for ILOs of 300 or more AUs include: • Operations of 1000 animal units or more must not discharge manure or process

wastewater to waters of the state. No discharge is allowed to a sinkhole, bedrock, well, tile intake, mine or quarry. If the feedlot does not meet the discharge standards, then requirements and timelines will be placed in an individual NPDES permit to correct the hazards.

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• Manure application rates must be limited so that the estimated plant-available nitrogen from all nitrogen sources does not exceed expected crop nitrogen needs for non-legumes and expected nitrogen removal for legumes. The rate determinations are to be based on the most recent publications of the University of Minnesota Extension Service or another land grant college in a contiguous state (some exceptions apply for rates above these levels).

The following testing, planning and record-keeping requirements must be met: • Manure Testing: Manure from all storage areas holding manure from more than 100

animal units must be tested for nitrogen and phosphorus at least annually for the first three years and at least once every four years thereafter. More frequent testing is required when management changes are expected to result in varying manure nutrient content.

• Soil Testing: Soil phosphorus testing is required at least once every four years on fields receiving manure applications.

• Manure Management Plans: Manure management plans must be completed for all operations with 1000 animal units or more upon submittal of a permit application.

• Record Keeping: Records must be kept of manure nutrient test results, field locations, rates and dates of application, available nutrients from manure and fertilizer, and soil test results.

Additional protective measures are required for application of manure in special protection areas, including land within 300 feet of lakes, streams, intermittent streams (excluding grassed waterways), public waters wetlands (e.g. over 10 acres) and drainage ditches without protective berms. Winter application is prohibited in these areas. In special protection areas without a 50 to 100 foot wide vegetated buffer, then the producer must maintain a 25 foot setback, incorporate the manure within 24 hours, and apply in ways that do not result in long-term soil phosphorus accumulation where phosphorus levels are already sufficient for crop growth. Manure must be incorporated within 24 hours if applied within 300 feet of an open tile intake (this does not apply to solid manure until October 1, 2005). Manure must also be incorporated within 24 hours when applied within 300 feet on the upslope side of a sinkhole. A 50-foot setback is required for all sinkholes, wells, mines, and quarries.

Some additional regulations apply to phosphorus specifically. Where manure from any size feedlot is applied in special protection areas to soils that have phosphorus test levels exceeding 21 ppm Bray PI or 16 ppm Olsen, and no permanent vegetated buffers exist along the protected water, re-applications of manure must not occur until phosphorus from the most recent application is removed by subsequent crops as based on soil test results or crop phosphorus removal tables. In the case where manure from feedlots with over 300 animal units is to be applied outside of special protection areas to soils with phosphorus levels exceeding 150 ppm Bray P1 or 120 ppm Olsen, or half of these levels inside Special Protection Areas and within 300 feet of open tile intakes, an interim permit application and manure management plan must be submitted to the MPCA or delegated county, describing how phosphorus is to be managed to minimize losses to surface waters before repeated manure applications occur in such areas. See Mulla (2003) and section 2.6.4 for a more detailed description of the use of the P index in Minnesota. Mulla (2003) states that this P index is not designed to be a regulatory tool; rather, its best use is as a voluntary tool in nutrient management planning.

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3.4.2.4 Maryland In 1998 the Maryland legislature enacted the Water Quality Improvement Act, which mandated nitrogen and phosphorus nutrient management planning for nearly all of Maryland’s commercial agricultural operations (see Coale et al. (2002) for more details). Included under the provisions were all farms with either at least $2500 annual gross revenue or 8 animal units (1 animal unit = 454 kg live weight). The regulations required that a P index be used for determining the potential for P losses from agricultural land to water. At that time, a reliable P index did not exist.

The Maryland Phosphorus Index was developed at the University of Maryland and was published in 2000 (Coale 2000). See the following website for details: http://www.agnr.umd.edu/MCE/Publications/Publication.cfm?ID=538&cat=13. The development and assessment of this index was constrained by a very aggressive implementation schedule imposed by state regulations. According to Coale et al,

“The Maryland PSI will be deployed for use in constructing farm nutrient management plans well before its predictive capabilities can be objectively and rigorously validated. Field validation is essential. In the meantime, the Maryland PSI should function adequately as a tool to assist in the prioritization of field P loss risk potential.”

3.4.3 European Jurisdictions In Europe, farming practices relating to nutrient management have been strongly influenced by the Common Agricultural Policy (CAP) of the European Union (EU). The CAP was originally agreed to in 1957 following World War II when memories of war and food shortages were vivid in the minds of many. The policy had the following objectives:

• To increase agricultural productivity • To ensure a fair standard of living for producers • To stabilize markets • To ensure availability of supplies • To ensure reasonable prices for consumers

These same objectives remain in place today. The policies through implementation effectively produced a bottomless market for agricultural products. The high levels of price support for grains provided a relatively cheap feed for livestock production and thus manure production as well.

In 1991 the European Commission (EC) passed the Nitrate Directive upon realization that the CAP policies had contributed to large surpluses of nutrients as well as food within the EU. This was the first EC legislation that linked agricultural nutrients and water quality and is the overriding European legislation that has the most far-reaching implications for European farmers.

Nitrogen surpluses were causing the greatest concern through its role in contaminating drinking water. There were many examples of drinking water nitrate levels exceeding the World Health Organization’s nitrate limit in drinking water of 50 mg/L. This was resulting in the need for extremely expensive water treatment. Phosphorus was also in oversupply with many soils being over-saturated resulting in phosphorus loadings to both surface waters and ground waters.

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The Nitrate Directive required EU countries to monitor their fresh surface waters and ground waters for one year to identify polluted waters. Those areas that contributed waters whose nitrate concentrations were approaching or exceeding 50 mg/L were classed as vulnerable areas.

The Nitrate Directive also required that EU countries each establish codes of good agricultural practice to reduce nitrate pollution. The codes were to address the following:

• Periods when the land application is inappropriate • The land application of fertilizer to steeply sloping ground • The land application of fertilizer to water-saturated, flooded, frozen or snow-covered

ground • The conditions for land application of fertilizer near water courses • The capacity and construction of storage vessels for livestock manure, including

measures to prevent water pollution by run-off and seepage into the groundwater and surface water of liquids containing livestock manure and effluent from stored plant materials such as silage

• Procedures for the land application, including rate and uniformity of spreading of both chemical fertilizer and livestock manure that will maintain nutrient losses to water at an acceptable level.

The Nitrate Directive also suggested but did not require that codes address the requirement for fertilizer plans and record-keeping systems related to fertilizer use. The Directive also required EU countries to provide training to producers on the implementation of the code requirements.

While member countries had the freedom to develop country-specific action plans for identified vulnerable areas, the plans had to include a limit of 170 kg N/ha (i.e.: max 2 LU/ha) on manure use. For the first 4 years of the program, 210 kg N/ha was allowable to help phase in this requirement. How effective the nitrate directive has been is not clear. Initially, there was strong resistance by a number of the member countries to execute the Directive. Therefore it has not been acted upon as effectively as it could have been. Nevertheless, there still remains a significant nitrate pollution problem in all EU countries. More information on and assessment of the European approach to nutrient management regulation can be found in Aldinger (2001), and Sibbesen and Runge-Metzger (1995).

The following sections touch on some the activities in selected countries (The Netherlands, Denmark, England and Ireland) and will help to illustrate how European countries are addressing the issue of nutrient management, particularly as it relates to intensive farming activities and phosphorus. Legislation has been introduced in a number of European countries to control application of animal manure. In some of them, e.g. in Denmark, France and Germany, the primary aim seems to be to control nitrogen losses. In others, e.g. the Netherlands and Belgium, legislative regulations are based on P addition.

3.4.3.1 The Netherlands Numerous Acts have been passed since the mid-eighties to try and address the excess minerals (nutrients) problem. The first law (1983) was the Interim Law for Restricting Pig and Poultry Farms. It prohibited the startup or expansion of such farms in sandy soil regions but proved ineffective in preventing increases in animal numbers. The Manure Law and the Soil Protection Act replaced it and introduced manure bookkeeping. It took a 3-phase approach. Phase III of this approach however was proving to be unrealistic in trying to meet the EC’s Nitrate Directive.

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In general, the program to reduce N and P losses in the Netherlands relies on the following basic components:

• Detailed minerals accounting on farms with high livestock density. • Establishment of mineral use and loss standards to be achieved. These are phased in

through to 2010. • Encouragement and restructuring • Reduction of ammonia losses

Ammonia emission, and related acid rain concerns, is as much of a problem in the Netherlands as nitrates. Programs are in place requiring manure injection at spreading, covering storage facilities and building low emission facilities. Producers have been very successful in reducing the water volumes in their liquid manure. This makes for reduced transport costs. They are also taking full advantage of phytase additives in feed.

The Minerals Accounting System (MINAS) forms the core of Dutch nutrient management regulations and policy. MINAS is essentially a registration system where farmers keep a detailed record of nitrogen and phosphate inputs and outputs on their farm. The difference between inputs and outputs reflects the farm’s minerals surplus. Farmers are required by law to complete an annual minerals return form. This system sets restrictively high levies on phosphate and nitrogen surpluses above a certain maximum allowed standards for nitrogen and phosphate per hectare (loss standard), effectively forcing farmers to take measures to minimize mineral losses to the environment.

The system was first made compulsory for intensive livestock farms in 1998; it was extended to all farms in 2001. In 1999, the government decided to progressively reduce the maximum allowable standards to reach their final level in 2003. Because dry sandy soils and loess soils are most vulnerable to nitrate leaching to groundwater, stricter standards apply to these soil types, so that environmental targets may also be met there.

In 2002 the Dutch government introduced a manure transfer contract system to complement MINAS. Together, the two systems should ensure that the volume of manure deposited on land is always equal to what the land can take. Under the manure transfer contract system, farmers must plan manure transfer before it is produced. Farmers can apply manure on their own land provided they do not exceed the MINAS loss standards. Surplus manure can be delivered to crop farmers or livestock farmers with a manure shortage, or to manure processors. Farmers who are unable to dispose of all of their surplus manure in this manner would be faced with the prospect of paying high levies or cutting down on the number of livestock. Many farmers do not meet the MINAS standards and a number of farms prefer to pay the levy, as the alternative is more costly.

More information on the “Dutch Approach” to nutrient management can be found in Ministry of Agriculture, Nature Management and Fisheries (2001). An assessment of the administrative costs can be found in Jacobsen (2002).

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3.4.3.2 Denmark The EC Nitrate Directive influences Denmark’s policy, like the Netherlands. Therefore, there is a lot of similarity between Denmark and other EU countries and the Netherlands. Similarities between the Netherlands and Denmark’s situation are as follows:

• Danish farmers are also subject to a deterrent levy if they exceed a fertilizer quota assigned to each farm.

• Denmark also corrects nitrogen losses in animal manure. • Denmark also tests newly formed groundwater (i.e.: groundwater just below the root

zone) as the point to see whether the nitrate standards are exceeded. • Their target reduction for nitrogen emissions to the sea is also 50%. Denmark expects to

reach this target in 2003.

Differences between the Netherlands and Denmark are as follows: • Denmark has implemented a more restrictive input standard for animal fertilizers: for

cattle maximum 170 kg N/ha, linked to 1.7 LU. • There is an exception for a small percentage of the cattle herd: 230 kg nitrogen/ha • Danish farmers have to set up a fertilization plan, taking account of nitrogen use and

lower yields. • The farms are generally tied to the land and must dispose of their manure surpluses in the

neighbourhood of the farm on a contract basis. • Denmark has been much more successful in establishing biogas plants.

The Danish Aquatic Programme I was implemented in 1987 with the aim of reducing nitrate leaching by 50%. Although a series of policy measures, including when to apply animal manure, storage of 9 months and fertiliser accounts, has been implemented, the aim has still not been achieved. Until 1997 the N-surplus has been reduced only by 20%. The Danish Aquatic Programme II was therefore necessary. For the first time the program did not just contain measures aimed only at reducing N-leaching. A wider range of measures was introduced. The measures can broadly be divided into farm measures affecting the nitrogen use on the farm and general state measures, such as making more wetlands and plantation of forest aimed at reducing N leaching as well as improving the landscape.

Earlier legislation in the 1980s was the mandate of the country’s environment ministry. In 1990, however the mandate was turned over to the agriculture ministry. Manure spreading in Denmark is prohibited from harvest until February 1. This results in a need for farmers to have 9 months of storage. Sine 1994, a statement has to be filed with the municipality stating that sufficient storage is available. Liquid manure tanks must have a cover. A floating straw cover is commonly used.

Beginning in 1997, the regulations required that manure storage be checked every 10 years for strength and tightness. A minimum 5 year written contract is required if the storage capacity is achieved by using storage facilities on other farms or at biogas plants. All farms exceeding 10 hectares are required to produce compulsory fertilizer and crop rotation plans. As well, 65% of the cropland must be sown to a green cover crop in the autumn.

Biogas production has proved successful at several large plants. Raw material includes both manure and municipal waste. Farmers deliver their manure directly to the plants which then process and store the manure and processed waste. Then from mid-March to mid-May, the plant delivers the processed waste to the edge of area fields in special trailers. The farmers then apply

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this material that contains 3% to 4% total N and 2% to 3% total P. Odors are very low. But there are sometimes odour problems around the biogas facilities themselves that raise concerns.

Although Denmark has been aggressive in dealing with controlling nitrates, little emphasis has been placed on phosphorus regulations.

The material on Denmark’s regulation of animal agriculture is based largely on information in the OMAF website, Nutrient Management in Other Jurisdictions, and can be found at http://www.gov.on.ca/OMAFRA/english/agops/nutmgtlinks.htm.

An assessment of the effectiveness of these regulations can be found in Jacobsen (2002), and Aldinger (2001).

3.4.3.3 England Information on England was drawn largely from the OMAF website: http://www.gov.on.ca/OMAFRA/english/agops/nutmgtlinks.htm, and the Department for Environment, Food, and Rural Affairs (DEFRA) Guidelines for Farmers in Nitrate Vulnerable Zones (UK DEFRA 2002). England established a Pilot Nitrate Scheme in 1990 in response to concerns over nitrate contamination of groundwater. This directive, while not initially in response to the EC’s Nitrate Directive, had many parallel features and is the basis of the program now in place to help England meet the EC Directive. The Pilot Nitrate Scheme was replaced in 1995 with the Nitrate Sensitive Areas (NSA) scheme. A total area of 35 000 ha within 32 NSA’s were affected and these areas all fell within the category of Nitrate Vulnerable Zones; these are areas with high levels of nitrate in surface and ground waters as classified under the EC Nitrate Directive.

The program in England to reduce mineral losses from their land has been largely voluntary. There are, however, significant government financial incentives to encourage participation. The three types of voluntary land use measures farmers can participate in are as follows:

• Premium Arable Scheme – which involves the conversion of arable land to extensive grass under a range of possible management schemes that have low fertilizer input.

• Premium Grass Scheme – which involves extensification of intensively managed grass. • Basic Scheme – a scheme involving low nitrogen input arable cropping. Low nitrogen

input refers to an average maximum of 150 kg/ha/year fertilizer additions.

Farmers entering their fields under these programs agree to stay in the program for a minimum of 5 years. They also agree not to remove fencerows, hedgerows or destroy other traditional and environmental features from the area. Payments are made in exchange for their voluntary participation.

In response to growing concerns about nitrate buildup in large parts of the country, the UK Department of Environment, Food, and Rural Affairs published a guide for farmers containing rules to be followed in manure and fertilizer applications in Nitrate Vulnerable Zones in May 2001 and revised in July 2002 (UK DEFRA 2002). In 2002, these zones covered about 47% of the country; in 1996, they covered only 8%. These guidelines cover such subjects as periods during which spreading is prohibited, nitrogen limits, spreading controls, slurry storage, and record keeping. A significant departure from past approaches was to move away from incentives, and towards “command and control” regulations. DEFRA does not regulate or even mention phosphorus in its rules or guidelines.

http://www.gov.on.ca/OMAFRA/english/agops/nutmgtlinks.htm


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3.4.3.4 Ireland A brief reference is made to Ireland in that the approach has been one of guidelines for nutrient management, but increasingly there is pressure to move towards regulations, especially as “spread lands” become scarcer.

A variety of factors are influencing the availability of spread lands. These include revised P nutrient advice for grassland, the Environment Protection Agency BATNEEC Guidelines, the national Code of Good Agricultural Practice to Protect Groundwater from Pollution by Nitrates, and the national Rural Environment Protection Scheme (REPS) all of which influence the supply of agricultural land that can be utilized for recycling of manure generated by intensive agricultural enterprises (IAEs). IAE manure cannot be applied to land where the existing soil test phosphorus (STP) level exceeds 15 milligram per litre (mg/L) (according to Morgan's Test). Enrolment of farmland into the Department of Agriculture and Food's REPS is significantly reducing the availability of suitable spread lands. The mandatory upper limit of organic N, 170 kg/ha/annum, for farmers participating in REPS constrains their potential to receive IAE manure

The EPA Guidance Notes establishes a STP limit of 15 mg/L for spread lands associated with intensive pig and poultry units. Soils with P levels greater than this are not permitted for the application of IAE manure. The EPA limit is based on the 1994 agronomic STP level for silage production above which no P was recommended. This is significantly lower than the suggested Teagasc STP limit of 30 mg/L on IAE manure spread lands with low vulnerability for nutrient losses. This higher STP limit was proposed to accommodate the special circumstances of IAE nutrient surpluses. However, since then more recent research has demonstrated a relationship between STP and P loss to water. This evidence, coupled with the demands for more sustainable nutrient management practices, has negated the applicability of the 30-mg/L STP limit.

It should be noted that the agronomic STP targets for crop production based on Teagasc fertilizer guidance should be adopted on tillage and grassland farms (i.e. nutrient deficit farms). Applications of P to land with STP in excess of these targets are not recommended in most situations, as no agronomic benefits will result.

Teagasc has developed a four-category index system as the basis for fertilizer P recommendations for tillage crops and grassland. The basis of the Teagasc system is a set of soil indices based on the extractable P measured (by the Morgan’s P chemical test) in the soil and the crops response to fertilizer applications as measured by field experimentation.

Although the storage and application of organic wastes is covered under existing legislation, it appears that this legislation is not effective in controlling the loss of nutrients to waters from agriculture. Controls that already exist include the provision for making bylaws under section 21 of Water Pollution (Amendment) Act, 1990; requirement to undertake nutrient management planning under section 21 A of Water Pollution (Amendment) Act, 1990; EPA Act, 1992, Waste Management Act, 1996; the designation of Nitrate Vulnerable Zones under the Nitrates Directive (Council Directive 91/676/EEC) which will result in the Code of Good Agricultural Practice becoming mandatory in these areas; and the controls which exist for farms participating in REPS.

Proposals need to be developed that specifically address the need for the provision of adequate winter storage of organic wastes on farms and the subsequent land spreading on agricultural lands in accordance with nutrient management planning. The types of land spreading equipment

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used and some land spreading practices should also be addressed e.g. the use of umbilical system to apply slurry during winter months on wet soils and spraying slurry off roads and farm lanes using high trajectory spreading equipment.

Details of these guidelines can be found in Carton and Magette (1999), and Brogan et al.(2001).

3.5 Conclusions This section summarizes key observations that may be relevant to Manitoba’s approach to regulating phosphorus management. It will then go on to drawing some broader conclusions as to an appropriate approach to regulating phosphorus in Manitoba in the context of the conceptual ideas mentioned in Section 4.2.

3.5.1 Summary Starting with a summary of the overall approach to the regulation of intensive livestock operations in selected jurisdictions, we move on to focusing specifically on phosphorus.

3.5.1.1 National level regulations of intensive livestock operations In the US, federal rules define large "concentrated animal feeding operations" as point sources of pollution. This implies that they should be regulated under the same National Pollution Discharge Elimination System (NPDES) that issues permits for industrial and municipal wastewater discharges. The USDA and USEPA established the Unified National Strategy for Animal Feeding Operations which sets forth a framework of actions that USDA and EPA can take under existing and legal regulatory authority to reduce water quality and public health impacts from improperly managed animal wastes. Regulations controlling non-point source pollution control is largely left up to individual states.

The European Commission (EC) established the Nitrate Directive in 1991 upon realization that the CAP policies had contributed to large surpluses of nutrients as well as food within the EU. This was the first EC legislation that linked agricultural nutrients and water quality and is the overriding European legislation that has the most far-reaching implications for European farmers. Individual countries then took a variety of approaches, with varying degrees of success, to meet the requirements of the Nitrate Directive.

Unlike the US and Europe, there are no national level regulations relating to intensive livestock operations, or water quality for that matter, in Canada. The only federal jurisdiction relating to the impact of agricultural activities on water quality is the Department of Fisheries and Oceans regulation of water for fish habitat.

3.5.1.2 Phosphorus-based regulations Overall, the most common approaches to regulating phosphorus applications to land (Landwise 2001) include:

• Water monitoring – limit phosphorus application rates to maintain the phosphorus load or concentration below a critical level in water being discharged from a property. This is the best method for minimizing phosphorus losses from any system. However, it is very difficult and expensive to monitor due to the point and nonpoint nature of phosphorus losses from the land. It is more often used to document performance on a watershed basis

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• Phosphorus balance – phosphorus application rates are limited to levels that meet crop requirements plus an allowable surplus. The Netherlands and Denmark use this approach to ensure that phosphorus does not build up in the soil.

• Soil monitoring – phosphorus application rates are limited when soil phosphorus levels exceed a critical level. Several approaches have been utilized:

o Single limit approach – Rates of phosphorus applications are reduced or eliminated when soils exceed a single critical value of extractable soil phosphorus. This approach is used in a number of jurisdictions but is not considered to have a strong scientific basis because of the importance of soil characteristics and landscape position on phosphorus losses. However, in regions with relatively uniform soils, this approach could provide a simple mechanism to ensure safe phosphorus levels in the soil profile.

o Percent saturation or threshold approach – The potential for phosphorus to stay in and be leached from most soils increases rapidly after a soil capacity of 25% has been exceeded. Hesketh and Brookes (2000) suggest a simple method of determining the critical value of soil test phosphorus based on the relationship between 0.01 M CaCl2 extractable phosphorus and soil test phosphorus. This approach is most appropriate for phosphorus losses through drainage. It may also be useful for regions with high levels of soluble phosphorus loss in runoff.

o Soil index approach – This approach integrates the risk of phosphorus transport with soil phosphorus test information in order to determine the appropriate areas and application rates. The index approach provides the most flexibility to assess and manage phosphorus losses through surface runoff.

Regulations were reviewed for four US states that have implemented phosphorus-based regulations - North Carolina, Iowa, Minnesota, and Maryland. North Carolina was reviewed as it is viewed in some circles as the “classic” situation for inadequate environmental regulations. No scientific analysis was found as to the effectiveness of its regulations. Minnesota, Iowa and Maryland regulations are more recent, and all have an emphasis on phosphorus-based regulations.

Iowa and Maryland have both placed considerable emphasis on the use of a phosphorus index. In both cases, people involved in the development and application of the P index in those states have expressed caution about the applicability of such an index for developing individual farm nutrient management plans or as an evaluation scale for determining whether land users are complying with water quality or nutrient management standards established by local, state or federal agencies.

In Europe, largely driven by the EU Nitrate Directive of 1991, legislation has been introduced in a number of countries to control application of animal manure. In some of them, e.g. Denmark and England, the primary aim seems to be to control nitrogen losses. In others, e.g. the Netherlands and Ireland, regulations have been broadened to include phosphorus application. There appear to be some questions as to the effectiveness of phosphorus regulations in that farmers are sometimes prepared to pay penalties rather than incur the costs of compliance. In other case, some farmers simply stopped raising livestock.

In Canada, only Quebec and Ontario have moved towards phosphorus based regulations. Quebec has moved aggressively towards regulating manure application based on phosphorus in regions where the phosphorus buildup has been deemed to be excessive. Ontario is in the process of

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introducing legislation to do the same. At this point, it is too soon to be able to judge the effectiveness of the Quebec and Ontario approaches.

3.5.2 Assessment of Regulatory Approaches As mentioned at the beginning of this chapter, policies and regulations relating to agriculture and the environment can be grouped into three broad categories. This review has identified all of these, often in combination. These include:

• Regulatory instruments, often referred to as “command and control” • Economic instruments, • Extension and participatory instruments, emphasizing a voluntary approach

Regulatory and economic instruments can be further classified according to where they are intervening in the production and resource management process. These subcategories would be: designed-based or preventative, and results or performance based. Design based instruments impact the production process directly; they regulate practices, distances, application rates, environmental certification, etc. Results based instruments tend to control emissions, erosion rates, impose penalties in situations of contamination problems, etc. They normally require a monitoring system to control the output and evaluate the impact of agricultural activities.

Historically, policymakers have generally opted for the more traditional ”command and control’ instruments involving explicit limitations on allowable levels of emissions and the use of specified abatement techniques, instead of taxes or “effluent fees” on polluting activities. Pricing measures and financial incentives as instruments for the regulation of pollution have been rare.

With the decline in publicly supported agricultural extension and research, there has been an increase in producer involvement and leadership in environmental stewardship initiatives. The work of the Manitoba Pork Council in environmental education for its members is a good example of such an approach. Another example is the New York State Agricultural Watershed Council. This is a voluntary venture between farmers and foresters in the New York City watershed and the City of New York to establish water standards that include best management practices that protect water quality and enhance economic viability (Coombe 2003).

To design the “ideal” system of regulating nutrient management in agriculture, let alone phosphorus in Manitoba, is well beyond the scope of this project. However, some general directions can be offered.

3.5.2.1 Assessment criteria In attempting to devise or recommend a set of policies for regulating nutrient management based on phosphorus, the multiplicity of social, economic, and natural realities calls for a framework that must satisfy a wide set of criteria. The Eastern Canada Soil and Water Conservation Centre (1997) was responsible for putting together a comprehensive review of policy instruments environmental protection in agriculture.

Their assessment of a broad range of policy instruments was based on the following criteria. These criteria have received elaboration where appropriate.

Equity criteria - Equity should not be interpreted as equality. • Equity among farmers - All farmers (both within and among regions) must be treated

equally with respect to applicability of regulations, and have the same access to support

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programs and costs for environmental protection. Zoning regulations affect landowners on an unequal basis, depending on location (close to water courses or neighbours). While policy makers (and farmers) are not responsible for this situation, compensation measures or other considerations may be needed to meet the equity criteria.

• Internalization of externalities - The costs of environmental protection and degradation should be borne by resource users. Subsidies for environmental protection imply a priority right for landowners over public resources such as water.

• Rural/urban equity - Policies and regulations, and planning/management processes should take into account the demographic unbalance between rural and urban as well as between farm and rural nonfarm. The off-farm impacts of agricultural activities, both positive and negative, should also be considered.

Implementation and enforcement efficiency - Costs and enforceability are very important criteria for policy efficiency. The efficiency will be low for policy instruments focusing mainly on legal actions for their implementation or requiring sophisticated and costly monitoring.

Political acceptability - The acceptability of a policy instrument for environmental protection can vary significantly between interest groups. A policy that is strongly unacceptable for one or more of the interests or provokes polarization will be rated lower. The social and political context can also vary between jurisdictions.

Environmental benefits - Significant measurable environmental benefits should accrue from the adoption and implementation of the policy instrument.

Effect on the competitiveness of farms and industry - Policy instruments that have an important negative impact on the competitiveness of enterprises have lower probabilities of being adopted and implemented (taking for granted that no financial assistance is available to cover a significant portion of the costs).

Overall administrative, environmental and economic efficiency - This criterion integrates the overall efficiency and feasibility of the policy instrument.

An important additional criterion that should be included in this list would be appropriate scientific rationale underpinning the choice of policy instrument.

3.5.2.2 Linking technical phosphorus considerations and regulations The link between technical phosphorus risk assessment tools and regulations based on these assessment tools is critical. Choosing an inappropriate risk assessment tool to regulate phosphorus transfer from soil to water will not only add to producers’ production costs, but also may unnecessarily constrain agricultural activities or even be ineffective in ameliorating environmental concerns.

Section 3.6 of this report provides a detailed review of methods of assessing risk of phosphorus transfer from soil to water. These methods can be grouped into three categories: agronomic phosphorus tests, environmental phosphorus tests, and total phosphorus tests. Looking at regional agronomic balances of phosphorus can help identify regions with large phosphorus build-ups. Such regions can then be targeted with water protection programs and more field specific methods of risk assessment (see section 3.6.2). The Dutch MINAS program is based on nutrient balance for both N and P.

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A number of jurisdictions base their regulations on the P index. These include Ontario, Quebec, Iowa, Maryland, Minnesota, and Ireland. See section 3.6.4 for a description of different modifications of the P index. As was mentioned earlier, researchers in Iowa, Maryland and Minnesota have expressed reservations about the applicability of such an index for developing individual farm nutrient management plans or as an evaluation scale for determining whether land users are complying with water quality or nutrient management standards established by local, state or federal agencies. As indicated in section 3.6.6, the development of an effective P Index for Manitoba will require further research in some key areas before one could be comfortable that it would be an effective regulatory tool.

3.5.2.3 Concluding observations Before reaching a final conclusion on the applicability of regulating manure management based on phosphorus loadings in Manitoba, it is necessary to investigate the feasibility and effectiveness of several science based options for regulating phosphorus management in Manitoba. At that juncture, it would be useful to apply the criteria mentioned above to a number of case studies to evaluate the impact of various regulatory options as to their suitability to the Manitoba situation. It would also be useful to explore in more depth and monitor the developments in Quebec, Ontario, Minnesota, Maryland, and Iowa, as well as the Netherlands and Ireland, to assess the ongoing effectiveness of their regulations.

Some regulatory recommendations and regulatory cautions follow.

Regulatory recommendations include: • Given that about 60% of phosphorus loadings in the Red River originate from US

sources, efforts need to be expended by the Government of Manitoba to work with US jurisdictions to reduce phosphorus loadings before they enter Manitoba.

• There is a need for a more collaborative approach among government departments and agencies in the development of a phosphorus management strategy for Manitoba, especially among Manitoba Agriculture and Food, Manitoba Conservation, Environment Canada’s National Water Research Centre, and Fisheries and Oceans Canada’s Freshwater Institute.


• Invest in research to reduce P in manure (e.g. phytase management and other feed additives), before regulating P.






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Regulatory cautions include: • Tighter environmental regulations will impact small-scale farm operators more

negatively than large-scale farm operators. Larger farm operations are in a better position to have the financial resources, technical knowledge, and human resources to know and follow increasingly complex regulations.


3.6 Abbreviations AAFRD Alberta Agriculture Food and Rural Development AU Animal unit CAFO Confined animal feeding operation CAP Common agricultural policy (Europe) CFO Confined feeding operations CNMP Comprehensive nutrient management planning DEFRA Department for Environment, Food, and Rural Affairs (England) EC European Commission ECSWCC Eastern Canada Soil and Water Conservation Centre EU European Union IAE Intensive agricultural enterprises IDNR Iowa Department of Natural Resources IMMAG Iowa Manure Management Action Group ILO Intensive livestock operation LU Livestock unit (Europe) MPCA Minnesota Pollution Control Agency MDS Minimum distance separation MINAS Minerals accounting system (The Netherlands) MMP Manure management plans N Nitrogen NPDES National Pollution Discharge Elimination System (US) NRCB Natural Resources Conservation Board (Alberta) NMP Nutrient management plan NMS Nutrient Management Strategy NSA Nitrate Sensitive Areas (England) OMAF Ontario Ministry of Agriculture and Food P Phosphorus REPS Rural Environment Protection Scheme (Ireland) SAFRR Saskatchewan Agriculture, Food and Rural Revitalization STP Soil test phosphorus USDA United States Department of Agriculture USEPA United States Environmental Protection Agency

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3.7 References Alberta Agriculture, Food, and Rural Development. Confined Feeding Operations Website. Available online at: http://www.agric.gov.ab.ca/navigation/livestock/cfo/index.html

Aldinger, Helmuth. 2001. Agricultural and Environmental Policies in the EU and their Impact on Fertilizer Consumption. 69th IFA Annual Conference. Sydney, Australia

Baumol, W.J., and Oates, W. 1988. The Theory of Environmental Policy (2nd edition). Cambridge University Press

Bertrand, R. and R. Beaulieu. 2003. Evolution of the Quebec Environmental Regulations for Agricultural Operations. Presentation to Agricultural Phosphorus Update 2003. April 16 2003, Winnipeg.

Brinkman, George. 1998. Canadian Agrifood Policy Handbook. University of Guelph

Brogan, Jane, Crowe, Matt and Carty, Gerry. 2001. Developing a National Phosphorus Balance for Agriculture in Ireland: A Discussion Document. Environmental Protection Agency of Ireland. Available online at http://www.epa.ie/pubs/docs/Phosphorus%20Balance.pdf

Carton, Owen T. and Magette, William L. May 1999. Land Spreading of Animal Manures, Farm Wastes & Non-Agricultural Organic Wastes Part 1 - Manure (and Other Organic Wastes) Management Guidelines for Intensive Agricultural Enterprises. TEAGASC (Ireland). Available online at http://www.teagasc.ie/research/reports/environment/4026/eopr-4026.htm

Coale, Frank. 2000. The Maryland Phosphorus Site Index Technical Users Guide. Soil Fertility Management Information Series, SFM – 7. Maryland Extension Service. Available online at: http://www.agnr.umd.edu/MCE/Publications/Publication.cfm?ID=538&cat=13

Coale, Frank J., Sims, J. Thomas, and Leytem, April B. 2002. Accelerated Deployment of an Agricultural Nutrient Management Tool: The Maryland Phosphorus Site Index. J. Environ. Qual. 31:1471-1476

Coombe, Richard. 2003. Working Together on Water Standards. Presentation to the Freshwater Forum, February 18, 2003, Winnipeg. Available online at http://www.cecmanitoba.ca/files/Coombe.ppt

Eastern Canada Soil and Water Conservation Centre. 1997. Policy Instruments for Environmental Protection in Agriculture: Analytical Review of the Literature. Available online at http://www.cuslm.ca/ccse-swcc/publications/english/policy_iep.pdf

Hesketh, N. and P.C. Brookes. 2000. Development of an Indicator for Risk of Phosphorus Leaching. Journal of Environmental Quality 29:105-110

Hilborn, D. and R.P. Stone. 2000. Determining the Phosphorus Index for a Field. Ontario Ministry of Agriculture and Food. Available online at http://www.gov.on.ca/OMAFRA/english/engineer/facts/98-079.htm

Iowa Manure Management Action Group. Website on Manure Plans and Permits - State Regulations. Available online at http://extension.agron.iastate.edu/immag/ppstothreg.html.

Jacobsen, Brian H. 2002. Reducing Nitrogen Leaching in Denmark and the Netherlands: Administrative Regulation and Costs. Poster Paper for the Xth EAAE Conference. Zaragoza

http://www.agric.gov.ab.ca/navigation/livestock/cfo/index.html

http://www.epa.ie/pubs/docs/Phosphorus Balance.pdf

http://www.teagasc.ie/research/reports/environment/4026/eopr-4026.htm

http://www.agnr.umd.edu/MCE/Publications/Publication.cfm?ID=538&cat=13

http://www.cecmanitoba.ca/files/Coombe.ppt

http://www.cuslm.ca/ccse-swcc/publications/english/policy_iep.pdf

http://www.gov.on.ca/OMAFRA/english/engineer/facts/98-079.htm

http://extension.agron.iastate.edu/immag/ppstothreg.html

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Landwise Inc. 2001. Phosphorus Standards in Alberta: Potential Impacts on the Agricultural Industry. Prepared for the soil Phosphorus Limits Committee of Alberta Agriculture, Food, and Rural Development

Minnesota Pollution Control Agency. Website. Available online at http://www.pca.state.mn.us/hot/feedlots.html

Mulla, David. 2003. Adaptation of the Phosphorus Index in Minnesota, Presentation to Agricultural Phosphorus Update 2003, April 16 2003, Winnipeg

Natural Resources Conservation Service. 2001. Iowa Phosphorus Index. Available online at http://www.ia.nrcs.usda.gov/Technical/Phosphorus/phosphorusstandard.htm

Natural Resources Conservation Service. National Conservation Practice Standard - Nutrient Management. Nutrient Management Code 590. Available online at http://www.nrcs.usda.gov/technical/ECS/nutrient/590.html

Netherlands Ministry of Agriculture, Nature Management, and Fisheries. 2002. Manure and the Environment: The Dutch Approach to Reduce the Mineral Surplus and Ammonia Volatilization. 2nd edition. Available online at http://www.minlnv.nl/international/policy/environ/

North Carolina Soil and Water Commission. 2001. State Plan for Managing Animal Manures and Animal Derived Nutrients in North Carolina. Available online at http://www.enr.state.nc.us/DSWC/pages/state%20nutrient%20plan.pdf

Olson, Barry, Brent A. Paterson, Joanne Little, and Sheilah Nolan. 2003. Alberta Soil Phosphorus Limits Project. Presentation to Agricultural Phosphorus Update 2003. April 16 2003, Winnipeg.

Ontario Ministry of Agriculture and Food. 2000. The Poop on Phosphorus. Available online at http://www.gov.on.ca/OMAFRA/english/livestock/dairy/facts/info_poop.htm.

Ontario Ministry of Agriculture and Food. 2002. Ministry Protocols for Ontario Regulation Made under the Nutrient Management Act, 2002. Available online at http://www.gov.on.ca/OMAFRA/english/agops/nut_units.htm.

Ontario Ministry of Agriculture and Food. 2003. Ewes Government Responds to Public Consultations with New Direction on Nutrient Management. News Release March 21, 2003. Available online at http://www.gov.on.ca/OMAFRA/english/infores/releases/2003/032103.html

Ontario Ministry of Agriculture and Food. Nutrient Management in Other Jurisdictions. Website. Available online at http://www.gov.on.ca/OMAFRA/english/agops/nutmgtlinks.htm

Quebec Ministère d'État aux Affaires municipales et à la Métropole, à l'Environnement et à l'Eau. 2002. Quebec Regulation Respecting Agricultural Operations – Highlights. Available online at http://www.menv.gouv.qc.ca/sol/agricole-en/explagriANG6.pdf

Saskatchewan Agriculture, Food and Rural Revitalization and Saskatchewan Environment. 2002. Environmental Review Guidelines for Intensive Livestock Operations. Available online at http://www.agr.gov.sk.ca/docs/livestock/pork/intensive_hog_operations/ILOreview2002.pdf

http://www.pca.state.mn.us/hot/feedlots.html

http://www.ia.nrcs.usda.gov/Technical/Phosphorus/phosphorusstandard.htm

http://www.nrcs.usda.gov/technical/ECS/nutrient/590.html

http://www.minlnv.nl/international/policy/environ/

http://www.enr.state.nc.us/DSWC/pages/state nutrient plan.pdf

http://www.gov.on.ca/OMAFRA/english/livestock/dairy/facts/info_poop.htm

http://www.gov.on.ca/OMAFRA/english/agops/nut_units.htm

http://www.gov.on.ca/OMAFRA/english/infores/releases/2003/032103.html


http://www.menv.gouv.qc.ca/sol/agricole-en/explagriANG6.pdf

http://www.agr.gov.sk.ca/docs/livestock/pork/intensive_hog_operations/ILOreview2002.pdf

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Sibbesen, Erik and Runge-Metzger, Artur. 1995. Phosphorus Balance In European Agriculture - Status And Policy Options. In H. Tiessen (ed). SCOPE 54 -Phosphorus in the Global Environment - Transfers, Cycles and Management. John Wiley & Sons Ltd

UK Department for Environment, Food and Rural Affairs. 2002. Guidelines for Farmers: Nitrate Vulnerable Zones – England. Available online at http://www.defra.gov.uk/corporate/regulat/forms/agri_env/nvz/nvz4.pdf

USDA and USEPA. 1999. Unified National Strategy for Animal Feeding Operations. Available online at http://cfpub.epa.gov/npdes/afo/ustrategy.cfm

http://www.defra.gov.uk/corporate/regulat/forms/agri_env/nvz/nvz4.pdf

http://cfpub.epa.gov/npdes/afo/ustrategy.cfm

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