BENEFICIAL MANAGEMENT PRACTICES FOR THE APPLICATION

OF MANURE ON TILE-DRAINED LANDS:

A REVIEW OF LITERATURE

Marcos R. C. Cordeiro

University of Manitoba, Postdoctoral Fellow

David A. Lobb

University of Manitoba, Professor

Don N. Flaten

University of Manitoba, Professor

Henry Wilson

Agriculture and Agri-Food Canada, Research Scientist

- Final Report, November 2016 -

ii

EXECUTIVE SUMMARY

This literature review provides an overview of the benefits and risks of applying livestock manure

to farmland with subsurface (tile) drainage for the climatic, physiographic, and agronomic conditions

prevailing in Manitoba. Specific aspects addressed in this review include:

o Description of factors that need to be considered when determining whether or not subsurface

drainage will result in agronomic and environmental benefits;

o Summary of research on nutrient and pathogen transport from tile-drained lands to surface

waters;

o Description of the materials, methods (installation, operation and maintenance) and approaches

used in drainage systems, including controlled drainage systems;

o Summary of beneficial management practices (BMPs) for soil, water, manure and crops that

decrease risk of nutrient and pathogen transport to water from subsurface drainage systems; and

o Suggestions for future research and development (R&D) needs/priorities in the use of tile

drainage in Manitoba through identification of gaps in the research that has been conducted to

date.

This report is not intended to provide guidelines, let alone regulations, for design or operation of

drainage systems, nutrient management, or BMPs. Rather, it provides technical background for the

discussion of tile drainage and BMPs, and their implications to hydrology and water quality. Thus,

key concepts of soil physics, standard design methods of tile drainage systems, and the major

hydrologic process impacted by tile drainage at field scale are discussed in the first sections of this

review. The following sections describe the benefits and risks of tile drainage from an agronomic and

environmental standpoint. The remainder of the document is dedicated to presenting selected BMPs

for maximizing the agronomic benefits and minimizing the environmental risks of manure application

on tile drained land, with considerations to application of these BMPs to Manitoba conditions.

While this report refers to research results and BMPs that have been adopted in other jurisdictions

and could potentially be relevant for adoption in Manitoba, recommendation of BMP regulations and

policies for this Province is beyond the scope of this work. Similarly, the discussion on BMP

effectiveness does not attempt to rank potential BMPs for adoption since such ranking would require

site-specific consideration of other aspects such as economical/financial implications to farmers,

public incentives through provincial or federal programs, as well as synergism between agronomic,

physiographic (e.g., soils), and climatic variables. Rather, the aim of this work is to provide a source

of science-based information for several stakeholders in Manitoba’s agricultural sector, including

RMs, government agencies, livestock and crop producers, water management contractors, and

researchers.

The key elements of this report are summarized below and are organized around the directives

listed in the original request for proposals:

Detail the factors that need to be considered when determining whether or not subsurface

drainage is beneficial.

The purpose of tile drainage is to remove excess soil water for improved crop growth and improved

conditions for field operations, as well as to remove salts within the root zone. For any specific field

situation, there are several factors that can determine the need for and the benefits from subsurface

drainage: (i) agronomic (crop yields and their uniformity and stability); (ii) economic (yield losses,

price of crops, cost and return on investment); (iii) technical: (availability of equipment and expertise

to install tile drainage, access to regional drainage to provide an outlet for tile outflow); and (iv)

environmental (potential increase in runoff from the farm and contribution to downstream flooding,

iii

potential increase export of nutrients and pathogens and contribution to downstream water

contamination).

There is a strong argument for the improvement of on-farm drainage in Manitoba based on crop

insurance claims for crop loss due to excess moisture, which is by far the largest reason for insurance

payouts. What often limits the efficacy of drainage, whether it is surface or subsurface drainage, is the

ability to remove the water from the drainage system – often, there is nowhere for the water to go,

after it leaves the field.

Review and provide a summary of research on nutrient and pathogen transport from tile-drained lands to waters. Comparison of nutrient and pathogen transport from non-tiled land and tiled land amended with manures would be particularly valued.

In removing excess water from farmland and delivering it to off-farm drainage networks, both

surface and subsurface drainage also deliver nutrients and pathogens to surface waters. Furthermore,

application of manure or synthetic fertilizers is often associated with increased losses of nutrients to

surface and subsurface drainage systems. With respect to the comparison between nutrient and

pathogen losses from tiled and non-tiled land, there is evidence of: increased delivery of water from

tile-drained lands; higher N and lower P concentration in tile drainage outflow; and pathogen

transport from the soil surface to tile drains. However, the delivery of nutrients and pathogens from

non-tiled lands will vary greatly depending on the presence, intensity and efficacy of surface drainage,

making the comparison of tiled to non-tiled situations extremely difficult for Manitoba. Simply put,

the appropriate comparison is a combination of surface and subsurface drainage versus surface

drainage alone.

Describe the structures, methods and approaches that are currently being used in controlled drainage systems (those that go beyond conventional tile drainage) and their suitability for use under Manitoba conditions.

Controlled drainage systems, broadly, consist of systems that employ technology at the tile outlet to

manage the outflowing drain water. What is normally referred to as controlled drainage consists of a

control structure that can be operated to regulate the outflow. By closing the control structure, the tile

drain outflow is stopped, and through regulation of the control structure, the downstream flow of

water, nutrients and pathogens from the tile drains can be manipulated and the negative impacts

lessened. Other structural technologies exist to affect the outflow from tile drains. The most promising

is likely the collection of the water in retention ponds for reuse on farm.

Summarize operating BMPs that decrease risk of nutrient and pathogen transport to water that can occur with subsurface drainage systems.

Summarize system maintenance BMPs that decrease risk of nutrient and pathogen transport to water that can occur with subsurface drainage systems.

Controlled drainage consistently reduces the amount of water and N in tile drainage outflow.

However, the effect of controlled drainage on P loss is not consistent.

The “4R” nutrient stewardship framework (the “right” source, rate, timing and placement) of

applied nutrients, offers useful recommendations for optimizing crop uptake and minimizing losses of

synthetic fertilizers and livestock manures. Among sources of nutrients, water soluble synthetic

fertilizers are at least as susceptible to losses as livestock manures; however, losses of manure

nutrients are highly variable, depending on manure composition, handling, and storage. Of special

note, however, is the risk of liquid manure flowing into tile drains if application rates exceed the

water holding capacity of the soil. From a timing perspective, it’s very important to avoid application

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of manure or fertilizer when tile flow is occurring or else is likely to occur shortly after application.

For placement, tillage to incorporate surface-applied manure and some types of liquid manure

injection systems help to reduce preferential flow of manure to tile drains by disrupting preferential

flow pathways.

In general, tillage that disturbs the soil prior to, during or immediately following the application of

nutrients and pathogens can disrupt the continuity of macropores from the soil surface to the drain

tiles. The need for and effectiveness of this practice is specific to soils and crops that develop

extensive macroporosity, e.g. soils with a deep and extensive network of earthworm burrows and/or

heavy clay soils that form cracks during dry periods. However, the degree to which these risks and

BMPs apply to tile drainage situations in Manitoba soils has not yet been determined.

Crop rotations, including cover crops and perennial forages, have inconsistent effects on total

losses of nutrients to water from surface and subsurface pathways.

End-of-pipe technologies have been proposed for treating tile drainage outflow, such as water

retention (e.g., wetlands), biofilters, and vegetative filter strips or buffers. Wetlands are generally

more efficient at removing N than P, but their lifetime may be limited if the trapped nutrients are not

periodically removed from the wetland. True biofilters are also more effective at removing N than P

and require chemical additives, such as iron oxide, to enhance their capacity for removing P; even so,

the capacity for biofilters to remove N is very limited when temperatures are cold, such as during

snowmelt, when are large portion of tile drainage flow occurs. Vegetative buffers are vulnerable to

the same problems as biofilters and are generally ineffective for intercepting nutrients during

snowmelt.

Outline any gaps in the research that has been conducted to date and make recommendations as to future R&D needs/priorities in the use of tile drainage in Manitoba.

Tile drainage has been the subject of study for many decades in some regions of North America

such as the Great Lakes Lowlands and the St. Lawrence Lowlands, and research in these areas

remains active in efforts to improve the efficacy of drainage systems for agricultural production and

to better understand the impacts of drainage on the environment. The focus of this research has shifted

over that time from engineering criteria and crop productivity to the delivery of nutrients, pesticides

and pathogens to surface waters and the emission of greenhouse gases. The continued intensity of this

research reflects the sensitivity of drainage system performance to specific local environmental

conditions. Given the lack of research on drainage systems in Manitoba, it could be argued that all

aspects of tile drainage require substantial research.

Certainly, there is a need to design tile drainage systems for the unique combination of climate,

topographic, soil and cropping conditions we have in Manitoba to ensure the agronomic benefits and

economic costs can be rationalized. To minimize off-farm impacts, there is a need to better

understand the transport of water, nutrients and pathogens into and through the soil and into and

through tile drains, and to understand the effects management can have on this transport. Equally, if

not more important is the need to better understand the transport of water, nutrients and pathogens

into and through in-field surface drains and the effects management. With respect to nutrients, the

timing and placement of livestock manure and synthetic fertilizers are critical factors determining

their effective use and potential loss, but the greatest research needs in nutrient management related to

drainage systems are in the use of cropping and tillage practices.

Although beyond the scope of this contract, it should be noted that there is a lack of attention given

to the management of in-field drainage systems, surface or subsurface, in the context of water

management at the farm scale and at the municipal scale. The connections between these levels of

water management are not well understood, particularly when it comes to the transport of nutrients

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from fields to surface waters. To illustrate, during major snowmelt and rainfall events, the excess

water that inundates farm fields and ultimately drains from these fields is often the result of the

controlled and regulated flow of water through the municipal and provincial drainage systems. The

nature of the connections and the integration of the management of drainage systems may be very

specific to Manitoba. For this reason, it is recommended that considerable attention be given to: (i)

integrated management of in-field surface and subsurface drainage; (ii) integrated management of in-

field drainage with on-farm water management; and (iii) integrated management of on-farm water

management with municipal and provincial water management.

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1. INTRODUCTION .................................................................................................... 1

1.1. OBJECTIVES ......................................................................................................................3

1.2. SCOPE ..............................................................................................................................3

1.3. TILE DRAINAGE OVERVIEW ............................................................................................... 4

2. KEY CONCEPTS IN SOIL PHYSICS AND HYDROLOGY ...................................... 5

3. GENERAL DESIGN OF TILE DRAINAGE SYSTEMS ............................................. 9

4. HYDROLOGICAL PROCESSES FOR TILE DRAINAGE ....................................... 15

4.1. SOIL STORAGE ................................................................................................................ 15

4.2. INFILTRATION ................................................................................................................. 17

4.3. PERCOLATION ................................................................................................................. 18

4.4. SURFACE RUNOFF ........................................................................................................... 19

5. BENEFITS OF TILE DRAINAGE SYSTEMS .......................................................... 21

5.1. AGRONOMIC BENEFITS .................................................................................................... 21

5.2. ENVIRONMENTAL BENEFITS ............................................................................................. 23

6. RISKS ASSOCIATED WITH TILE DRAINAGE SYSTEMS ....................................24

6.1. AGRONOMIC RISKS ........................................................................................................ 24

6.2. ENVIRONMENTAL RISKS ................................................................................................. 26

6.2.1 Nutrient movement (matrix and preferential flow) .......................................................... 27

6.2.1.1 Manure ................................................................................................................................ 27

6.2.1.2 Synthetic fertilizer ............................................................................................................... 32

6.2.1.3 Manure versus synthetic fertilizer ........................................................................................ 38

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6.2.2 Pathogen transport (matrix and preferential flow) ......................................................... 39

7. BENEFICIAL MANAGEMENT PRACTICES (BMPS) FOR MAXIMIZING THE

AGRONOMIC BENEFITS AND MINIMIZING THE ENVIRONMENTAL RISKS OF

MANURE APPLICATION ON TILE DRAINED LAND.......................................................42

7.1. CONTROLLED TILE DRAINAGE .........................................................................................43

7.1.1 Design .......................................................................................................................... 46

7.1.2 Operation .......................................................................................................................47

7.1.3 Maintenance ................................................................................................................. 48

7.2. NUTRIENT MANAGEMENT ............................................................................................... 48

7.2.1 Sources (manure types and manures versus synthetic fertilizers) ................................... 49

7.2.2 Rates of nutrients (relative to crop requirements and removals) ...................................... 53

7.2.3 Matching volumes of liquid manure applied to available water holding capacity in soil 56

7.2.4 Placement (including practices to minimize risks of preferential flow of manure, such as

injection) .................................................................................................................................... 57

7.2.5 Timing relative to rainfall and snowmelt events ............................................................. 59

7.3. AGRONOMIC PRACTICES................................................................................................. 60

7.3.1 Tillage ............................................................................................................................ 61

7.3.2 Crop rotation ................................................................................................................ 63

7.3.3 Cover crops ................................................................................................................... 65

7.4. END-OF-PIPE MANAGEMENT ........................................................................................... 67

7.4.1 Natural (wetlands) and constructed reservoirs .............................................................. 67

7.4.2 Biofilters ....................................................................................................................... 69

7.4.3 Vegetative filter strips/buffers ......................................................................................... 72

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8. CONSIDERATIONS FOR ADAPTING TILE DRAINAGE BMPS FOR MANITOBA

CONDITIONS ................................................................................................................... 74

8.1. PHYSIOGRAPHIC CONDITIONS (CLIMATE, RELIEF, SURFICIAL GEOLOGY, SOILS, ETC.) AND THE

INFLUENCE OF SNOWMELT, FROZEN SOILS, CRACKING SOILS, SHALLOW WATER TABLES, ETC. .............. 74

8.2. AGRONOMIC PRACTICES................................................................................................. 76

9. KNOWLEDGE GAPS AND RESEARCH NEEDS/PRIORITIES ............................... 78

10. DISCLAIMER ....................................................................................................... 81

11. ACKNOWLEDGEMENTS ..................................................................................... 81

12. REFERENCES ....................................................................................................... 81

1

1. INTRODUCTION Agricultural drainage has been an important aspect for the socio-economic development of

southern Manitoba. It has shaped the landscape of this part of the Province since the early stages of

settlement (Bower, 2006). In the late 1800’s, drainage projects were perceived as a strategy to achieve

the agricultural potential of the province, which led to a drastic change of the landscape in the

following decades by means of surface drainage (Bower, 2007). However, the development promoted

by surface drainage also gave rise to challenges. Increased downstream flooding was observed since

the implementation of the major surface drainage projects in the first decades of the 20th

century

(Bower, 2006), which continues to be a major issue today (Brunet and Westbrook, 2012). Moreover,

surface drainage can contribute to water quality problems in downstream water bodies as a result of

increased phosphorus export in years with higher volumes of stream discharge in the Red River Basin

(McCullough et al., 2012).

More recently, a new form of drainage has been increasingly adopted by farmers, namely,

subsurface (tile) drainage. Tile drainage installations have steadily grown since the late 1990’s with

reported increases in agricultural production in marginal land (Lyseng, 2004). However, little research

is available about tile drainage agronomic and environmental performance in Manitoba (Farming for

Tomorrow, 2011). Negative environmental impacts due to nutrient export through tile drainage have

been reported in other agricultural regions of the North America, including several US states

(Schilling and Libra, 2000; Jaynes et al., 2001; Randall and Mulla, 2001; Strock et al., 2004b;

Petrolia and Gowda, 2006; Royer et al., 2006; David et al., 2010), and eastern provinces of Canada

[i.e., Quebec, Ontario, New Brunswick, and Prince Edward Island; Milburn et al. (1990); Valero et al.

(2007); Jiang and Somers (2008); Drury et al. (2009)]. However, findings from other jurisdictions

with contrasting characteristics may not apply directly to Manitoba due to differences in climate,

physiography, and agronomic practices. Thus, contextualization of research results to Manitoba’s

condition is crucial.

2

Besides agronomic and environmental performance, technical information about important aspects

of tile drainage such as design criteria, operation and maintenance, agronomic impacts, and

hydrological alterations is also scarce in the Canadian Prairies (Cordeiro and Sri Ranjan, 2012).

General design criteria proposed for Manitoba are usually adopted in tile drainage installations

without further consideration of the particularities of individual fields (Cordeiro and Sri Ranjan,

2015), which could result in inferior agronomic and environmental performance of tile drainage

systems.

This lack information about tile drainage in Manitoba has led to recent controversies about negative

impacts to surface and subsurface water resources and a moratorium on tile drainage installations in

lands receiving livestock manure in some Rural Municipalities [RMs; Harris (2015b)]. While the

efforts to protect water resources are acknowledged and encouraged (Harris, 2015a), such efforts

could lead to environmental and economic loss if they are not based on sound technical information.

For example, manure application at agronomic rates (Moore et al., 1995) and tile drainage (Scott et

al., 1998) have both been proposed as beneficial management practices (BMPs) to reduce export of

contaminants to receiving waters. However, this evidence seems to have been overlooked in the

recent discussions about the risk of manure application on tiled land.

Prohibition of tile drainage installations should not be discussed before giving proper consideration

to agricultural water management in Manitoba, as well, with special emphasis on water retention at

field and watershed scale. While this topic is relatively new in the Province and not yet well

documented in the scientific literature, it has been advocated by institutions such as the Manitoba

Agricultural Water Management Association (MAWMA, 2015). Researchers of the University of

Manitoba have also promoted water management and water retention as a means to reduce nutrient

loss from farmland and mitigate flooding. Among the strategies discussed, maximization of soil

infiltration, moisture storage in the soil profile, and drainage of excess water have been proposed

3

(Lobb, 2015), which can be achieved through tile drainage.

1.1. OBJECTIVES

The general objective of this literature review is to provide an overview of the benefits and risks of

organic fertilization of farmland featuring tile drainage systems for the climatic, physiographic, and

agronomic conditions prevailing in Manitoba. Specific aspects addressed in this review include:

Description of factors that need to be considered when determining whether or not

subsurface drainage will result in agronomic and environmental benefits;

Summary of research on nutrient and pathogen transport from tile-drained lands to waters;

Description of the materials, methods (installation, operation and maintenance) and

approaches used in drainage systems, including controlled drainage systems;

Summary of beneficial management practices (BMPs) for soil, water, manure and crops that

decrease risk of nutrient and pathogen transport to water from subsurface drainage systems;

and

Suggestions for future research and development (R&D) needs/priorities in the use of tile

drainage in Manitoba through identification of gaps in the research that has been conducted

to date.

1.2. SCOPE

This work is not intended to provide guidelines, let alone regulations, for design or operation of

drainage systems, nutrient management, or BMPs. Rather, this document provides technical

background for the discussion of tile drainage and BMPs, and their implications to hydrology and

water quality. Thus, key concepts of soil physics, standard design methods of tile drainage systems,

and the major hydrologic process impacted by tile drainage at field scale are discussed in the first

sections of this review. The following sections describe the benefits and risks of tile drainage from an

agronomic and environmental standpoint. The remainder of the document are dedicated to presenting

4

selected BMPs for maximizing the agronomic benefits and minimizing the environmental risks of

manure application on tile drained land, with considerations to application of these BMPs to Manitoba

conditions.

While this report does refer to research results and BMPs that have been adopted in other

jurisdictions and could potentially be relevant for adoption in Manitoba, recommendation of BMP

regulations and policies for this Province is beyond the scope of this work. Similarly, the discussion

on BMP effectiveness will not attempt to rank potential BMPs for adoption since such ranking would

require site-specific consideration of other aspects such as economical/financial implications to

farmers, public incentives through provincial or federal programs, as well as synergism between

agronomic, physiographic (e.g., soils), and climatic variables. Rather, the aim of this work is to

provide a source of science-based information for several stakeholders in Manitoba’s agricultural

sector, including RMs, government agencies, livestock and crop producers, water management

contractors, and researchers.

1.3. TILE DRAINAGE OVERVIEW

Although a more detailed treatment of tile drainage is given in section 3, an initial description is in

order to contextualize the following sections and to introduce the general principles behind this

practice. Skaggs et al. (2012) outline the different conditions and stages during the tile drainage

process, which are summarized here. The major condition required for operation is that the soil must

be saturated above the tile drains. However, there are several saturations states characterized by

different water table elevations above the tile drains. When the soil profile is completely saturated and

water is ponded on the soil surface as a continuous water surface, runoff can take place from all areas

towards the tile drains. As a result, most of the flow will take place on top of the tile drain, with a

large percentage of the drainage water entering the profile within a horizontal distance equal to the

drain depth. Subsurface drainage rate is at maximum for this condition. When the continuous water

5

surface is disrupted due to infiltration and evaporation, separated puddles of water occur. At this

point, water can no longer move across the surface to the vicinity of the drains and the water table

near the drain is drawn down. Drainage rates continue to decline as the ponded water is removed until

the water table midway between the drains is just coincident with the surface and attains an

approximately elliptical shape. The drawdown process as the water table falls from the soil surface

and finally to drain depth is characterized by decreasing subsurface drainage rates, until it reaches

zero when the water table is below the drain depth.

Since the ponded conditions do not always occur in the field and are usually short when compared

to the drainage process characterized by the lowering of the water table, only the latter process is

usually described in the literature. The drainage process in the soil profile (i.e., lowering the water

table) will be the focus of this review and is described in more detail in following sections, although

surface movement of water has implications for the transport of nutrients and pathogens through tile

drains (sections 6.2.1 and 6.2.2). As well, the hydrological processes affected by tile drainage will be

discussed with regards to lowering of the water table. Another implication of saturated-conditions

requirement for tile outflow to occur is that water percolating through the soil profile will not enter

the tile drain unless it brings the water table above the drainage depth. This is further discussed in

sections 2 and 4.3.

2. KEY CONCEPTS IN SOIL PHYSICS AND HYDROLOGY The soil profile is comprised of soil particles (soil matrix) and pore space, which can be occupied

by air or water (Hillel, 1998). However, the packing of soil particles of different diameters give rise to

different pore sizes due to interlock of small particles among large ones (Figure 1). Pore size is

associated with particle size, where smaller soil particles create smaller the pore sizes; however,

structured soils may have larger pore sizes associated with secondary pore space, even though the

primary pore space is characterized by a small pore size (Corey, 1994).

6

Figure 1. Hypothetical packing of soil particles of different diameters.

Water is held in the pore space between soil particles by capillary forces and adsorption forces

(Smedema et al., 2004). Capillary forces increase as the pore size decreases; as a result, water is held

more strongly in smaller pores and less strongly in larger pores (Corey, 1994). This difference in

capillarity has a direct impact on soil internal drainage. When soils are saturated (i.e., pore space

completely occupied by water; Figure 2a) after a heavy rainfall or snowmelt, water will drain from

large pores by action of gravity. As the large pores empty, smaller pores will start to drain. When the

pore size is reached where capillary forces counterbalance the gravitational force, drainage will stop.

This state characterizes the soil moisture status called field capacity (FC; Figure 2b). The water held

by capillary forces can still be utilized by plants since the root systems can exert suctions much higher

than those typical of field capacity [i.e., –33 kPa; Givi et al. (2004)]. In fact, it is assumed that plants

can extract water held by capillarity up to –1500 kPa, which is called permanent wilting point [PWP;

Figure 2c; Givi et al. (2004)].

7

Figure 2. Soil moisture status representing saturation (A), field capacity (B), and permanent wilting

point (C).

The water drained by gravity between soil saturation and field capacity is called gravitational water,

and is the only portion of the soil moisture that could be intercepted by tile drains. The negative

pressure typical of field capacity water content or lower will prevent water from seeping into the

drains. Thus, the volume of the pore space between saturation and field capacity is the one of interest

with regards to tile drainage. This volume is called drainable pore space [µ; Smedema et al. (2004)],

but it is also referred to as drainable porosity or specific yield (Johnson, 1967; Hillel, 1998).

The variation in pore sizes arising from the arrangement of soil particles of different diameters

affects not only the ability of the soil to hold water, but also water movement through the soil profile.

Hydraulic conductivity (K) pertains to the ability of the soil to convey water or, alternatively, it is the

ease with which the water moves through the soil matrix. Since this matrix is porous, the hydraulic

conductivity is subject to a number of factors that govern the movement of water at pore scale. The K

is related to both liquid and medium properties, as defined by Marshall et al. (1996):

/gkK , (1) (1)

where ρ is the density of water, g is the acceleration of gravity, k is the intrinsic permeability, and η is

the dynamic viscosity of the water. With regards to the liquid, both density and viscosity change with

temperature. Hydraulic conductivity thus increases with increasing temperature as both density and

viscosity decreases. However, these parameters are usually assumed to be uniform over the scales of

interest for tile drainage. As a result, the intrinsic permeability, which encompasses factors depending

8

on the soil matrix, is the factor that most determines K in practice. The permeability of a medium is

related to its porosity (Marshall et al., 1996):

k = r2/8, (2) (2)

where is the porosity of the medium, defined as the fraction of the bulk volume of the soil occupied

by voids [ = VV/VT; Collins (1961)]; r is the pore radius; and k is defined as above. Porosity is

dependent on the size of the particles and the grain size distribution. For particles of uniform size, the

porosity increases as the particle size decreases. That is why clay soils have larger porosity than sandy

soils. For mixed unconsolidated materials such as soils, a large variety of grain sizes will decrease

porosity as the smaller particles sit in between the large ones, thus reducing the pore size through

which the soil water can move [Figure 1; Collins (1961)]. As a result, soils with poorly sorted grains

will exhibit lower porosity than soils with well sorted grains, even if these grains are fine. Of more

importance for hydraulic conductivity, however, is the concept of effective porosity, which is the

fraction of the total volume constituted by interconnected pores. The second variable affecting k in

Eq. (2), namely, pore radius, also depends on soil texture because soils with coarse texture (e.g.,

sands) have larger pores than do soil with fine texture [e.g., heavy clays; Hillel (1998)]. Thus,

hydraulic conductivity is usually larger for sandy soils and decreased through the textural spectrum

towards clays [Chow (1964); Table 1] .

Table 1. Hydraulic conductivity of unconsolidated porous media [Adapted from Chow (1964)]. Soil texture Hydraulic conductivity (cm s

-1)

Sand 10-5

– 1

Silt 10-7

– 10-3

Clay 10-9

– 10-5

Hydraulic conductivity can also vary for a single soil at different water contents (Fetter, 1999). In

practice, K will decrease with decreasing water content from saturation because the pores will become

emptier and the water has to travel in the thin film around the soil particles. As the pores get empty,

the route around the soil particle also become longer due to an increase in tortuosity, which is defined

9

as the ratio between the length of the soil domain and the capillary or water path length. Another

feature of unsaturated hydraulic conductivity is that the matric suction increases as the water content

decreases due to water being held at smaller pores.

While hydraulic conductivity is a key parameter used in soil physics and hydrology, it considers

only water movement through the soil matrix. However, other forms of flow through bio-pores,

cracks, and other major conduits of water may create distinct flow paths in the soil (Smedema et al.,

2004). These types of flow, referred to as preferential flow, promote a faster movement of water to

greater depths via distinct pathways while bypassing a large portion of the soil matrix pore-space

(Hillel, 1998; Šimůnek et al., 2003). In an agricultural context, the hydraulic conductivity of

preferential flow paths has been reported to be one (Skaggs et al., 1980) to three orders of magnitude

(Øygarden et al., 1997) larger than that of the soil matrix.

3. GENERAL DESIGN OF TILE DRAINAGE SYSTEMS Design of tile drainage systems involves many different aspects, including selection of tile material,

soil conditions (e.g., acidity, freezing/thawing), need of envelop or filters for soil stabilization and

selection of their material, and definition of design criteria (ASABE, 2015). Specific design

requirements encompass site topography; direction and amount of surface flow; outlet accessibility;

soil profiles; existing subsurface drain system; utilities; history of cropping patterns; and economics

(ASABE, 2015). Discussion of these characteristics is out of the scope of this review but a number of

sources discuss the design of tile drainage systems in detail (Skaggs and Nassehzadeh-Tabriz, 1986;

USBR, 1993; USDA-NRCS, 2001; Smedema et al., 2004; King and Willardson, 2007; ASABE,

2015). Only one critical aspect pertaining to performance and design of these systems, namely, tile

10

lateral1 spacing, will be discussed since it has direct influence on tile outflow, which affects the water

balance at field scale and, consequently, nutrient dynamics.

The Province of Manitoba suggests a general tile spacing between 12 and 15 meters (MAFRD,

2015), which is usually adopted by contractors (Cordeiro and Sri Ranjan, 2015). This

recommendation can be adjusted based on soil texture and permeability [ASABE (2015); MAFRD

(2015); Table 2]. However, a broad range of lateral spacings may still be suitable for a single soil

textural class, making it difficult to tailor this design criterion for specific field conditions based on

soil texture alone. For example, Cordeiro and Sri Ranjan (2015), using DRAINMOD to simulate corn

yields under different lateral spacings in a sandy loam soil in Manitoba, found that the tile spacing

used in the field (i.e., 15 m) could have been increased to 40 m with minimal effects on yields.

Moreover, the 15-m spacing used could lead to water deficit stress in dry years, which could result in

yield losses. Thus, the general recommendation proposed by the Province should be used only as a

starting point for design and not as a definitive spacing.

Table 2. Pipe drains lateral spacing for various soils Soil Lateral spacing (m)

†

Clay and clay loam soils 9-20

Silt and silty 18-30

Sandy loam soils 30-90

Muck 15-60 † Considers a base depth around 1.2 m (ASABE, 2015).

In order to address the inadequacy of a single spacing for different soil types, the recharge regime

as well as the physical properties of the soil and pipe characteristics should be taken into

consideration. Recharge regime can be defined as steady or non-steady (also referred to as unsteady)

state conditions. Steady state conditions assume constant flow through the soil to the drain, resulting

in tile discharge being equal to recharge and constant water table head (h; Figure 3). In non-steady

1 The present review will focus on materials commonly used in recent installations of tile drainage systems in

Manitoba. Tile drainage networks usually consist of PVC parallel circular pipe drains (i.e., laterals) that

individually outlet to a pipe drain system main or outlet.

11

state conditions, all these parameters vary with time and the water table fluctuates during the drainage

process (Smedema et al., 2004). Soundly based and tested procedures and criteria generally exist only

for steady state design; consequently, design is usually based on steady state formulae and criteria,

although it is recognised that the drainage process is in fact non-steady (Smedema et al., 2004), which

is the case for the Canadian Prairies (Cordeiro and Sri Ranjan, 2012). However, King and Willardson

(2007) argue that water tables fluctuate between drains regardless of whether a steady state or a

transient method is used to determine spacing, and that drainage systems in both humid and arid areas

can be designed using steady state theory when the appropriate design parameters are selected. In fact,

the uncertainty in calculations due to variability in soil parameters can be larger than that arising from

the formula used (USDA-NRCS, 2001). Thus, only steady station conditions will be discussed below.

Figure 3. Variables for determining tile drain spacing [adapted from Smedema et al. (2004)].

A number of steady state formulae are available (Smedema et al., 2004), although the Hooghoudt

equation is often described in the literature (Skaggs and Nassehzadeh-Tabriz, 1986; USDA-NRCS,

2001; Smedema et al., 2004; King and Willardson, 2007; ASABE, 2015). The equation provides

12

acceptable lateral spacing for typical agricultural drainage needs (ASABE, 2015). In this method, a

number of variables pertaining to soil physical properties, environmental conditions, and tile

characteristics have to be defined a priori, including soil saturated hydraulic conductivity (Ksat) above

and below the tile laterals; the depth to the impermeable layer (D), which is used to calculate an

equivalent depth (de) that is always smaller than D; the drainage coefficient (i.e., the amount of water

to be removed by the drainage system); the field drainage base (i.e., depth of the drains; W); and the

maximum allowable water table height above the drains (h). Ksat and depth to impermeable layer can

be defined from field measurements. The drainage coefficient is defined based on the designer’s

experience but typical values range from 1 to 3 mm d-1

in arid regions and from 7 to 10 mm d-1

in

humid areas (King and Willardson, 2007; ASABE, 2015). The field drainage base should be defined

to balance performance, cost, and agronomic practices. The tiles must be deep enough to prevent

damage from tillage and keep costs down, but shallow enough to respond quickly to precipitation

events (MAFRD, 2015). In Manitoba, a general depth of 0.9 m is usually adopted (Cordeiro and Sri

Ranjan, 2015). However, drains should be installed on or above a low-permeability layer and in

highly permeable layers that are at or near the proper depth (USDA-NRCS, 2001; King and

Willardson, 2007). The maximum allowable water table depth above the drains (h) should aim at

keeping good soil aeration for proper root development. It has been proposed that saturation

conditions should be removed from the top 0.3 m of the soil profile (H in Figure 3) within 24-48 h

after a significant rainfall event in order to maintain good aeration for crops such as corn, soybeans,

and wheat (Jacobsen et al., 2010). Thus, the maximum allowable water table depth (h) can be

calculated as the drainage base depth (W) minus the target aeration depth (H).

Once those variables have been defined (Figure 3), Hooghoudt’s equation for layered soils can be

written as (Smedema et al., 2004):

13

2

2

1

2

2 48

L

hK

L

hdKq e , (3)

where q is the recharge or drainage coefficient (mm d-1

); K1 and K2 are the hydraulic conductivities of

first and second soil layers, respectively (m d-1

); de is the equivalent depth (m); h is the maximum

allowable water table depth above the drains (m); and L is the tile spacing (m). Rearranging Eq. (3)

for tile spacing gives:

q

hK

q

hdKL e

2

122 48 . (4)

The equivalent depth (de) can be calculated as:

LDfor

u

D

L

D

Dde 41

1ln8

(5)

or

LDfor

u

L

Lde 41

ln8

, (6)

where u is the wet perimeter of the tile drain (m). Since the tile spacing L depends on the equivalent

depth de and vice-versa, Hooghoudt’s equation has to be solved iteratively by assuming an

approximate value for L to calculate de. The de result is plugged back in Eq. (4) to calculate L. This

process is repeated until the drain spacing converges to a stable result, which is the appropriate

spacing for the conditions assumed. According to Smedema et al. (2004), important features of these

equations are that the equivalent depth becomes independent of the depth to the impermeable layer for

D > ¼ L [Eq. (6)]; also, tile spacing increases with increasing K, decreasing q, and increasing h.

Moreover, shallower drains require narrower spacing to achieve equivalent drainage (King et al.,

2015b) because h decreases with decreasing W for a constant H.

Tile drainage systems with gravity outlets operate continuously and automatically with minimum

14

maintenance. However, if the outlet is a sump with a pump, then continuous maintenance of that part

of the system is necessary to assure adequate drainage (King and Willardson, 2007). In Manitoba,

sump outlets discharging into municipal ditches are used if a gravity outlet is not feasible.

Envelope materials around subsurface drains are used to prevent the movement of soil material into

the drain (which may settle and clog the drain); to increase the permeability around the drain

openings; to provide suitable bedding for the drain; and to stabilize the material on which the drains

are laid (Willardson, 1974). Usually, pipe drains installed in stable soils will give satisfactory

performance without specific requirements for soil stabilization; however, unstable soil situations,

such as fine sand and non-cohesive soil, require stabilizing (ASABE, 2015). Envelope/filter materials

provide structural stability and improved hydraulic performance, while aggregate and geotextile

stabilizing materials provide structural stability and, to a lesser extent, improve the hydraulic

performance of the system (ASABE, 2015). In Manitoba, geotextile envelopes (usually called

“socks”) around corrugated plastic tubing are normally used.

Plant roots can invade and clog tile drains, while deep-rooted plants can clog drains in a relatively

short period during times of water shortage (King and Willardson, 2007). Trees or other permanent

vegetation may interfere with operation and maintenance of the system (ASABE, 2015). In Manitoba,

such conditions may occur close to shelter belts and render drains potentially susceptible to root

clogging. It is recommended that the pipe components of tile drainage systems should be located,

when possible, 30 m away from water-loving trees (e.g., willow, elm, cottonwood, and maple) and 15

m away from other tree species (ASABE, 2015). If trees near pipe drains cannot be eliminated, use of

rigid, non-perforated drain pipes is recommended for that section of the system (King and Willardson,

2007; ASABE, 2015).

The appearance of new wet spots near a drain line in a field is an indication that the drain is

clogged at that point; the drain can then be excavated, examined, and repaired (King and Willardson,

15

2007). Water jet cleaning using high-pressure jetting equipment similar to that used for sewer cleaning

is effective for many clogging and sealing problem (Grass et al., 1975). Sections of drains up to 200

m long can be easily cleared of roots, sediment, and iron ochre deposits by jet cleaning at pumping

pressures between 4 and 5.5 MPa (Grass et al., 1975; King and Willardson, 2007). The drains are

cleaned in sections from excavated access holes; after an appropriate plug is placed into the

downstream section of drain at the access hole, an upstream direction from this plug is cleaned and

the water and debris are pumped out of the access hole (King and Willardson, 2007). When cleaning

of this section is finished, the drain pipe and envelope are repaired and the access hole is filled.

Cleaning costs are usually less than 20 % of the cost of installing a new drain (King and Willardson,

2007).

4. HYDROLOGICAL PROCESSES FOR TILE DRAINAGE

4.1.SOIL STORAGE

The soil’s ability to hold water, also referred to as soil water storage, is an important component of

the water balance, and it is usually calculated as the difference between precipitation minus drainage

and evapotranspiration (Kengni et al., 1994):

ETDRS (7)

where ΔS is change in the soil water storage (mm), R is recharge by precipitation or irrigation (mm),

D is drainage of the soil profile at the depth considered (mm), and ET is evapotranspiration (mm). The

dependence of soil storage to important variables for agricultural production such as precipitation,

irrigation, and drainage make it a central aspect for agricultural water management. In this context, it

has been traditionally represented by the “bucket model” [Hillel (1998); Figure 4].

16

Figure 4. The "bucket model" representing soil moisture availability between field capacity and wilting

point [adapted from Hillel (1998)].

As discussed in section 2, the soil profile is able to hold water at different degrees of suction due to

the presence of pores of different sizes. Any moisture between saturation2 and field capacity is not

used by plants and is subject to drainage; soil moisture status between field capacity and wilting point

are held by negative pressure (i.e., suction) and are available for plants; soil moisture below the

wilting point is not available for plants due to the high negative pressure at which the water is held

(Hillel, 1998). Equation (7) and Figure 4 indicate that drainage is an important variable affecting soil

storage. Modelling and field studies have shown that tile drainage can increased soil moisture storage

(Skaggs et al., 1994) by quickly removing gravitational water that otherwise would take longer to

drain. By increasing moisture storage, drainage also influences other hydrological processes at field

scale, such as infiltration, percolation, and surface runoff.

2 The boundary between the saturated and unsaturated zone in the soil profile is not always defined by the

water table but at some elevation above it corresponding to the upper extent of the capillary fringe, which is

also saturated but under negative pressure equal to the air-entry value for the soil (Hillel, 1998).

17

4.2.INFILTRATION

The movement of water from the soil surface into the soil is defined as infiltration (Dingman, 2008)

and depends on the soil moisture status, which is a directly related to soil storage. Common

physically-based descriptions of infiltration such as the Green-Ampt and Phillip models take water

content or matric suction, which are related, into consideration (Mishra et al., 2003). Empirical

approaches such as Horton’s infiltration model also imply dependence on soil moisture status

(Chahinian et al., 2005). The relationship between infiltration and soil moisture content is also evident

from typical field measurements of saturated hydraulic conductivity, where infiltration rates are large

when the soil is dry but decreases at higher water contents (Beasley et al., 1980; Hillel, 1998).

By lowering the water table, tile drainage increases the pore space (i.e. storage) available for

infiltration (Skaggs et al., 1994). Expressed in terms of soil moisture status, tile drainage can shorten

the time the soil takes to go from saturation to field capacity, thus creating an opportunity for water to

infiltrate. The drainage intensity of the tile network also has an impact on soil moisture and

consequently, on infiltration. For example, Larney et al. (1988), investigating the impact of tile

spacing on soil moisture in a poorly-drained silt loam soil in Indiana, found that tile spaced at 20

meters resulted in drier soil moisture regime than tiles spaced at 40 meters. Their results also showed

that soil moisture was lower close to the drains and higher further away from them, suggesting that

infiltration may not be uniform in tile drained fields.

In cold climates such as that of the Canadian Prairies, infiltration into the soil profile can be

restricted by frozen soils (Shook and Pomeroy, 2012). Infiltration can be restricted in mineral and

other soils frozen in a wet condition (e.g., >70–80 % pore saturation) or if a basal ice lens forms at the

soil surface due to rain that occurs near freeze-up or the freezing of percolating or infiltrating

meltwater (Zhao and Gray, 1999; Gray et al., 2001). Infiltration can also be restricted when soil

infiltrability is governed primarily by the soil moisture content (water + ice) and soil temperature at

the start of snow ablation and the infiltration opportunity time (Gray et al., 2001). That is, infiltration

18

is governed primarily by the snow cover water equivalent (SWE) and the frozen water content of the

soil layer extending from the surface down to 0.3 m (Gray and Prowse, 1993), as well as time window

during which infiltration can occur (i.e., opportunity time). In the Cold Region Hydrological Model

(CRHM) platform (Pomeroy et al., 2007), the opportunity time is calculated as the period beginning

when the snowmelt exceeds 2 mm per day and finishing when the average SWE is less than 2 mm.

Deep frost penetration up to 2.0 m in Manitoba (Selezneva et al., 2008; Satchithanantham and Sri

Ranjan, 2015) goes beyond the drainage depth usually adopted in the Province and certainly affects

infiltration during winter months. Recent research on tile drainage in Manitoba indicates no flow

between December and March (Satchithanantham et al., 2014; Satchithanantham and Sri Ranjan,

2015). However, insulation provided by the snowpack increases the temperature of the soil profile and

enhances infiltration during spring (Kahimba et al., 2008). Thus, although tile drainage outflow has

not been observed in Manitoba before March, it is possible that small infiltration rates could still

occur in the soil profile above the drains due to infiltration promoted by the higher temperatures in the

soil profile arising from snowpack insulation. However, these small amounts of infiltration would not

be large enough to generate tile outflow due to either limited penetration (i.e., not reaching tile drain

depth) or limited volume.

Unsaturated soil conditions can also enhance infiltration into frozen soils in the Canadian Prairies

(Fang and Pomeroy, 2007), which has led to suggestions that tile drainage can be used to lower the

water table in the fall and create air-filled pore volume to accommodate the infiltration during the

spring snowmelt (Cordeiro et al., 2015b). This enhanced infiltration during spring would potentially

reduce surface runoff, suggesting that tile drainage can be used as a potential management tool to

reduce runoff volumes during critical spring snowmelt periods.

4.3.PERCOLATION

The downward flow through the unsaturated zone of the soil profile is referred to as percolation

19

and is essential for groundwater recharge (Dingman, 2008). While groundwater can be recharged by

other mechanisms such as through areas where the water-bearing formation is exposed to the

atmosphere or from surface water bodies (Wolfe, 2000), percolation is of more importance in an

agricultural context since a vast land base (through which percolation occurs) is the predominant

feature in agro-ecosystems when compared to exposed water-bearing formations or water bodies.

High recharge rates (e.g., rainfall, irrigation, snowmelt) combined with slow natural subsurface

drainage can lead to shallow water table (Smedema et al., 2004). Manitoba, although being in a semi-

arid to sub-humid region, is subject to seasonal shallow water tables due to snowmelt infiltration and

poor drainage (Cordeiro et al., 2015a). In contrast, deep percolation of irrigation return flow can be a

major source of groundwater recharge beneath irrigated areas in arid regions (Schmidt and Sherman,

1987).

When the water table is at or above the drainage depth, the tile drains will intercept gravitational

water (Skaggs et al., 2012) that otherwise would move downwards towards the aquifer by natural

drainage. When the percolation rate is larger than that of the natural soil drainage, the water table will

rise; the percolating water will thus drive an increase in water table elevation (i.e., water table

becomes shallower in relation to the soil surface). If the water table elevation is higher than that of the

drainage base depth (W; section 3), it will be intercepted by tile drains and gravitational water will

become tile outflow. This concept is well illustrated in intensively irrigated areas, where a leaching

fraction that is applied to leach salt out of the root zone can cause waterlogged conditions (Asawa,

2006). In such situations, drainage is used to control the waterlogging and leach salts out of the soil

profile (Smedema et al., 2004).

4.4.SURFACE RUNOFF

Recent field investigations are shaping new conceptions about the hillslope runoff generation

process. Surface runoff had previously been described as continuum process but now is perceived as

20

threshold-mediated and connectivity-controlled (McDonnell, 2003; Spence, 2010; McDonnell, 2013).

Threshold-mediated runoff in agricultural land has been known for several decades, where

depressional storage must be filled before runoff can begin (Gayle and Skaggs, 1978; Skaggs et al.,

2012). In this context, increased infiltration promoted by tile drainage reduces the proportion of total

outflow occurring as surface runoff (Walter et al., 1979; Skaggs et al., 1994; Calhoun et al., 2002;

Stillman et al., 2006; Schilling and Helmers, 2008) since runoff can be triggered by excess saturation

in the soil profile (Beven, 2011), which is delayed by increased infiltration and storage promoted by

tile drainage. However, the changes in water yield (i.e., total runoff) from a field or small watershed

due to tile drainage tends to be relatively minor, with increases around 10 % (Blann et al., 2009).

Drainage intensity of tile drainage systems varies according to field conditions, with higher drainage

rates when ponding occurs in surface depressions (Skaggs et al., 2012).

The seasonal patterns of surface runoff in cold regions need to be taken into consideration when

applying recommendations from regions where soils are not frozen throughout the winter. The

hydrology of cold regions is unusual as almost all surface runoff events are due to the spring melt of

winter snowpacks, rather than to runoff from rainfall (Shook and Pomeroy, 2010). For example, in a

two-year study in Ontario, Tan et al. (2002) found that 65 % of the surface runoff (as well as tile

drainage outflow) occurred between November and March. Moreover, runoff in the Canadian Prairies

is primarily due to spring snowmelt over frozen soils (Tiessen et al., 2010; Shook and Pomeroy, 2012;

Liu et al., 2013a; Liu et al., 2014). Cold-region processes such as frozen soils are known to affect

hydrology of tile drained fields (Luo et al., 2000). For example, accounting for cold-region processes

in tile drain simulations in Nova Scotia using DRAINMOD – one of the most widely used models for

design and performance evaluation of drainage systems (Hassanpour et al., 2011), led to more

accurate predictions of drainage flow in terms of timing and magnitude. In Manitoba, a 12-year

simulation of the water balance in tile-drainage fields having fine sandy loam soils indicated that, on

21

average, only 10 mm out of the 518 mm yearly precipitation was lost as runoff (Satchithanantham and

Sri Ranjan, 2015), suggesting that runoff amounts under tile drained conditions are small (equivalent

to <2 % of precipitation).

5. BENEFITS OF TILE DRAINAGE SYSTEMS

5.1.AGRONOMIC BENEFITS

The advantage of using subsurface drainage has long been recognized and reported in the literature.

The enhancement of agricultural use of land may take place through direct effects on crop growth, by

improving the efficiency of farm operations, or by maintaining a favourable salt regime in the soil

profile of irrigated areas (Van Schilfgaarde, 1974). In other words, the major benefit is removal of

excess water and salinity control (Bernstein, 1974). These benefits are very relevant to Manitoba,

which is the second largest potato producer in Canada (Satchithanantham et al., 2012). Due to its

sensitivity to stress caused by water excess and deficit, potato production is usually supplemented by

irrigation in Manitoba (Satchithanantham et al., 2012). Tile drainage is also becoming popular in the

potato-growing regions of the Province (Satchithanantham et al., 2014). The combined effect of

irrigation and drainage was found to increase potato yields from 15 % to 32 % in Manitoba

(Satchithanantham et al., 2012); however, the results depend on annual variations in the hydrological

cycle. Similar results were also found for corn production in Manitoba, which is often grown in

rotation with potatoes (Cordeiro and Sri Ranjan, 2012). For this crop, the benefit of drainage was

more pronounced in wet springs and the combined effect of irrigation and drainage corresponded to

yield increases around 20 % (Cordeiro and Sri Ranjan, 2012). Unfortunately, the sole effect of tile

drainage could not be directly assessed since this practice was always combined with irrigation in the

studies carried in Manitoba. Yield increases promoted by tile drainage have been attributed to

increased aeration of the root zone, enhanced transport of CO2 produced by roots and microflora,

enhanced chemical reactions, and increased soil temperature (Hiler, 1969; Wesseling, 1974).

22

Reduction of the moisture content in the upper soil layers due to lowering of the water table also

increases the bearing strength of the soil, increasing trafficability, and contributing directly to the

timeliness of field operations (Young and Ligon, 1972; Kandel et al., 2013). DRAINMOD

simulations in two soils in Ohio suggest that, all other variables remaining the same, field working

days increase with decreasing tile spacing (Nolte et al., 1983). Increasing the number of working days

in a growing season is crucial for crop yields, especially if it allows for an early planting date. For

example, the second week of May is the latest time window to plant corn and spring wheat in

Manitoba without yield losses, based on long-term seeding dates and yield information (Manitoba

Agricultural Services Corporation, 2016). A modelling study in Ohio comparing the cost of investing

in tile drainage or bigger farm machinery to achieve an early planting date found that the effect of

drainage on timeliness cost is more significant than that of machinery size (Ozkan et al., 1991). Thus,

tile drainage is an efficient strategy to achieve timely planting and avoid yield losses. However,

drainage systems with both low and very high drainage intensity may result in low remaining income

no matter what size of machinery systems are used (Ozkan et al., 1991), which emphasizes the need

for correct designs of tile drainage systems to maximize economical returns.

Another agronomic benefit of tile drainage is disease control. Excess moisture has long been

known to favor diseases in potato and wheat (Colhoun, 1973). For example, pink rot, caused by

Phytophthora erythroseptica, is a widespread potato disease observed in several countries in North

and South America, Europe, Asia, and Australia (Rowe and Nielsen, 1981). The disease, which causes

canopy wilting and tuber rotting, develops in soils approaching saturation from poor drainage or

excessive precipitation or irrigation. Development of fusarium head blight (FHB) and mycotoxin

production in wheat ears was also found to increase with increasing length of wetness period (Xu et

al., 2007). Similarly, maize ear rot in Ontario, Canada, caused by Fusarium graminearum increased

with wetness (Vigier et al., 1997; Reid et al., 1999).

23

5.2.ENVIRONMENTAL BENEFITS

The reduced surface runoff associated with tile drainage (see section 4.4) can reduce potential for

flooding and transport of nutrients with surface runoff, which helped tile drainage to be regarded as a

hydrologic control BMP (Scott et al., 1998). Fraser and Fleming (2001), reviewing the environmental

benefits of tile drainage in Ontario, stated that peak flows can also be reduced because increased

storage capacity allows more water to infiltrate, making the soil act as a buffer for rainfall and

distribute the runoff over a longer period of time. However, their review reinforced that reductions

depend on antecedent soil conditions; very wet conditions prior to precipitation can promote peak

flow reductions of 20 %, while dry conditions prior to precipitation can cause reductions as larger as

87 %. This reduction in peak flow is consistent with research in the US Midwest where tile drainage

outflow is thought to affect predominantly the baseflow portion of a stream hydrograph (Schilling and

Helmers, 2008). While surface inlets could lead to increase the contribution from tile to stormflow

(Schilling and Helmers, 2008; King et al., 2015b), the authors of this review are unaware of

widespread adoption of this practice in Manitoba. Surface inlets or intakes (risers extended from

underground pipes to the surface) remove surface water from depressed areas in fields, which provide

a direct and significant pathway for sediment, solids, and nutrients in agricultural runoff to enter

waterways via tile drainage systems (Blann et al., 2009).

Although tile drainage is not strictly a practice to control pollutant export, the reductions it

promotes in runoff at field scale can have a positive impact on water quality. Surface runoff is one of

the major causes of sediment transport and is associated with nutrient transport (Sharpley et al., 1992;

He et al., 1995; Sharpley et al., 2002; Aksoy and Kavvas, 2005; Koiter et al., 2013). In the Canadian

Prairies, sediment and nutrient transport is substantial during the spring runoff and dissolved forms

are responsible for much of the nutrient loads exported from agricultural land during this season

(Tiessen et al., 2011; Cade-Menun et al., 2013). For example, Liu et al. (2013b) studying the critical

factors affecting field-scale losses of nitrogen and phosphorus in spring snowmelt runoff in the

24

Canadian Prairies, found that 91 % of total nitrogen and 73 % of total phosphorus in the snowmelt

runoff were in dissolved forms. Similarly, Tiessen et al. (2010) studying the influence of tillage

practices on seasonal runoff and nutrient losses in the Canadian Prairies found that more than 80 % of

the total nitrogen and phosphorus in the spring snowmelt runoff were in dissolved forms. Although

potential for infiltration of snowmelt runoff may be limited by frozen soils, tiling is likely to improve

infiltration and reduce surface runoff and contact between runoff and phosphorus sources at the soil

surface.

Another benefit of tile drainage is the indirect effect on nutrient and water use efficiencies, which

minimizes the opportunity for leaching and runoff. In general, efficiencies of nitrogen fertilizers will

increase with good drainage conditions (Wesseling, 1974), thus reducing residual nitrogen that could

be subject to mobilization. Alleged reduced variability in crop yields (Stonehouse, 1995) would also

contribute to increased agronomic stability. This enhanced stability would render nutrient use more

predictable and, consequently, further promote nutrient use efficiency.

6. RISKS ASSOCIATED WITH TILE DRAINAGE SYSTEMS

6.1.AGRONOMIC RISKS

While subsurface drainage systems, by design, target keeping the water table at desired depths

during wet periods, they may intensify deficit water conditions during dry periods of the growing

season (Fouss et al., 2007). Low water table and excessive drainage can increase the risk of fire,

increase subsidence, and cause soil erosion (Irwin, 1967), but the major agronomic risk is related to

crop performance. Temporal and spatial variability of rainfall may result in excessive soil water and

related reduced crop growth early in the growing season followed by deficit soil water conditions later

in the growing season (Fouss et al., 2007; Singh et al., 2014). Wet conditions in the spring and dry

conditions in the summer were reported by Cordeiro and Sri Ranjan (2012) in the second year of their

study on the impact of irrigation and tile drainage on corn yields in Manitoba in 2011. Early excess

25

water stress in naturally drained systems often aggravates water deficit conditions later in the season

because shallow root depths caused by high water tables early in the growing season may not be

sufficiently deep to access deeper soil water needed by the crop later in the season (Fouss et al.,

2007). Again, this condition has been reported in a study of water redistribution in the corn root zone

in Manitoba for 2010, which had a wetter-than-normal spring that could have caused a shallow root

system (Cordeiro et al., 2015a). As a result, the water uptake by the corn crop during period of high

evapotranspiration was restricted to the top 0.4 m of the soil profile.

Yield reductions due to over-drainage have been reported in the literature. Wesseling (1974), based

on the results of extensive research in different crops, soil types, and soil moisture conditions, stated

that crop yields have a common response to wet and dry conditions across several soil types and

textures, where yields increase as the water depth increases up to an optimum after which yields

decrease. The reason for yield decreases at water table depths shallower and deeper from the optimum

is lack of aeration due to water excess and water deficit stress, respectively. This yield response is

illustrated with results from different field investigations, with the exception of some vegetables

(Table 3).

Table 3. Crop yields at varying water table depths [adapted from Wesseling (1974)]

Crop Water table depth (m)

0.15 0.30 0.40-0.50 0.60 0.75 0.80-0.90 1.0 1.2 1.5

Relative yield (%) .

Ladino clover 100 97 88 – – – – – –

Orchardgrass 100 70 92 – – – – – –

Fescue 100 70 92 – – – – – –

Alfalfa – – – 100 – – – 97

Corn – 41 82 85 100 85 45 – –

Soybeans 64 63 78 100 86 – – – –

Potatoes – – 90 100 – 95 92 – 96

Grain sorghum 73 86 93 100 93 – – – –

Millet 41 69 80 87 98 100 93 – –

Research in North America also shows similar trends. Doty et al. (1984), discussing different

studies on the impact of over-draining in North Carolina, argued that drought stress may arise in

26

drained areas. According to those authors, soils with little water holding capacity (e.g., sandy soils),

which are drained too rapidly or too deeply, may develop drought stress within 4 to 7 days. They

stated, for example, that lowering the water table from 0.8 to 1.0 m below the soil surface for corn and

soybean growing in sand, sandy loam, and silt clay loam soils can reduce yields from 45 to 100 % of

potential yield. Recent modelling exercises in sandy loam soils of Manitoba using DRAINMOD

suggest that, while corn yield losses are more frequently caused by water excess, which justifies the

adoption of tile drainage in the Province, excessive drainage can also result in poor crop performance

(Cordeiro and Sri Ranjan, 2015). In their 20-year simulation, the authors found that excessive

drainage resulted in yield loss in 3 years (15 % of the time), and that losses decreased with increasing

tile drain spacing.

6.2.ENVIRONMENTAL RISKS

The hydrological alterations of tile drainage discussed in section 4 can promote environmental

benefits (sub-section 5.2) but also represent a risk for nutrient export. The following sub-sections will

focus on the movement of the two major macronutrients for crop production, namely, nitrogen and

phosphorus (Fageria, 2008). Doubling the agricultural food production during the past 35 years of the

20th

century was only possible with a 6.9-fold increase in nitrogen fertilization and a 3.5-fold increase

in phosphorus fertilization (Tilman, 1999). These are also the major nutrients associated with

eutrophication of aquatic ecosystems (Conley et al., 2009), although phosphorus usually limits

productivity of freshwater lakes (Schindler, 1977) and is thought to be the major cause of

eutrophication in Lake Winnipeg (Schindler et al., 2012).

Nitrogen and phosphorus can be in different pools (i.e., organic or inorganic) and forms [i.e.

dissolved and particulate; Bianchi et al. (2013)]. These different nutrient forms can be mobilized

through different transport mechanisms (e.g., overland flow and through the soil profile). Thus, the

nutrient source can play a role in nutrient movement through tiles. The major sources of nutrients for

27

crop production are manure (e.g., livestock manure, green manure, compost) and synthetic fertilizers

(Mongillo and Zierdt-Warshaw, 2000), which vary in composition and proportion among the different

pools and forms. The remainder of this work will focus on livestock manure and synthetic fertilizer,

which also differ in cost (Mooleki et al., 2002; Girard and Girard, 2014). Combinations of these

differences in nutrient sources, along with contrasting application rates and timing, have direct

implications for nutrient transport. Furthermore, due to its animal origin, livestock manure can also be

a source of pathogen contamination (Pachepsky et al., 2006). Thus, investigation of pathogen

movement through tile drainage is also important from an environmental and public health standpoint.

6.2.1 Nutrient movement (matrix and preferential flow)

6.2.1.1 Manure

While livestock manure from different sources can be a valuable resource for crop production

(Eghball and Power, 1994; Moore et al., 1995; Fleming et al., 1998a), manure management can

become a problem due to intensification of animal production systems (Sims et al., 2005). For

example, in Canada, inventories of cattle, pigs, and poultry increased by 113 %, 144 %, and 171 %

between 1982 and 2002, respectively, after a similar intensification between 1962 and 1982. Such

increases may create a nutrient management issue, unless managed correctly, that leads to long-term

over-application of manures to cropland due to the lack of economically viable off-farm uses for

manure (Sims et al., 2005). Over-application springs from practical limits imposed by the distances

that manure can be hauled and its cost (Kellogg et al., 2000; Nguyen et al., 2013). Moreover,

synthetic fertilizer applications are carried out without giving credit for nutrients already applied in

manure, which leads to soil fertility levels that exceed agronomic requirements (Beegle et al., 2000).

Nutrients from manure can be transported by surface runoff (Sharpley, 1997; Schroeder et al.,

2004; Smith et al., 2007) and by leaching through the soil profile (Eghball and Power, 1994; Beegle et

al., 2000; Sims et al., 2005). While tile drainage can potentially reduce losses by reduced runoff (sub-

28

section 4.4), it may enhance leaching, resulting in potentially increased loss of nutrients in tile

outflow. Research regarding nutrient export by tile drainage from manured fields is not available for

Manitoba but there are a few studies conducted in more humid regions of eastern Canada that provide

insight on important factors influencing movement of nutrients from manured land to tiles.

Among these factors, manure application rate seems to be a prominent one. A 6-year study in

Ontario, where liquid manure was applied to continuous corn at 224, 560 and 897 kg N ha-1

every

year found that average nitrate and orthophosphate concentrations in tile drainage increased with

application rate (Phillips et al., 1981). Also, Fleming et al. (1998b) monitored nitrate levels in tile

drainage water between 1995 and 1998 in 20 farms in Ontario under a variety of farm management

systems and where the most commonly grown crops were corn, soybeans, and wheat. They found that

the mean tile water concentration of nitrate-N for the manured lands (26.5 mg L-1

) was significantly

higher than the concentrations for land not receiving manure (13.8 mg L-1

). Likewise, mean

concentrations of total phosphorus were significantly higher for manured land (1.05 mg L-1

) when

compared to non-manured land (0.29 mg L-1

). The authors attribute the higher concentrations

observed for manured land to over-application of nutrients. Similarly, Ball Coelho et al. (2007)

measured higher concentrations of nitrogen and phosphorus in tile outflow after liquid swine manure

was applied to standing corn in Ontario. These authors found that injection of manure at 75 and 94 m3

ha-1

led to instantaneous detection of ammonium (nitrogen) and dissolved reactive phosphorus in

samples, while concentrations of these analytes were much lower at the application rate of 56 m3

ha-1

,

even when tiles flowed freely during manure application. Concentrations of ammonium-N and

dissolved reactive phosphorus were about 5, 40, and sevenfold greater at the highest than at the

second highest application rate of injected manure in 2000, 2001, and 2002, respectively, but did not

vary over time at rates below 56 m3

ha-1

. Concentrations of dissolved reactive phosphorus were as

high as 4.62 mg L-1

for application rates above 56 m3

ha-1

but were smaller than 1.3 mg L-1

for

29

application rates at or below this threshold. Similar trends of increased nutrient concentrations in tile

outflow for higher manure application rates have also been reported in the New York state (Hergert et

al., 1981), North Carolina (Evans et al., 1984), and Iowa (Cook and Baker, 2001). Experiments in

Manitoba also showed that leaching of nitrate in sandy soils increased with increasing manure

application rates (Nikièma et al., 2013). The results showed leaching of 47.7, 51.4 and 77.4 kg ha-1

from plots receiving low (64 kg N ha-1

), medium (128 kg N ha-1

) and high (192 kg N ha-1

) rates of

manure application, respectively, while leaching from control plots was much less (i.e., 23 kg ha-1

).

Notably, from the three years of experiment leaching occurred only when precipitation was above

normal, indicating that frequency of leaching is likely to be low, but will be dependent on weather

pattern

In addition to application rate, manure type may also influence nutrient export through tiles. For

example, Tan et al. (2009), investigating nitrate concentrations in tile outflow from plots receiving

either solid and liquid cattle manured applied on a phosphorus-basis of 100 kg P ha-1

and 200 kg N ha-

1, found higher concentrations for solid manure (flow-weighted mean of 11.2 mg L

-1) than those for

liquid manure (flow-weighted mean of 7.3 mg L-1

). The authors argued that nitrogen in liquid manure

could have been lost as ammonia-N, which could explain the lower nitrate concentration in tile

drainage outflow in this treatment. Also, weekly monitoring of tile drainage outflow from 39

agricultural fields of varying textures and under variety of crops in Nova Scotia between 2002 and

2003 (Mehlich-3 soil test P ranging from 29.7 to 433.8 mg kg-1

) indicated that poultry and swine

manure contributed to high phosphorus content in soils, leading to consistently high total phosphorus

concentrations comprised mostly of soluble reactive forms (Kinley et al., 2007). The concentrations

of total phosphorus among all the fields ranged from virtually zero to 5.5 mg L-1

, but no averages

were reported. These results corroborate experiments conducted in Manitoba where the proportion of

total labile phosphorus in manure samples was larger for liquid swine manure than for solid cattle

30

manure, which influenced leaching and runoff potential (Kumaragamage et al., 2012). Similar

comparisons in the literature among different manure types were not found for other parts of North

America. However, a study in South Eastern United Kingdom found that pig slurry showed increased

concentrations of dissolved phosphorus in the first monitoring season when compared to untreated

control and other manure types that included broiler litter and cattle farmyard manure (Hodgkinson et

al., 2002). The authors of that study in the UK highlighted the role of preferential flow in nutrient

transport, which could have been exacerbated by the recent installation of the tile drainage system

immediately prior to monitoring.

Preferential flow has been, in fact, reported to enhance nitrogen and phosphorus movement from

annual cropland receiving manure to tile drains in Ontario for liquid swine manure (Ball Coelho et al.,

2007). Movement of liquid municipal biosolids from agricultural fields to tile drains has also been

observed to be expedited by preferential flow in Ontario (Lapen et al., 2008). Based on studies in

Quebec and the UK, phosphorus transfer through preferential flow pathways may be particularly

important after storm events that rapidly follow periods or drought and/or surface input as manure

since preferential flow exerts considerable control on vertical flow rates, especially when rainfall

follows a period of drought (Simard et al., 2000). However, these conclusions may not apply to the

prevailing conditions in the Canadian Prairies, where most of the runoff is generated due to snowmelt

(Shook and Pomeroy, 2010). Investigations of vadose zone flow in cracked heavy textured soils in the

Canadian Prairies indicate that water is flushed down rapidly via cracks and fissures in the root zone

during snowmelt, and that water moves downward by diffusion-dominant advective-diffusive flow

mechanisms in late spring after the entire soil profile thaws (Joshi and Maulé, 1999). Thus, manure

applications to tiled land should be target to periods when soil cracks have disappeared in the spring.

This is particularly important to Manitoba, where it has been estimated that more than 40 % of

farmlands had moderate to very high likelihood of crack flow, which is the highest likelihood among

31

the Prairie Provinces (i.e. Saskatchewan and Alberta) and one of the largest in Canada (Dadfar et al.,

2010).

Other studies on preferential flow of manure conducted in the United States are also very

informative for understanding under which conditions preferential flow to tile drains occur. For

example, intricate experiments to assess if burrows created by Lumbricus terrestris L. earthworm

species can contribute to rapid movement of injected manure to tile drains in Ohio retro-fed smoke

into tile drains using a turbine blower and measured infiltration rates using dyes (Shipitalo and Gibbs,

2000). The results showed that dyed water was observed in the tile when added to smoke-emitting

burrows, but not when added to burrows that did not produce smoke. The authors concluded that

earthworm burrows in close proximity to tile lines may expedite transmission of injected wastes

offsite. They suggest that movement of injected wastes to tiles via earthworm burrows and other

preferential flow paths may be reduced by using precision farming to avoid manure application near

tile lines or by modifying application procedures. Targeted application and modification of application

procedures, however, may not be practical for current Manitoba conditions due to lack of precision

technology, inadequacy of current equipment design specifications, and difficulties imposed by

equipment maneuvering.

By analogy to the study on biopore preferential flow carried out by Shipitalo and Gibbs (2000), the

risk of nutrient movement from manured land by preferential flow paths of different origins (e.g.,

cracking soils) would be greater if those paths were hydraulically connected to tile drains. Moreover,

Hoorman and Shipitalo (2006) compiled a database of liquid animal waste violations entering tile

drains in Ohio between 2000 and 2003 and found that out of 98 cases, 21 violations were related to

preferential flow (i.e., 13 cases due to dry, cracked soils and 8 cases due to earthworm burrows; Table

4). Notably, most violations were due to management issues such as timing of application (i.e., excess

rain near application or application on saturated soils), application rate (i.e., over-application or

32

application error), or manure storage management. Applying manure when tile drains were flowing

were also a cause of violation in 20 % of the cases, which highlights the importance of avoiding

manure application under these conditions. For cold-climate regions with a well-defined “drainage

season” such as Minnesota, this period spans from March to June (Jin and Sands, 2003). A drainage

season has also been suggested for Manitoba but has not been defined (Satchithanantham et al.,

2014). Based on weather and physiographic similarities, this season in Manitoba is likely similar to

that proposed for Minnesota. This well-defined drainage season contrast to observations in Ontario

(Van Esbroeck et al., 2016) and Ohio (King et al., 2015a), where drainage can occur during the

winter, although P concentrations being higher between March and June in the case of Ohio.

Table 4. Reasons why liquid animal wastes entered tile drains in Ohio between

2000 and 2003 [adapted from Hoorman and Shipitalo (2006)] Cause Number of cases out of 98

Excess rain or saturated soils 41

Over-application or application error 35

Manure storage management 33

Ponding manure or excessive irrigation 26

Drainage lines flowing or plug failure 20

Broken tile or shallow drainage 14

Equipment or storage failure 13

Dry, cracked soils 13

Feedlot runoff 11

Snowmelt 10

Eggwash water (thin waste, low solids) 10

Earthworm burrows 8

6.2.1.2 Synthetic fertilizer

The application of nitrogen and phosphorus fertilizers has steadily increased since the 1960s, a

trend that will likely continue due to the large demand for grains projected for 2050 (Tilman et al.,

2002). Without the use of synthetic fertilizers, the world food production could not have increased at

the rate it did in the last decades of the 20th

century and more natural ecosystems would have been

converted to agriculture (Tilman et al., 2002). A summary of results from long-term studies in the

USA, England, and the tropics, along with the results from an agricultural chemical use study and

33

nutrient budget information suggest that at least 30 % to 50 % of crop yield is attributable to synthetic

fertilizer nutrient inputs, and that is probably a conservative estimate (Stewart et al., 2005).

Application of synthetic fertilizers is more prevalent in the developed north (i.e. United States,

Canada, and much of Europe) as well as parts of India and China (Potter et al., 2010). As a result, the

negative impact of synthetic fertilizer to water quality due to export through tile outflow is well

documented for some regions of North America.

In Canada, there are only a few studies reporting nutrient export through tile in the western

provinces since this practice is relatively new in this region. For example, assessment of nutrient

export from potato fields in Manitoba report flow-weighted mean nitrate concentrations in freely-

draining tile outflow ranging from 64 to 70 mg L-1

between 2010 and 2011 (Satchithanantham et al.,

2014). However, the maximum values reported were 161 and 92 mg L-1

in 2010 and 2011,

respectively. The exported loads of nitrate were 161 and 73 kg ha-1

in 2010 and 2011, respectively.

Higher concentrations of nitrate in 2010 could be potentially associated with higher fertilizer

application rates since 152 kg N ha-1

were applied in that year, while only 77 kg N ha-1

were applied

in 2011. These higher concentrations, combined with higher-than-normal precipitation in 2010 led to

larger exported nitrates loads in that year (Satchithanantham et al., 2014), which reinforces the

influence of annual variability in weather on the environmental performance of tile drainage. Flow-

weighted mean concentrations and exported loads of phosphate followed a trend similar to that

observed for nitrate. Flow-weighted mean phosphate concentrations were 0.56 and 0.14 mg PO4 L-1

in

2010 and 2011, respectively, while exported loads were 1.28 and 0.15 kg PO4-P ha-1

. However, an

inverse correlation was observed between fertilizer application rate and phosphate export since higher

concentrations and loads in 2010 corresponded to a reduced application of phosphorus in that year (61

kg P ha-1

) than in 2011 (67 kg P ha-1

). While the application rates were not substantially different, they

suggest that higher phosphorus application rates can result in reduced export, highlighting the

34

complex processes involved in phosphorus export through tile drains (i.e., presence of different

phosphorus pools in the soil; section 7.2.2).

Another study reporting nutrient export from tile-drained corn fields in Manitoba indicates the

same trends observed for potatoes (Cordeiro et al., 2014). The flow-weighted mean nitrate

concentration in tile outflow was larger in 2010 (98.9 mg L-1

) than in 2011 (41.2 mg L-1

). Increased

concentrations associated with larger drainage volumes in 2010 led to increased exports of nitrate in

that year (138 kg ha-1

) when compared to 2011 (36 kg ha-1

). Concentrations and loads of phosphate

followed a similar trend. Flow-weighted mean phosphate concentrations in tile drainage were 0.4 mg

L-1

and 0.32 mg L-1

in 2010 and 2011, respectively, while exported loads were 0.6 kg ha-1

and 0.27 kg

ha-1

in 2010 and 2011, respectively. Again, the influence of weather was significant factor on nutrient

export since 2010 was wetter-than average, but fertilizer application rates also seem to have played a

role. A total of 150 kg N ha-1

was applied in 2010 to the corn fields as urea (46-0-0) and ammonium

polyphosphate (10-34-0), while the amounts of N fertilizer decreased to 122 kg N ha-1

in 2011.

However, different than the inverse phosphate trend reported by Satchithanantham et al. (2014) for

potatoes, the concentrations and loads of phosphate in corn fields followed fertilizer application rates.

In 2010, 60 kg P ha-1

were applied as ammonium polyphosphate (10-34-0), while only 34 kg P ha-1

were applied in 2011. These results support the premise that increased exports and concentrations of

nitrate and phosphate in tile outflow will be linked to increased application rates of synthetic fertilizer,

which is similar to the trend observed for manure (sub-section 6.2.1.1), although other aspects

pertaining to field hydrology such as seasonal timing of flow will also be influential and interact with

nutrient availability in the soil.

The nitrate loads exported from potato (Satchithanantham et al., 2014) and corn fields (Cordeiro et

al., 2014) in Manitoba (>130 kg ha-1

) were larger than most values reported for Canada (which ranged

from 7 to 39 kg ha-1

·yr) and the US [which were as high as 122 kg ha-1

·yr in Minnesota; Randall and

35

Goss (2008)]. They were also close to (corn) and larger than (potato) the amount of N fertilizer

applied, which suggests the flush of nitrate accumulated in the soil profile. This possibility was

discussed by Cordeiro et al. (2014), who attributed the larger amounts exported in 2010 to the high

amounts of nitrate accumulated in the soil profile being flushed through the newly installed drainage

system by the excessive precipitation in that year. Nutrient excess in the soil profile is in agreement

with the cropping history of the field, which was planted potatoes, corn, canola, and wheat. However,

exported loads in 2011 (36 kg ha-1

) decreased to the range of values reported in the literature for

North America (Randall and Goss, 2008). These results suggest increased nutrient export in the first

year of installation of tile drainage systems in the Canadian Prairies, which has also been reported in

Europe for farmland receiving manure fertilizers (Hodgkinson et al., 2002)

Nutrient export from synthetically fertilized land through tile drains is better documented in Eastern

Canada, which has a longer history and wide adoption of tile drainage. Surveys in southern Ontario in

the 1990’s indicated that 87 % of farmers were using tile drainage, which was promoted by both

government subsidies and alleged reduced variability in crop yields (Stonehouse, 1995). The yearly

studies on nutrient export from synthetically fertilized land in Ontario showed similar trends to those

observed for manure, where higher application rates lead to higher nutrient concentrations in tile

outflow. For example, total nitrogen and total phosphorus concentrations in tile drainage effluent were

monitored in Ontario by Bolton et al. (1970) over a 7-year period for three cropping systems (i.e.

continuous corn, continuous bluegrass and a four-year rotation of corn-oats-alfalfa-alfalfa). One

treatment received no fertilizer application while the other received 5-20-10 fertilizer at the rate of

336 kg ha-1

per year, applied to all crops except the first- and second-year alfalfa in the rotation. Corn

plots in the fertilized treatment received an additional application of N, side-dressed at 112 kg ha-1

per

year. The results showed that the mean total nitrogen concentration in tile outflow for all three

cropping systems was 8.1 mg L-1

for the fertilized plots, while the control plots had concentration of

36

6.4 mg L-1

. Likewise, mean total phosphorus concentrations were 0.21 and 0.18 mg L-1

for fertilized

and unfertilized plots, respectively. Mean nutrient losses followed the same trend and were larger for

fertilized plots (8.1 kg N ha-1

; 0.19 kg P ha-1

) when compared to control plots (4.4 kg N ha-1

; 0.12 kg

P ha-1

).

An assessment of nutrient losses through tile drains from two potato fields receiving synthetic

fertilizer in Quebec reported total nitrogen and total phosphorus concentrations ranging from 1.7 to

40.0 mg L-1

, and from 0.002 to 0.052 mg L-1

, respectively (Madramootoo et al., 1992). Although not

explicitly discussed in the study, the data suggests that phosphorus in tile drainage is connected to

fertilizer application rate. Further analysis of the results reported by Madramootoo et al. (1992) for

two years and two fields (n=4) indicated a strong correlation between annual P fertilizer rate and

mean annual concentrations of total phosphorus in tile outflow (r2=0.87). Such correlation was not as

strong for total nitrogen (r2=0.46), although nitrogen concentrations in tile outflow increased

following synthetic fertilizer application. These results are consistent with increased nitrate leaching

with increasing fertilization rates in Europe (Bergström and Brink, 1986) and the US (Gast et al.,

1978).

A 43-year long-term study in southern Ontario comparing phosphorus losses through tile drains

from a clay loam soil under different cropping systems found that flow-weighted mean concentrations

of total phosphorus was larger for fertilized fields compared to unfertilized ones, although

performance varied for different cropping systems (Zhang et al., 2015). For example, plots under

continuous corn receiving 16.8 kg N ha-1

as ammonium nitrate and 67.2 kg P2O5 ha-1

as triple

superphosphate each spring had flow-weighted mean concentrations of total phosphorus 33 % higher

than the unfertilized plots. These results are consistent to other recent studies in Ontario where

elevated P losses were observed for fields under reduced tillage receiving either manure or synthetic

fertilizer applied in the fall when compared to fields not receiving fertilizer (Van Esbroeck et al.,

37

2016). Interestingly, the long-term comparison among cropping systems reported by Zhang et al.

(2015) showed that total phosphorus in tile drainage water from plots receiving fertilizer were larger

for continuous grass, followed by the corn–oats–alfalfa–alfalfa rotation, and by continuous corn. The

same trend had been observed by Culley et al. (1983) in Ontario, where mean total phosphorus

concentrations in tile discharge from Bluegrass sod (0.63 mg L-1

) were almost three times larger than

those from continuous corn (0.23 mg L-1

).

These differences between perennial and annual cropping systems could have been caused by

enhanced infiltration of the former. Improved infiltration of perennial forages have been reported in

the Canadian Prairies (van der Kamp et al., 2003; Kahimba et al., 2008), which has been attributed to

preferential flow pathways promoted by the development of macropores, such as root holes,

desiccation cracks, and animal burrows (van der Kamp et al., 2003). However, preferential flow has

also been observed in annual cropland. For example, soil loss promoted by preferential flow was

observed in Ontario for tile-drained corn under different tillage treatments (Gaynor and Findlay,

1995). Also, Beauchemin et al. (1998) investigating forms and concentration of phosphorus in

drainage water of 27 different soils in Quebec in 1994 and 1995, reported total phosphorus

concentrations ranging from 0.01 to 1.17 mg L-1

, resulting in the water quality standard of 0.03 mg L-1

for total phosphorus being exceeded in 14 out of 27 sites in the first year but only in six out of 25 sites

in the second year. The authors also reported a trend for phosphorus to be exported from P-rich,

cracking clay soils, in which case particulate phosphorus was the dominant form, and argue that

preferential flow is the leading cause for the phosphorus export in these soils. These results agree with

more recent studies in Quebec that report preferential flow as an important mechanism of phosphorus

transport through tile drains in clay loam soils (Eastman et al., 2010). This study also showed that

phosphorus speciation was linked to soil texture, since clay loam soils lost 80 % of their total

phosphorus in particulate form (filtered through a 0.45 µm cellulose filter), while that fraction

38

corresponded to only 20 % of the total phosphorus lost from sandy loam soils. Thus, most of

phosphorus lost from tiled drained, clayey soils was in particulate form through preferential flow,

which has been confirmed by other studies in Quebec (Jamieson et al., 2003). This trend is still to be

studied in the Canadian Prairies, where dissolved forms are responsible for much of the nutrient loads

exported from agricultural land through surface runoff (section 5.2).

The effect of preferential flow on nutrient export through tile has also been observed in the US.

Experiments using tracers and hydrograph separation techniques in Iowa found that although

preferential flow contributed less than 2 % of the total drain outflow, this mechanism was responsible

for considerable fraction of the loads of tracers leached (i.e., between 9 % and 24 %) during two

irrigation events (Everts and Kanwar, 1990). Another study in Iowa attributed differences between

nitrate-N concentrations from different tillage systems to preferential flow (Bjorneberg et al., 1996a).

According to the authors, simple dilution (i.e., high drain flow results in lower nitrate-N

concentration) could not explain the results since plots with the lowest drainage volumes had the

lowest concentrations in most cases. They argued that water bypassing the soil matrix through

preferential flow had less opportunity to pick up nitrate en route to the drains when compared to

matrix flow. Thus, in this case, the effect of preferential flow on nitrate leaching seems to be opposite

to that on phosphorus leaching (i.e., preferential flow decreases nitrate concentration in drainage

outflow).

6.2.1.3 Manure versus synthetic fertilizer

The discussion on sub-sections 6.2.1.1 and 6.2.1.2 indicates that management aspects are the key

factors driving nutrient export through tile drains, rather than fertilizer source. Some research trials

directly comparing exports from manure versus synthetic fertilizer help illustrate this point. In the

UK, for example, total P in drainage water from plots receiving manure was lower than that from

plots receiving inorganic fertilizer containing similar amounts of Olsen-P concentrations in the soil,

39

although the reasons for this behaviour could not be explained (Heckrath et al., 1995). Also, Randall

et al. (2000), comparing nutrient losses in subsurface drainage water from corn fields receiving either

dairy manure or urea in Minnesota, concluded that nitrate-N losses in subsurface tile drainage did not

differ between the urea and manure treatments. Moreover, the mean total phosphorus concentration

was not different for the two nutrient sources. Results from a long-term (1998 to 2009) experiment in

Iowa comparing poultry manure and urea ammonium nitrate applied to loamy soils in a corn-soybean

rotation found that nitrate leaching was greater for plots receiving 168 kg N ha-1

as synthetic fertilizer

than plots receiving poultry manure at the same rate (Nguyen et al., 2013). However, poultry manure

applied at 336 kg N ha-1

resulted in even greater nitrate losses. Again, these results support the

discussion in previous sub-sections about the effect of fertilizer rates on nutrients export through tile.

Finally, experiments done in Ontario between 2008 and 2012 compared nitrate export through

uncontrolled tile drains from plots having clay loam soils under a corn-soybean rotation and receiving

200 kg N ha-1

·yr either as liquid cattle manure, solid cattle manure, or ammonium nitrate (Tan et al.,

2015). The results indicated that the average nitrate loss was larger for synthetic fertilizer (33.3 kg

ha-1

) and smaller for liquid cattle manure (20.4 kg ha-1

), with solid cattle manure showing

intermediate values (24.3 kg ha-1

). The same trend was also observed for cumulative loss, where plots

receiving synthetic fertilizer, solid cattle manure, and liquid cattle manure exported 135, 98, and 81 kg

of nitrate over the four years of experiment, respectively.

6.2.2 Pathogen transport (matrix and preferential flow) While manure and synthetic fertilizer have similar potential for nutrient export through tile drains,

manure has an inherent risk associated with the movement of pathogenic bacteria. Animal manure

may contain pathogenic microorganisms such as Salmonella, Escherichia coli, and Campilobacter,

which can be spread by application manure onto land (Bicudo and Goyal, 2003). Guan and Holley

(2003) recommend that untreated manure should not be applied to fields where there is a potential for

40

runoff, since most pathogens can survive freezing or low temperatures for long periods of time.

The risk of pathogen transport is not strictly associated with tile drainage since it can also occur due

to overland runoff; however, tile drainage can enhance pathogen transport to surface water due to

leaching (Reddy et al., 1981). Much of the risk seems to be associated with management issues. For

example, a study in Ontario using a nalidixic acid resistant strain of E. coli (E. Coli NAR) as a

biotracer in liquid manure applied at 56 m3 ha

-1 found no connection between transport of bacteria

and flow rate in the tile drains (Joy et al., 1998). Rather, the strongest association was observed

between bacteria concentrations in the tile outflow and precipitation after manure application. These

results agree with those reported for Ontario by Patni et al. (1984), who found that precipitation had a

greater influence on bacterial counts in tile outflow than manure application in fields receiving 110 m3

ha-1

of dairy cattle liquid manure in the spring and 70 m3 ha

-1 in the fall. The authors found that fecal

bacterial counts from manured land were comparable to land not receiving manure due to

contamination from wildlife, which has also been observed in control plots (i.e., not receiving

manure) from studies in Iowa (Cook and Baker, 2001). However, low numbers in manured plots

reported by Patni et al. (1984) could be due to manure storage since 60% of the manure was applied

in the spring after being winter-stored, which had much lower bacterial concentrations than the

relatively fresh manure applied in the fall. Similarly, spreading of liquid manure in Ontario resulted in

no change in bacterial quality of tile water when the tiles were not flowing at the time of application,

and insignificant changes in bacteria counts were observed when the soil had been tilled prior to

manure application (Dean and Foran, 1992). Results reported by Ball Coelho et al. (2007) also

indicate that counts of E. coli in tile water occurred between 1 and 3 hours following application rates

larger than 74.8 m3 ha

-1. Notably, these application rates were large enough to induce tile drainage

outflow. Thus, the effect of high application rates of liquid manure in this case can be compared to

that of precipitation following application. The quick appearance of bacteria in tile outflow was also

41

observed in Nova Scotia, where bacterial levels were normally highest at the beginning of each

drainage flow event and then decrease as the event progresses, although viable fecal bacteria could be

recovered from both runoff and tile drainage water many months after manure application (Stratton et

al., 2004). Thus, the risk is associated with any farmland receiving manure, not only tile-drained land.

Timing of application was also a significant factor in bacteria counts in tile drainage water. The

results from studies in loamy soils in Iowa showed that Enterococcus numbers were significantly

higher in plots where manure had been applied in late winter, when compared to plots receiving

manure in the fall or spring, as well as plots receiving urea‐ammonium nitrate fertilizer (Pappas et al.,

2008). These results were attributed to preferential flow during the winter since the manure had to be

broadcast to frozen soil rather than injected. The authors argue that preferential flow through

macropores is a major pathway for bacterial leaching, and that fall and spring applications by

injection disrupted this pathway, resulting in decreased bacterial counts.

The prominence of macropores as the primary mechanism by which bacteria from surface-applied

manure can be transported through the soil profile has been widely described (Jamieson et al., 2002)

since most cells are either adsorbed to soil particles or filtered through the soil matrix (Reddy et al.,

1981). However, recent studies indicate that macropores can also attenuate the bacterial counts of the

water percolating through them. For example, Fox et al. (2012) described the attenuation of bacterial

contamination during downwards transport towards the tile drains in Iowa. Direct injection of diluted

swine manure in two naturally occurring, surface-connected macropores that penetrated to the

subsurface drain depth indicated that the soil surrounding the macropore filtered approximately 90 %

of the E. coli that entered the macropore at the soil surface. However, the macropores identified in the

experiment were created by roots, earthworms, and other biological activities (i.e., biopores), which

may be different from the macropores created by soil cracking in terms of bacteria load reductions.

Some additional aspects that may play a role in bacteria transport include macropore scale and shape

42

(Fox et al., 2012).

7. BENEFICIAL MANAGEMENT PRACTICES (BMPS) FOR MAXIMIZING THE

AGRONOMIC BENEFITS AND MINIMIZING THE ENVIRONMENTAL RISKS OF

MANURE APPLICATION ON TILE DRAINED LAND Nutrient export from agricultural land, which is a nonpoint source, is an issue not particularly

linked to tile drainage (Sharpley et al., 1994; Carpenter et al., 1998; Daniel et al., 1998; Howarth et

al., 2002). For example, tile is only one of many factors influencing phosphorus loss from agricultural

systems (Table 5). While the list provided in Table 5 is not exhaustive and does not include important

agronomic practices (e.g., tillage, crop rotation, cover, crops), it does illustrate the multiform nature of

nutrient export from farmland. In fact, recent research in Ontario indicates a synergistic effect among

management practices and suggest that these practices should be considered together to avoid

phosphorus losses from agricultural land (Van Esbroeck et al., 2016).

Table 5. Factors influencing phosphorus loss from agricultural systems

[adapted from Howarth et al. (2002)] Factor Description

Surface runoff Water has to move off or through a soil for P to move

Tile outflow P can leach through the soil in sandy, organic, or P-saturated soil

Soil texture Influences relative amounts of surface runoff and tile outflow

Irrigation runoff Improper irrigation management can induce surface runoff and erosion of P

Connectivity to stream The closer the field to the stream, the greater the chance of P reaching it

Proximity of P-sensitive water Some watersheds are closer to P-sensitive waters than others

Sensitivity P input Shallow lakes with large surface area tend to be more vulnerable to

eutrophication

Soil P As soil P increases, P loss in surface runoff and tile outflow increases

Applied P The more P (fertilizer of manure), the greater the risk of P loss

Application method P loss increases in this order: subsurface injection, plowed under, and surface

broadcast with no incorporation

Application timing The sooner it rains after P is applied, the greater the risk of P loss

This section discusses some of the beneficial management practices that have shown positive

results in terms of nutrient export from tile drained land. It builds on the previous sections and

presents BMPs that apply to several aspects of tile-drained agricultural systems. The first sub-section

(i.e., controlled tile drainage) pertains to enhanced design an operation of tile drainage for water

conservation and decreased nutrient export. The second sub-section discusses nutrient management,

43

which was identified as a major aspect affecting nutrient export from tile drained land (section 6.2.1).

The third sub-section presents agronomic practices regarded as BMPs not only in the context of tile

drained land but also in a broader agricultural context. Finally, end-of-pipe BMPs are discussed in the

last sub-section.

7.1.CONTROLLED TILE DRAINAGE

Controlled tile drainage, also referred to as water table management, started to be developed in the

mid 1980’s in the US to conserve water resources, use natural rainfall effectively, reduce plant

stresses from excess or deficient soil water, reduce energy costs, and maximize operating profits

(Fouss and Reeve, 1987). However, controlled drainage is also a promising practice to reduce the

nutrient export from tile-drained fields to surface water; in contrast to uncontrolled tile drainage,

controlled drainage restricts tile drain discharge through water flow control structures that can create a

higher water table and reduce tile outflow by emulating a shallower drain (Frey et al., 2013).

Initial trials in Manitoba have shown controlled drainage to be very effective to reduce nutrient

export from farmland. For example, average nitrate-N export from controlled drained plots under

potato was reduced by 98 % in 2010 (larger precipitation than the 30-yr long-term average) and 67 %

in 2011 (precipitation similar to the 30-yr long-term average) when compared to tile drained plots

having uncontrolled outflow (Satchithanantham et al., 2014). Similarly, the average phosphate export

was reduced by 94 % in 2010; however, it was 15 % higher than uncontrolled drained plots in 2011.

Reductions in nutrient export were caused by reductions in drainage outflow. On average, controlled

drainage reduced the flow volume by 91 % in 2010 and 54 % in 2011 compared to uncontrolled

drainage. The large reductions in flow and consequently in exported loads observed in 2010 could be

due to delays in monitoring, which started in May 31st and did not capture the peak flows caused by

spring snowmelt. Thus, the very good performance for 2010 should be seen with caution. Also, the

larger export of phosphate in 2011 was mainly due to higher concentrations in the controlled drained

44

plots, despite their reduced lower flow volume. The operation of the system may also have influenced

these results since controlled drainage was operated as free drainage until June, which resulted in

drainage outflow being 22 % higher during the spring due to the closer spacing of this treatment (8 m)

compared to uncontrolled drainage [15 m; Satchithanantham et al. (2014)]. Operation as uncontrolled

drainage during this spring is therefore a critical factor for water conservation and nutrient loss. A

balanced approach between conserving and draining excess moisture is necessary to maximize the

benefits of controlled drainage; the transition from uncontrolled to controlled drainage mode in each

growing season should be timed so that controlled drainage can conserve soil moisture without

causing any adverse effects (Satchithanantham et al., 2014).

Controlled drainage showed similar performance under corn in Manitoba (Cordeiro et al., 2014).

For 2010, >99 % reduction in nitrate loads and 87 % reduction in phosphate loads were observed in

the controlled drainage plots due to a 97 % reduction in drainage outflow. However, the late operation

of the system after the peak flows during spring snowmelt and the unusually high exports in that year

(i.e., > 130 kg NO3-N ha-1

) due to the newly installed tile drainage system render these results less

typical than what would be expected. In 2011, results were within the expected range, showing a 71 %

reduction in nitrate export and 69 % reduction in phosphorus export due to the combined effect of

reduced analyte concentrations and 39 % reduction in drainage outflow.

The reductions in nitrate export promoted by controlled drainage in Manitoba agree with those

reported for humid regions of eastern Canada. For example, Drury et al. (2009), comparing nitrate

export from plots in Ontario under a corn-soybean rotation featuring uncontrolled tile drainage

(UTD), controlled tile drainage (CD), and controlled tile drainage with subsurface irrigation (CDS),

found that CD and CDS reduced average annual nitrogen losses via tile drainage by 44 % and 66 %,

respectively, relative to UTD in plots receiving 150 kg N ha–1

applied to corn as 18–46–0 synthetic

fertilizer (no N applied to soybean). When the fertilizer rates increased to 200 kg N ha–1

applied to

45

corn (50 kg N ha–1

applied to soybean), the average annual losses from CD and CDS decreased by 31

% and 68 %, respectively, compared to UTD. These results were similar to reported by Tan et al.

(2007) also in Ontario, who found that CDS reduced total nitrate loss by 41 % when compared to

uncontrolled drainage from plots under corn-soybean with corn receiving 168 kg N ha-1

.

Different than nitrate, the effect of controlled drainage on phosphorus loads does not seem to be

very well stablished in the Canadian Prairies, as illustrated by the contrasting results found by

Cordeiro et al. (2014) and Satchithanantham et al. (2014). Results from Quebec also indicate

increased loads from controlled drainage plots equipped with subsurface irrigation. Stampfli and

Madramootoo (2006) found total dissolved phosphorus concentrations from controlled drainage plots

to be equal to or higher to those from uncontrolled drainage which could have been caused by high

phosphorus levels in the irrigation water and a higher P solubility caused by the shallow water table.

Similar results were found by Valero et al. (2007), who report increased P loads in tile drainage from

controlled drainage/sub-irrigation plots compared to uncontrolled drainage plots, even with reductions

of 27 % in total outflow volumes promoted by controlled drainage. Their study also indicates that

dissolved forms of phosphorus comprised more than 96 % of total phosphorus, which is

understandable since the soil is a fine sandy loam and not as susceptible to preferential flow as soils

with high clay content.

Despite the documented environmental and agronomic benefits of controlled tile drainage, adoption

remains low in areas where tile drainage is largely adopted, such as eastern Ontario (Dring et al.,

2015). Surveys and semi-structured interviews with producers and drainage contractors/experts

indicated that although nearly 70 % of producer respondents were aware of this management practice,

lack of extension support, increased farm labor, and cost were ranked as the major disincentives for

adoption of controlled drainage (Dring et al., 2015).

46

7.1.1 Design Currently, there are no specific design procedures for controlled drainage systems in either humid

or arid regions (Ayars et al., 2006). Several design alternatives may satisfy the objectives and

environmental requirements of the controlled drainage system, which depends upon the location,

crop, and soil properties (Fouss et al., 2007). However, some general rules based on economics and

crop requirements have been proposed. For example, as the slope of the field approaches 1 %, the

number and cost of control structures to maintain a water table depth within a desired range for a

specific crop may become economically prohibitive; also, at slopes above 2 %, a uniform water table

depth is difficult to maintain because of lateral seepage when the conductivity exceeds 0.5 m/day

(Fouss et al., 2007).

Controlled drainage may be an option with an existing drainage system, as well as a new system, if

the existing system can be adapted to control the water table without waterlogging a portion of the

field (Ayars et al., 2006). However, for most locations, it is not clear whether the greatest demands on

the system design are to provide good drainage under shallow water table conditions, or to provide

sufficient water supply to meet evapotranspiration (ET) demands during the driest periods (Fouss et

al., 2007). For these reasons, computer model simulations are recommended to conduct a complete

analysis and final design of the water table control system (Fouss et al., 2007; ASABE, 2013). The

DRAINMOD model (Skaggs et al., 2012) specifically calibrated for soils of the field under

consideration (ASABE, 2013) has been suggested for predicting the performance of the controlled

drainage system over 20 to 30 years and selecting the most appropriate design (Fouss et al., 2007).

Alternative control structures in Manitoba depend on the type of outlet. For gravity outlets, stop-log

control structures (Agridrain Corp., Adair, Iowa) have been successfully used (Cordeiro and Sri

Ranjan, 2012; Satchithanantham et al., 2012). However, other structures such as a Water Gate Valve

(Schafer, 2011) or Innotag drainage control systems (Tan et al., 1999) are also available. The former

system is a fully automatic, float-operated head pressure valve installed underground that does not

47

interfere with field operations; the latter is fitted to the collector drains and opens and closes

automatically based on head pressure on a rubber membrane. For sump outlets, a pump control is set

by switches that turn the sump pump on and off (Ayars et al., 2006). Switches can be programed to

operate when the water level in the sump reaches a pre-defined maximum and turn-off when a

minimum level is reached. Variable-speed drives can also be used to match the drainage rate of

discrete events and avoid frequent on-off cycles that can damage pump equipment.

7.1.2 Operation Controlled drainage systems should be operated so that plant roots are supplied with soil air and

can obtain sufficient water (ASABE, 2013). However, switching between uncontrolled and controlled

drainage depends on the weather patterns of the region and the crop being planted. In Manitoba,

controlled drainage for potato production was started in May 31, 2010 and June 14, 2011, whereas

planting dates were April 28 and May 19 of each year, respectively (Satchithanantham et al., 2014).

For corn, controlled drainage started around May 15 in 2010 and 2011 (Cordeiro et al., 2014) in an

attempt to keep the water table constantly at 0.75 m below the soil surface (Cordeiro and Sri Ranjan,

2012). In both potato and corn studies done in Manitoba, the stop-logs were removed during the

winter to avoid damage to the tile drainage system due to freezing (ASABE, 2013). However,

controlled drainage over the winter (i.e., starting in November) has been reported for Minnesota in

some instances (Stacey et al., 2010). This practice is recommended to reduce nutrient losses. In fact,

after the crop harvest and completion of fall farm activities, the system should be switched to

controlled drainage and operated in that mode throughout the winter to reduce drain flow and hence

nutrient losses (ASABE, 2013), although it could potentially reduce the hydrologic benefits (e.g.

reduced surface runoff) during the spring seasons following wet a fall or winter. Controlled drainage

operated in continuous mode, except for brief periods during planting and harvest to prevent delays or

compaction caused by excess moisture, have been reported for Ontario (Tan et al., 1998; Drury et al.,

48

2009; Tan et al., 2015). However, this practice seems more feasible in Ontario where there is runoff

outflow during the winter (Van Esbroeck et al., 2016), which indicates that frost penetration is not

deep enough to cause damage to the tile drainage system.

7.1.3 Maintenance Since controlled drainage systems are essentially a regular tile drainage system with a control

structure, the same maintenance described in section 3 applies to these systems. A specific

recommendation include a more frequent inspection (i.e., twice a year, at planting and harvest) to

determine the presence of sediment or any other obstructions (ASABE, 2013). This is particularly

important for stop-log control structures such as those manufactured by Agridrain, which has an open

top for insertion or removal of the stop-logs. The top opening creates opportunity for sediment, litter,

or even small animals to enter the structure if the rim is close to the soil surface or if it is located in an

area where water ponding occurs. Moreover, all moving parts of the control drainage system should

be protected from freezing (ASABE, 2013). For Manitoba conditions, this last recommendation

usually means removing the stop-logs during the winter months to avoid water being collected in the

structures. Lastly, some control structures can be automated (Johnson et al., 1993), which require a

power supply. In these cases, electrical components (e.g., pumps, multiplexers) should be isolated and

properly installed (ASABE, 2013).

7.2.NUTRIENT MANAGEMENT

As discussed in sub-section 6.2.1, nutrient movement is an important aspect related to the

environmental performance of tile drainage. Plant nutrient management is fundamental to many

current global sustainability issues in agriculture (Mikkelsen, 2011) and is not particularly linked to

tile drainage. In order to address these sustainability issues, the “4R” nutrient stewardship framework

has been proposed (International Fertilizer Industry Association, 2009), which suggests that fertilizer

management can be improved by utilizing scientific principles to achieve the right source, right rate,

right time, and right place during application.

49

The 4R framework acknowledges the three pillars of sustainability, namely, economic, social, and

environmental goals (International Fertilizer Industry Association, 2009). As such, various

stakeholders have input in the selection of nutrient management practices, and their objectives may

not always coincide (Mikkelsen, 2011). The scientific principles governing the 4R framework also

depend on cropping system (e.g., region and crop combination) and socio-economic context under

consideration [e.g., equipment availability, income level; International Fertilizer Industry Association

(2009)]. The following sub-sections will try to address the individual components of the 4R

framework taking into consideration the agronomic and physiographic context of the Northern Great

Plains, with emphasis on the Canadian Prairies and relating the findings to tile drainage when

possible.

7.2.1 Sources (manure types and manures versus synthetic fertilizers) The potential nutrient export through tile drainage is not inherently linked to any specific source of

fertilizer, as discussed in section 6.2.1. Rather, it depends on a number of factors such as agronomic

practices, land use, soil texture, precipitation, runoff potential, and other site-specific features [e.g.,

topography; Beaulac and Reckhow (1982)]. However, the fact that manure is comprised of different

nutrients may represent a higher risk of nutrient flux from farmland where it is applied, depending on

which nutrient is used as the basis for the rate of manure application. Manure applications based on

nitrogen crop needs may result in phosphorus accumulation in the soil because the

nitrogen:phosphorus (N:P) ratio in manure is usually smaller than N:P uptake ratio of crops (Eghball

and Power, 1999). According to Gburek et al. (2000), the N:P ratio of manure ranges from 2:1 to 6:1,

while crop uptake ratio varies from 7:1 to 11:1, which results in excess phosphorus being added to the

soil when manure is applied on a nitrogen basis. In the Canadian Prairies, nitrogen-based application

rates were the accepted practice (The Prairie Provinces’ Committee on Livestock Development and

Manure Management, 2004), although recent manure regulations implemented by the Province of

50

Manitoba require applications to be based on soil phosphorus thresholds (i.e. Manitoba’s Livestock

Manure and Mortalities Management Regulation, last revised in 2008).

The range of N:P ratios in manure arise from different types of livestock and other factors such as

age and diet (Gasser, 1987). For example, the N:P ratios of poultry (broiler), beef cattle (finishing

operation) and swine (grow-finish operation) are 3.3:1, 7.6:1, and 6.2:1, respectively [ratios calculated

from nitrogen and phosphorus amounts excreted in urine and feces combined, based on diet values;

ASABE (2005)]. These N:P ratios contrast to those reported by Sharpley and Moyer (2000) for dairy

manure (4.3:1), poultry manure (2.0:1) and swine slurry (1.9:1), which reflect variability in manure

composition even for the same source. Other factors affecting manure composition at time of

application include type and amount of bedding material used, accumulation time, amount and quality

of water used to flush the facility, location in a storage pit from which the waste is removed, and

length of storage before land application (Sharpley et al., 1994). Therefore, manure composition will

depend on individual livestock operations. Due to this variability, it has been suggested that manure

analysis should be performed before application in order to determine the real nutrient value of

manure (Sharpley et al., 1994).

The release of nutrients from manure also varies with source due to nutrient speciation. Sharpley

and Moyer (2000), analyzing manure composition and phosphorus leaching in runoff from simulated

rainfall, observed that the potential for phosphorus leaching was closely related to water-extractable

inorganic P. Although total phosphorus increased from dairy manure (3,990 mg kg-1

) to poultry

manure (28,650 mg kg-1

) to swine slurry (32,950 mg kg-1

), the total leached phosphorus increased

from dairy manure to swine manure to poultry manure since the latter had the largest water-

extractable inorganic P fraction. Similar results were obtained by Kleinman and Sharpley (2003), who

found that concentrations of dissolved reactive phosphorus in rainfall runoff were larger for poultry

and swine manure than for dairy manure. According to the authors, the differences were driven by

51

differences in water-extractable phosphorus. In Manitoba, research has shown that sodium

bicarbonate plus water extractable P was a better at predicting P loss by leaching and surface runoff

than water-extractable phosphorus alone, highlighting the usefulness of this extraction method for

managing the risk of phosphorus losses from manured land (Kumaragamage et al., 2012).

When the data from the studies discussed above are combined, a pattern emerges where manures

with low N:P ratio and with phosphorus forms more susceptible to runoff represent a higher risk to be

exported. That is, soluble forms of phosphorus applied in excess of crop requirements are prone to be

mobilized by water movement. Although phosphorus is the nutrient usually supplied in excess

through manure applications, nitrogen can also leach from manure applications. Measurements of

nitrate leaching in sandy soils receiving different types of manure in the UK revealed that losses were

consistently lower from cattle farmyard manure than from broiler litter and separated pig or cattle

slurry. Experiments in Saskatchewan showed that application of liquid hog manure at agronomic rates

(i.e., 37,000 L ha–1

yr-1

) presented higher levels of nitrate accumulation in the top 1.5 m of the soil

profile than solid cattle manure at agronomic rates (i.e., 7.6 Mg ha–1

yr-1

), which creates more

opportunity for residual nitrate to be leached (Stumborg et al., 2007). Thus, there is a trend similar to

that of phosphorus where manure containing large amounts of available nitrogen, such as slurries and

poultry manure, increase the risk of nitrate leaching (Chambers et al., 2000).

While different manure types may have different potential for nutrient release, synthetic fertilizer

can represent an even greater risk. Comparison of simulated-rainfall runoff from pasture land (tall

fescue) in a fine-silty soil receiving either poultry litter at 4.5 Mg ha−1

or synthetic N-P-K fertilizer

equivalent to 218 kg N ha−1

and 87 kg P ha−1

showed that average nitrate, total phosphorus, and

phosphate concentrations were lower for poultry litter (1.1 mg NO3-N L-1

; 15.4 mg TP L-1

; 10.4 mg

PO4-P L-1

) than for synthetic fertilizer (2.6 mg NO3-N L-1

; 26.2 mg TP L-1

; 26.1 mg PO4-P L-1

)

(Nichols et al., 1994). Likewise, the exported loads were lower in poultry litter runoff (0.1 kg NO3-N

52

ha-1

; 1.3 kg TP ha-1

; 0.9 kg PO4-P ha-1

) than in synthetic fertilizer runoff (0.2 kg NO3-N ha-1

; 2.0 kg

TP ha-1

; 2.0 kg PO4-P ha-1

). In a similar experiment, Heathwaite et al. (1998) found that total nitrogen

and total phosphorus concentrations in runoff from hillslope plots receiving synthetic fertilizer at 250

kg N ha-1

and 100 kg P ha-1

were higher than those receiving liquid cattle slurry at 50 m3 ha

-1 (138 kg

N ha-1

and 69 kg P ha-1

) and solid cattle farmyard manure including bedding material at 50 t ha-1

(364

kg N ha-1

and 118 kg P ha-1

). While higher nutrient content justifies the higher concentrations

comparing synthetic fertilizer to slurry, it does not explain the higher concentrations when compared

to farmyard manure.

As for manure, different sources of synthetic fertilizers also represent variable risk of nutrient

export. Among the factors affecting runoff and leachate losses, fertilizer solubility has been listed as a

prominent one by Easton and Petrovic (2004). These authors monitored nitrate on runoff and leachate

from Kentucky bluegrass and perennial ryegrass in a coarse loamy soil receiving either natural (i.e.

swine compost, dairy compost, and biosolids) or synthetic [i.e. readily available urea (35-3-5), and

controlled-release urea (24-5-11)] fertilizers and found that synthetic fertilizers were more easily

leached due to higher solubility. They state that their results were similar to those reported by Brown

et al. (1977), who found that sources with high solubility such as ammonium nitrate (NH4NO3)

produced much higher nitrate concentrations in leachate than less soluble sources such as biosolids or

urea formaldehyde.

Phosphorus dynamics is equally complex since its solubility depends on concentrations in the soil.

If a critical level of phosphorus in the soil is exceeded, the risk of transport offsite increases; this

critical threshold is marked by a rapid increase in phosphorus solubility (Dayton et al., 2014), referred

to as change point (Hartz and Johnstone, 2006). Despite this behaviour, application of soluble

fertilizer can have an overriding influence in the characteristics of short-term phosphorus runoff (Hart

et al., 2004). This overriding effect has led to suggestions that soil phosphorus contributes little to

53

phosphorus losses by runoff when a recent surface application of labile phosphorus sources occurred

(Kleinman et al., 2002). As for manure, this overriding effect arises from dramatic, temporary

changes in water extractable phosphorus (Kleinman et al., 2002), also defined as labile phosphorus

(Kumaragamage et al., 2012).

The preceding discussion indicates that different nutrient forms with different solubility and

mobility are source-dependent and play an important role in nutrient transport. While not specifically

studied in a tile drainage context, these findings have direct implications to this practice since

different sources of manure and synthetic fertilizer may present varying degree of risk depending on

composition, handling, and storage. Proper nutrient management should take this aspect into

consideration in order to match nutrient source to crop needs and environmental risk. In this case,

solubility is an important criterion. While one of the major factors influencing crop nutrition (e.g.,

Chien et al., 1990), it is also critical for leaching potential, as discussed above.

7.2.2 Rates of nutrients (relative to crop requirements and removals) Application rate has been suggested as one of the most important factors affecting nitrate losses

from tile-drained land, being more influential than timing, method of application, and mineralization

(Power and Schepers, 1989; Dinnes et al., 2002). It has also been linked to increased phosphorus

concentrations in runoff from manured land (Volf et al., 2007). The large effect of application rates on

nutrient movement though tile drainage has been emphasized in sub-section 6.2.1. Specific studies

assessing fertilizer rates in the prairies also confirm this fact. Early studies in Minnesota showed that

nitrate losses through tile drainage were proportional to fertilizer application rates; cumulative losses

after three years of treatment from plots on continuous corn receiving 20, 112, 224 and 448 kg N ha-1

yr-1

as urea during the spring were, respectively, 19, 25, 59, and 120 kg ha-1

(Gast et al., 1978). These

losses correspond to 32 %, 7 %, 9 % and 9% of the total applied over the experiment, which suggests

that less than 10% of the fertilizer applied at agronomic (i.e. 112 kg N ha-1

·yr) or higher rates was lost

54

through tile drainage. Similar results were found in Iowa for 4-yr corn-soybean rotations fertilized

with anhydrous ammonia in the spring for corn only (Jaynes et al., 2001). The total exported loads

were 29, 35, and 48 kg N ha-1

for plots receiving low (57 and 67 kg N ha-1

yr-1

), medium (114 and 135

kg N ha-1

yr-1

) and high (172 and 202 kg N ha-1

yr-1

) application rates, respectively. The proportion

fertilizer lost through tile drainage was also similar to that of Minnesota, with 23%, 14%, and 13% of

the fertilizer being mobilized for the low medium, and high application rates, respectively.

The relationship between application rates and concentrations in tile effluent is also true for manure

and synthetic phosphorus fertilizers (King et al., 2015b). For example, the flow weighted mean

concentrations in tile effluent from plots receiving 200 tonnes ha-1

yr-1

of dairy manure were 116, 441,

and 96 ppb during the three years of monitoring in New York, while the concentrations were 22, 10,

and 9 ppb for plots receiving 35 tonnes ha-1

yr-1

(Hergert et al., 1981). Higher synthetic fertilizer rates

broadcast onto the soil surface also resulted in higher phosphorus losses in Finland (Turtola and

Jaakkola, 1995).

Over-application of fertilizer can occur due to a number of reasons such as: i) the perception that

the suggested rate is too conservative and that profit can be maximized by applying more fertilizer

than recommended; ii) optimal use of one nutrient is linked to its substitutability with inputs such as

other nutrients and water; iii) failure to appreciate the true opportunity costs of farm inputs, which can

lead to a mistaken conclusion that farmers lose money by applying excess nutrients; and iv)

uncertainty about weather and soil characteristics, which lead both risk-averse and risk-neutral

farmers to over-apply nutrient (Sheriff, 2005). Among these, the perception that recommended rates

are not adequate has been suggested as the major cause of over-application (Rajsic and Weersink,

2008). The complexities involved in nutrient availability certainly support this perception.

Regarding nitrogen fertilizer, inputs higher than the minimum required to achieve maximum crop

growth were traditionally applied because nitrogen fertilizers were relatively cheap compared to the

55

expected economic benefits arising from a maximized crop yield (Lemaire and Gastal, 1997).

However, farmers have had to work with potential rather than maximum yields due to recent

economic and environmental constrains. Thus, nitrogen applications are determined before each

growing season based on climatic conditions (Lemaire and Gastal, 1997). This was the case reported

by Cordeiro et al. (2014) in their investigation of nutrient export through tile drains in Manitoba,

where nitrogen fertilizer application rates were reduced due to a reduction in yield goals as a result of

late planting.

In the case of phosphorus, several pools are available in the soil system, but since the readily

extractable pool supplies the bulk of phosphorus for plant uptake, it is only necessary to accumulate a

certain amount of phosphorus in this pool to achieve optimum crop yield (Kirkby and Johnston,

2008). According to Kirkby and Johnston (2008), a critical value below which there is economic loss

to the farmer and above which there is no increase in crop yield with further P additions is usually

sought. Those authors state that this critical value depends on soil type, crop and farming system;

however, in the absence of detailed information, Olsen P values in the range of 15–25 mg kg-1

are

satisfactory for most crops.

Manitoba Agriculture, Food and Rural Development makes recommendations for both nitrogen and

phosphorus applications for synthetic fertilizers based on soil test (Manitoba Agricultural

Sustainability Initiative, 2007), while recommendations of manure applications are provided by a

joint effort between all three Prairie Provinces (The Prairie Provinces’ Committee on Livestock

Development and Manure Management, 2004). These recommendations aim to match fertilizer

application to crop needs.

Regarding crop requirements, distinction should be made between nutrient uptake, which refers to

the nutrient amount absorbed by the plant, and nutrient removal, which refers to the amount of

nutrient obtained in the harvested portion of the crop (Goswami et al., 1990). Crop nutrient removal

56

depends on several factors such as plant hybrid, plant population, yield potential, fertilizer practice,

and soil conditions (Heckman et al., 2003). However, for most grain crops, both straw and grain are

removed from the field, resulting in crop uptake and removal being about the same (Goswami et al.,

1990). For cropping systems where residue is not removed, residue management is critical to

determine the long-term nutritional needs of the crops (Blair et al., 1990). Even when all these

variables have been taken into consideration, environmental conditions during the growing season

also influence crop uptake. For example, Stumborg et al. (2007), studying nitrogen dynamics in soil

receiving manure in Saskatchewan, observed nitrate accumulation in the top 0.6 m of the soil profile

as a result of low crop removal driven by dry conditions. The authors recommend reduced rates of

applied manure nitrogen when crop N removal potential is diminished by high frequency of drought.

Therefore, application rates may be highly variable even for the same field and crop, especially when

the crop is grown in rotation.

7.2.3 Matching volumes of liquid manure applied to available water holding capacity in soil The discussion in previous sections showed that the water holding capacity of the soil (section 2)

and manure form (i.e., slurry; sub-section 7.2.1) affect the transport of manure nutrient and bacteria in

the soil profile. Infiltration of liquid manure is influenced by the water-holding capacity of the slurry

itself and that of the soil. While high viscosity or dry matter content of the slurry increases its water-

holding capacity, thus decreasing the transport of water from the slurry into the soil, the soil’s water

holding capacity is affected by water content and porosity (Sommer and Jacobsen, 1999). The water

holding capacity of the soil also influences bacteria transport from liquid manure since bacteria

survive longer periods in soils with high water-holding capacity (Unc and Goss, 2004).

These aspects are behind the rationale for developing BMPs for liquid manure application. Since a

high soil water content will decrease the soil’s ability to absorb liquid from the slurry (Sommer and

Jacobsen, 1999), it has been recommended that application rates do not exceed the amount needed to

57

bring the soil to its water holding capacity (Hoorman and Shipitalo, 2006). While this criterion is

relatively simple to meet through an assessment of soil moisture status prior to liquid manure

application, violation of this premise has been listed as the number one reason for animal waste to

enter tile drains in Ohio [Table 4; Hoorman and Shipitalo (2006)].

7.2.4 Placement (including practices to minimize risks of preferential flow of manure, such as injection)

Fertilizer placement depends on its form. Different manure types, namely, liquid, slurry, or solid

(Reidy et al., 2005) will be applied by different methods such as broadcasting (Schmitt et al., 1995),

surface application followed by incorporation, or injection (Rasmussen, 2002). Similarly, synthetic

fertilizers placement methods can be grouped in two major categories: broadcasting or localized

placement along a plant row or in the hill (Lamer, 1957). While a detailed discussion of placement

methods is out of the scope of this review since it affects agronomic (e.g., Schmitt et al., 1999;

Baiyeri and Tenkouano, 2007), economic (e.g., Osei et al., 2003), and environmental performance

(Maguire et al., 2011), a few general guidelines are relevant when fertilizer is applied to tile drained

land.

Regarding manure, surface application is the only practical option for solid forms. This placement

method is itself highly variable, being influenced by many spreading parameters such as material

transport along the rotor, friction, initial position of the manure pieces, and rotational frequency

(Duhovnik et al., 2004). Interaction among these variables can influence manure distribution in the

field and, consequently, the potential for leaching. Moreover, surface application minimizes

interaction between manure and soil particles (Kleinman et al., 2009), which renders nutrient in

manure readily available for mobilization by surface runoff (Vadas et al., 2004). Due to this risk,

incorporation of surface-applied manure by tillage has been recommended. Phosphorus

concentrations in runoff are higher if manure is left unincorporated in the soil surface (Vadas et al.,

2004). Incorporation of manure also reduces phosphorus leaching (Kleinman et al., 2009), which is of

58

concern for tile drainage.

Injection methods have also been advocated if manure is in liquid form. It has been suggested that

injection of manure is a more environmentally friendly practice compared to other applications

methods (Dell et al., 2011), but some authors argue that this practice may also exacerbate nutrient

leaching losses (Kleinman et al., 2009). Regardless of the debate whether incorporation or injection is

more beneficial, disruption of preferential flow is usually presented as the most important factor to

avoid leaching. Incorporation of manure by tillage (Kleinman et al., 2009) or injection at moderate

rates (Ball Coelho et al., 2007) are among the strategies proposed to prevent nutrient leaching from

manure by preferential flow.

Studies on the agronomic effect of synthetic fertilizer placement have also been carried for a long

time (e.g., Olson and Dreier, 1956). A review of placement methods done by Randall and Hoeft

(1988) suggests that, at high soil test levels, crop yield response differences due to placement methods

are rare. However, those authors stress that when soil tests are low, soil moisture and/or precipitation

is limiting, or land is rented, and when fertilizer efficiency and economic return are to be maximized,

localized placement methods giving a zone of high fertilizer concentration within the effective rooting

area should be strongly considered for corn and small grains. Regarding leaching, very few studies

compare the efficacy of different placement methods, and the results are not conclusive. For example,

Tomar and Soper (1981) compared leaching from plots receiving 100 kg N ha-1

that was either

broadcast or banded, in the spring, prior to being planted to barley. The authors found that leaching

losses were negligible for both methods of application. These results could have been caused by a

number of reasons such as the low rate of fertilizer application or weather pattern for that particular

growing season. Similarly, Rees et al. (1997) found little nitrate leaching from fields planted corn or

wheat and receiving N fertilizer by broadcasting, incorporated, or localized placement (corn) or

banding (wheat) although evidence of leaching was observed up to 0.9 m on several plots. Again, the

59

reason for reduced leaching could have been due to the deep layer considered in the study (i.e., up to

2.0 m), although application rates were not necessarily low (i.e., 210 kg N ha-1

for corn and up to 150

kg N ha-1

for wheat).

7.2.5 Timing relative to rainfall and snowmelt events Nitrogen and phosphorus leaching from manures or synthetic fertilizer is sensitive to timing of

application, which has implications for the quality of water in tile drainage outflow (van Es et al.,

2004; van Es et al., 2006). Research in the UK on nitrate leaching suggest that manure containing

large amounts of available N, such as slurries and poultry manure, should not be applied to free-

draining soils; in that region, these conditions occur between autumn to early winter (Chambers et al.,

2000). However, the authors argue that increases in leaching is much less if manure has only a small

amount of readily available N, such as straw-based cattle farmyard manure. These results are

consistent with those found in New York state, where the risk of leaching of nitrogen and phosphorus

into tile drains was increased if manure were applied in the early fall (van Es et al., 2004; van Es et

al., 2006). In the case of nitrogen, precipitation pattern greatly affected leaching concentrations, while

for phosphorus, most of the losses occurred during and immediately following periods of high

precipitation, presumably through preferential flow. These results suggest that manure application

should not match periods prone to tile drainage outflow.

In the case of Manitoba, drainage outflow seems to be concentrated around a well-defined drainage

season in the spring (see section 6.2.1); thus manure applications should be avoided during that time.

However, if tile drains are running during the fall, manure applications should also be restricted.

While timing of application to weather conditions is easy in theory, it is hard to achieve in practice.

According to Hart et al. (2004), the concept of better timing of P applications in relation to climate

conditions, etc., is fine in principle and should be encouraged as much as possible, but since the

weather in many parts of the world is unpredictable, it cannot be relied on as a major method for

60

managing potential P losses.

Timing of synthetic fertilizer application has also been shown to affect leaching. For example, a

two-year experiment on nitrate leaching from potatoes planted in loamy sand soils showed that

reducing N application at planting and increasing amount at emergence and hilling stages for irrigated

Russet Burbank potato can reduce the potential for nitrate leaching, while increasing N uptake and

improving marketable yield (Errebhi et al., 1998). Similarly, side-dressing N for corn early in the

growing season has proven to be a sound strategy for optimizing yields and minimizing nitrate losses

in tile drains, whereas delaying side-dressing until mid-season (just before reproductive growth) has

been shown to negate much of the yield and nitrate leaching loss benefits (Jaynes, 2013). As for intra-

seasonal changes in application timing, inter-seasonal changes in application of synthetic fertilizer

have also proved to be effective. For example, annual losses of nitrate in the tile drainage water from

experiments in Minnesota averaged 36 % higher with fall application compared with spring

application; the 4-yr average annual concentrations and exported loads were 20 mg L-1

and 264 kg ha-

1 when fertilizer was applied in the fall, while the values if applied in the spring were 16 mg L

-1 and

190 kg ha-1

, respectively (Randall and Mulla, 2001). These results suggest that fertilizer should be

applied when crops do need and use it (i.e., spring), with enhanced efficiency if applications are split

in the spring. Economic assessments in Minnesota suggest that switching from fall to spring

application is feasible if timing is combined with a reduced application rate. Switching from fall to

spring pre-plant N application along with a rate reduction would achieve an 11 % reduction in

exported nitrogen loads; the timing switch would cost money but not enough to offset the fertilizer

cost savings (Lazarus et al., 2014).

7.3.AGRONOMIC PRACTICES

Nutrient loss from agricultural land to water resources depends on a number of factors, including

agronomic ones such as tillage, and crop selection (Dinnes et al., 2002). The influence of tillage on

61

nutrient loss has been investigated in Manitoba (Tiessen et al., 2010). Likewise, the effect of crop

rotations on nutrient dynamics and leaching has been established for some rotations in the Canadian

Prairies (Campbell et al., 1990). Nonetheless, the impact of these practices in tile-drained land is not

available for this region since tile drainage is a relatively new practice. This section discusses tillage,

crop rotation, and cover crops as BMPs to reduce nutrient export through tile drains based on the

research conducted elsewhere.

7.3.1 Tillage Soil tillage is a fundamental operation in agriculture because of its influence on soil properties,

environment, and crop production (Boydaş and Turgut, 2007). The primary purposes of tillage are to

manage soil structure, to manage weeds, and to manage crop residues (Raney and Zingg, 1957). Since

tillage operations bear an effect on soil structure, they modify the bulk density (i.e., porosity) and pore

size distribution of the soil, thus affecting its hydraulic properties (Klute, 1982). As a result, nutrient

movement is also affected.

The influence of tillage on tile drainage outflow and nitrate loss was investigated in Ontario by

Patni et al. (1996). Those authors compared no-till to conventional tillage in a loam soil under corn in

a 3-yr experiment and found that flow was significantly higher under no-till than conventional tillage.

The nitrate loss in tile effluent over the 3-yr study was not significantly different for the two

treatments, although it was significantly higher under no-till in one crop-year. The larger drain

outflow was the main reason influencing nitrate loss, since flow-weighted mean concentrations tended

to be higher in the conventional tillage treatment. Among the reasons for greater flow under no-till,

the authors suggest preferential flow and larger water holding capacity. Soil macropores remain

undisturbed under no-till, reducing the resistance to downward movement of water when compared to

conventional tillage where pores near the surface are broken down, while the higher water retention

capacity in the soil under no-till results in greater antecedent moisture content (Patni et al., 1996).

62

Notably, lowering the moisture content of the soil in the fall is the rationale behind adopting tile

drainage as a BMP to mitigate floods during the spring in Manitoba. Thus, tillage seems to enhance

the potential for water storage in the soil profile while lowering the potential for nutrient export in tile

drained land.

The results found in Ontario are similar to those found in a 3-yr study in Iowa by Bjorneberg et al.

(1996a), who compared drainage outflow and nitrate loss from continuous corn and corn-soybean

rotation plots tilled by four different systems: chisel plow, moldboard plow, ridge till, and no-till.

According to the authors, no-till plots had significantly higher subsurface drain flow than moldboard

plow plots under continuous corn, although this variable was not statistically different for the other

tillage treatments and cropping system. Nitrate concentrations in drain effluent from moldboard and

chisel plow systems were significantly greater than concentrations from no-till and ridge till systems

for all crop rotations. The authors suggest that the higher nitrate concentration for some tillage

treatments compared to no-till was a result of water moving through preferential flow in the latter,

which leached less nitrate from the soil profile. This is consistent with the estimation of preferential

flow in corn or soybean rotation in Iowa subject to chisel plow or no-till, where the annual

preferential flow estimations were usually higher for no-till than chisel plow for both cropping

systems (Bjorneberg et al., 1996b). The reason for a larger proportion of preferential flow in no-till

systems is because tillage disrupts preferential flow paths near the surface (Ehlers, 1975). As a result,

downward movement of water has to take place through the soil matrix, which reduces total drain

outflow (Bjorneberg et al., 1996a).

Tillage has also been shown to prevent movement of phosphorus through preferential flow to tile

drains. For example, an experiment on a clay-loam soil in Minnesota showed that ridge tillage had 26

times higher total P losses in tile outflow than the moldboard treatment, mainly due to preferential

flow of surface runoff that carried soluble P from surface-applied manure and crop residues (Zhao et

63

al., 2001). However, the effects of tillage are short-lived. According to King et al. (2015b),

desiccation of soils prone to cracking, freezing and thawing cycles, and earthworm activity may

recreate macropores in the topsoil, which would offset the tillage effect. Those authors suggest that

the impact of tillage on preferential flow paths is minimized on an annual scale. Thus, tillage

practiced on a frequent basis in order to be effective to prevent preferential flow towards tile drains.

Due to the frequently short duration of tillage effects on disrupting preferential flow paths,

performing tillage operations immediately prior to manure application has been suggested as a way to

enhance its effects (Hoorman and Shipitalo, 2006). For example, Kay et al. (2004) found that tillage

prior to slurry application may limit the transport of veterinary pharmaceuticals through tile drains

and suggest that this practice could be used as a management tool by farmers. This pre-pass concept

has been incorporated into the design of farming implements in order to combine both operations into

one pass, such as a slurry applicator sweep that creates a horizontal slice or furrow beneath the soil

surface where the slurry is injected (Dietrich Sr, 2005). The horizontal furrow is created by a sweep

shank placed before the injector and that causes minimum disturbance of the soil surface. By creating

this furrow, this implement can potentially decrease nutrient and pathogen movement towards tile

drains through disruption of preferential flow pathways. An assessment of this this type of one-pass

implement in Ontario displayed low risk of nitrate movement to groundwater in silt loam soils with

no artificial drainage (Coelho et al., 2009), but its impacts on tile-drained fields is yet to be assessed

by research in tile-drained fields.

7.3.2 Crop rotation Crop rotation increases yield and profit and allows for sustained production, which are likely

brought about due to improvements in soil physical properties and soil organic matter, although the

process behind them are not well understood (Bullock, 1992). Crop rotation experiments in the

Canadian Prairies have been conducted to assess its benefits on crop production and quality, control of

64

weeds, disease, and insects, as well as influence on soil moisture, nutrient dynamics, and soil quality

(Campbell et al., 1990). However, studies on the effect of tile drainage on the nutrient dynamics of

crop rotations are scarce in the literature. According to the few studies available, losses within a crop

rotation vary depending on the crop-year. For example, in the corn-oats-alfalfa-alfalfa rotation

reported by Bolton et al. (1970), losses nitrogen and phosphorus from fertilized plots were larger in

the corn year, followed by the second alfalfa year. Also, Logan et al. (1994) examined nitrate losses in

a corn-soybean rotation and found that nitrate concentrations in soybean years were equal to or larger

than those in corn years. However, the authors highlighted that a significant fraction of nitrate in tile

flow is due to N carried over from the previous crop, which was supported by the high concentrations

in the spring prior to application of N fertilizer. The annual exported loads, however, did not show any

consistent trend.

The results above agree with those obtained from long-term simulations using the Root Zone Water

Quality Model (RZWQM) in Iowa, where although N losses were higher under a corn-corn rotation

than under a corn-soybean rotation, N losses could be comparable to those of the corn-soybean

rotation under certain circumstances (Ma et al., 2007). The contrasting or inconclusive results from

the studies above, when assessed in light of annual variation in nutrient export within a cropping

system, suggest that the potential for nutrient export may be hard to define quantitatively for a single

crop within a rotation, although the studies discussed above do assess this potential qualitatively.

Thus, it might be more useful to assess the potential for nutrient export considering cropping systems

rather than individual crops within a crop rotation due to the annual influence of weather (e.g.,

precipitation pattern), fertilizer source and application rate, agronomic practices (e.g., tillage), etc.

Such discussion has been provided in section 6.2.1.2, where nutrient transport through tile drains is

enhanced in perennial forages when compared to annual crops.

65

7.3.3 Cover crops Integrating cover crops into cropping systems promotes pest-suppression, weed management, soil

and water quality, nutrient cycling efficiency, and cash crop productivity (Sarrantonio and Gallandt,

2003; Snapp et al., 2005). The benefits of cover crops on tile drainage water quality have also been

assessed in tile-drained fields in Canada. A 6-yr investigation of the effectiveness of winter wheat

cover crop in reducing nitrate losses from corn-soybean rotation was conducted in Ontario by Drury

et al. (2014). The results showed that the cover crop reduced the 5-yr flow-weighted mean nitrate

concentration in tile drainage water by 21 % and cumulative nitrate loss by 14 % relative to no cover

crop, despite an increase in tile drainage volume of 7 %. This increase in tile outflow was driven by

an increase in near-surface (20 cm) saturated hydraulic conductivity, which also resulted in reduced

runoff.

Similar results were also obtained in Minnesota, where winter rye was tested as a cover crop

following corn in a corn-soybean cropping system (Strock et al., 2004a). The study showed that, over

three years, tile-drainage discharge was reduced 11 % and nitrate loss was reduced 13 % for the

cropping system having the winter rye cover crop, although the magnitude of the effect varied

considerably with annual precipitation. The authors stress the potential of winter rye as a management

tool to reduce nitrate losses in tile drainage outflow in the north-central US, despite the challenges to

establishment and spring growth due to low rainfall and cool temperatures. The influence of

precipitation on effectiveness of cover crops was also observed in Iowa, where Kaspar et al. (2012)

found that rye winter cover crop planted after harvest significantly reduced annual flow-weighted

nitrate concentrations by 48 %, while an oat fall cover crop broadcast seeded into living corn and

soybean crops before harvest in late August or early September and killed by cold temperatures in late

November or early December reduced concentrations by 26 %. Despite the reductions in

concentrations, both cover crops did not significantly reduce nitrate loads, likely due to high volumes

and variability of cumulative drainage.

66

Despite the results above showing decreased leaching of nitrates into surface water and

groundwater, Riddle and Bergstrőm (2013) argue that cover crops have the potential to increase

phosphorus loadings to natural waters due to plant destruction during freezing–thawing events. Using

indoor lysimeter experiments where plant material from four different species (i.e., perennial ryegrass,

honey herb, chicory and oilseed radish) was applied to clay and sand topsoil and subject to simulated

rainfall and freezing, these authors found larger P-leaching fractions from clay when compared to

sand. Phosphorus leaching from clay soil decreased in the following sequence: chicory (2.6 kg ha-1

) >

ryegrass (2.3 kg ha-1

) > oilseed radish (2.2 kg ha-1

) > honey herb (1.3 kg ha-1

). These results contrast

to those found by Bechmann et al. (2005), who found that phosphorus leaching from 0.3-m intact soil

columns was not increased by freezing of the catch crop, although repeated freezing and thawing

significantly increased water-extractable phosphorus from catch crop biomass and resulted in

significantly elevated concentrations of dissolved phosphorus in runoff.

Regarding the potential for different legumes to be used in the Canadian Prairies as cover crops, a

field study was conducted in Alberta to determine the merits of establishing alfalfa, red clover or

Austrian winter pea as cover crops in the fall or spring with winter wheat (Blackshaw et al., 2010).

The results showed that spring-planted legumes emerged well within the winter wheat crop, but their

growth was limited under the prevailing semi-arid conditions. Fall-planted red clover had low plant

densities following winter, while fall-planted winter pea reduced winter wheat yield compared to the

control. In contrast, fall-planted alfalfa exhibited good winter hardiness, provided some weed

suppression without reducing winter wheat yield, caused only a slight reduction in soil water content,

and contributed some extra soil N at the time of seeding the following spring crop. These

characteristics are likely responsible for the large adoption of alfalfa, which remains the most

common and widely grown forage in western Canada with over 3 million ha (Acharya, 2014). Results

obtained by Blackshaw et al. (2010) suggest that alfalfa is a good legume to be tested as a cover crop

67

for reducing nutrient content in tile outflow in Manitoba. However, recent research in this Province

has shown that phosphorus losses in surface runoff from perennial forage fields can be greater that

those from annual crops (Liu et al., 2014).

7.4.END-OF-PIPE MANAGEMENT

Treatment of industrial and municipal wastes, as well as of liquid or waterborne wastes from

concentrated animal feeding operations has successfully been achieved through end-of-pipe

technologies (John and Donald, 1978). Some of these technologies have also been proposed for

treating tile drainage outflow, such as water retention (e.g., wetlands), biofilters, and vegetative filter

strips or buffers. These approaches are discussed in the following sub-sections.

7.4.1 Natural (wetlands) and constructed reservoirs Wetlands perform several environmental functions such as nutrient retention, sediment control,

carbon storage, wildlife habitat, and flood mitigation (Richardson et al., 2011; Molnar and

Kubiszewski, 2012). The water-quality improvements promoted by wetlands such as reduction of

nitrogen and phosphorus are driven by treatment mechanisms such as sedimentation, filtration,

chemical precipitation and adsorption, microbial interaction, and uptake by vegetation (Watson et al.,

1989). Most investigations on the use of wetlands for the treatment of agricultural nonpoint-source

pollution have focused on surface runoff, with less information on wetland performance in treating

subsurface drainage (King et al., 2015b). This is the case in Canada, where evaluation of the wetland

effectiveness to reduce nutrient loads is limited. A study carried out in Ontario intercepted both

surface runoff and tile drainage outflow in a wetland-reservoir and routed the intercepted water back

to the field for sub-irrigation (Tan et al., 2007). Although the combination of both surface runoff and

tile drainage outflow did not allow for a specific assessment of the efficiency of the wetland to reduce

nutrient loads in tile outflow alone, the wetland-reservoir water had substantially lower nitrate and

phosphorus concentrations relative to the tile drainage water. According to the authors, this result was

probably due to nitrate and phosphorus uptake by aquatic plants and algae. While this is a possibility,

68

the authors did not comment on other processes that could have reduced nutrient concentration in the

wetland water, such as dilution by surface runoff or precipitation. Since most of the phosphorus loss

occurred primarily through tile drainage rather than by surface runoff (Tan et al., 2007), dilution by

runoff was likely, but this could not be assessed since the study did not report surface runoff volumes.

More conclusive studies have been carried out in the US. For example, water and nutrient budgets

were determined for three years of monitoring for three constructed wetlands receiving tile drainage

outflow in Illinois (Kovacic et al., 2000). The surface area, volume, tile drainage area to wetland

surface area ratio, and average depth for the three wetlands were, respectively, 6, 3, and 0.8 ha; 5400,

1200, and 5200 m3; 25, 17, 32 m

2 m

-2; and 0.9, 0.4 and 0.7 m. The wetlands intercepted tile outflow

from fields planted to corn and soybean. The results indicated that 95 % of the annual wetland

hydraulic loading occurred during winter and spring, although precipitation during these seasons

accounts for only 52 % of total precipitation. Inflow was proportional to the land area drained and 75

% of the hydraulic load of the wetlands was filled vary rapidly due to the flashy nature of tile drainage

outflow. During the three years of monitoring, the nitrate inflow and outflow to and from the wetlands

were 4,445 and 2,749 kg, respectively, which corresponded to a reduction of 1,696 kg or 38 % of the

total. A somewhat similar reduction (i.e., 30 %) was estimated for restored wetlands in the highest

nitrate contributing areas of the tile-drained regions of the upper Mississippi River and Ohio River

basins (Crumpton et al., 2006).

Regarding phosphorus, wetlands removal efficiencies reported by Kovacic et al. (2000) were

smaller than that of nitrate. A total of 63.3 kg of dissolved P entered the wetlands and 49.1 kg were

exported, corresponding to a total removal efficiency of 22 %. The removal of total phosphorus was

much smaller than that of dissolved phosphorus, where only 1.5 kg or 2 % of the overall 63.6 kg total

P load were removed. The authors argue that removal of total phosphorus varied within the wetlands

depending on flow and retention time, with a wetland-year case where exported loads exceeded the

69

inputs for both dissolved and total phosphorus.

The results presented above illustrate that wetlands are efficient to remove nitrate loads and their

efficiency for phosphorus depends on the hydraulics of the wetland in a particular year, which

illustrate the influence of hydraulic variables (e.g., number of tile flow days, time to achieve wetland

hydraulic loading, and retention time) on nutrient dynamics within the wetland (Kovacic et al., 2000).

Performance comparison across sites in Maryland, Illinois, and Iowa indicates that among the

hydraulic variables influencing performance, large wetlands relative to the contributing drainage area

is the most important criterion to effectively improve water quality (Woltemade, 2000). This aspect

has implications to crop yields since more agricultural land would have to come out of production to

enhance wetland nutrient removal. Another consideration is the longevity of this function; as wetlands

accumulate sediments and nutrients their filtering ability may be deminished.

7.4.2 Biofilters One potential practice to reduce nitrate and phosphate transfer from tile drained farmlands is the

use of fixed-bed, in-field subsurface bioreactors (or biofilters). Bioreactors are trenches filled with

carbon material (usually wood chips) that serve as a medium for denitrifying bacteria to grow, thus

reducing the amount of nitrate that enters water bodies from tile drains (Cooke and Bell, 2014).

Denitrifying drainage bioreactors can provide cost-effective treatment of tile water at the field scale

while requiring minimal system maintenance; moreover, they do not cause yield reductions or take

land out of production (Christianson et al., 2009). Biofilters have been recently proposed to mitigate

nitrate export from tile drained land in Quebec and Ontario (Rasouli et al., 2014).

According to Christianson et al. (2012b), the first published work on biofilters using a solid carbon

source was conducted in Ontario by Blowes et al. (1994). In that work, three 200-L fixed-bed

bioreactors containing porous-medium material of coarse sand and organic carbon (tree bark, wood

chips and leaf compost) were partially buried in the stream bank at a location 100 m downstream from

70

an active tile drainage system. The results indicated that the bioreactors were able treat tile drainage

outflow containing nitrate concentrations of 3-6 mg L-1

and reduce the concentrations to < 0.02 mg L-

1 at a rate of 10-60 L day

-1 over a 1-yr period. In a more recent study, van Driel et al. (2006) used a

lateral flow design bioreactor to reduce nitrate concentrations in tile outflow from a cornfield in

Ontario. The bioreactor was able to reduce the average nitrate concentrations from 11.8 mg L-1

to 3.9

mg L-1

at an average flow rate of 7.7 L min-1

over 26 months, resulting in an average removal rate of

2.5 g NO3-N m2 day

-1. The reduction in average nitrate concentrations corresponds to a 67 %

decrease, which is in agreement with other studies in the US. For example, a field-scale bioreactor in

Illinois indicated reductions around 47 % in nitrate concentrations 15 hours after a nitrate pulse was

routed through the filter with a retention time of 4.4 h (Chun et al., 2010). Similarly, nitrate reductions

between 30 % and 70 % were reported for retention times between 4 and 8 h in Iowa (Christianson et

al., 2011). It has been highlighted that the bioreactors removal rates were one order of magnitude

higher than those of other passive treatment systems such as constructed wetlands, even at low

operating temperatures between 7 °C and 9 °C (van Driel et al., 2006).

Temperature is actually an important factor influencing biofilters’ performance. Analysis of

removal performance of four bioreactors in Iowa found that temperature was the most important

factor affecting percent bioreactor nitrate load reduction among all the factors analyzed, which

included retention time, influent nitrate concentration, temperature, flow rate, age, length-to-width

ratio, and cross-sectional shape (Christianson et al., 2012a). Since nitrogen loads in tile drainage are

greatest in the spring, when temperatures are just above freezing, optimization of bioreactor treatment

becomes a challenge (Christianson et al., 2012b). Bioreactors are recommended to be managed at

longer retention times if temperature is low, which can be achieved through the use of control

structures (Christianson et al., 2012b). However, this means another challenge in Manitoba since the

spring, with its cold temperatures, usually corresponds to the highest tile flow rates (Cordeiro et al.,

71

2014; Satchithanantham et al., 2014; Satchithanantham and Sri Ranjan, 2015). In fact, one of the

greatest challenges associated with the treatment of agricultural runoff is the handling of peak flows

associated with spring snowmelt and major rainfall events (van Driel et al., 2006), which results in

lack of economic feasibility if bioreactors are designed to handle peak flow rates (Partheeban et al.,

2014). Thus, it has been suggested designing bioreactors to treat approximately 20 % of the peak

flow, which corresponds to approximately 70 % of the total flow volume during the year

(Christianson et al., 2009). Regarding the operating performance of bioreactors, the estimates are

often on the order of several decades, with empirical data showing at least 10 years (Christianson et

al., 2012b). Findings from a bioreactor long-term monitoring in Ontario indicate that nitrate removal

rates, measured subsequently in the sixth and seventh years of operation, varied with temperature in

the range of 2 to 16 mg N L-1

·day, but remained similar to rates measured in the second year of

operation (Robertson et al., 2009).

Since bioreactors are mostly designed for removal of nitrate by denitrification (Christianson et al.,

2009), results dealing with phosphorus removal are much more scarce. In order to address phosphate

removal, Salo et al. (2015) tried a woodchip bioreactor system using iron media as post-treatment to

simultaneously remove nitrate and phosphate from subsurface drainage. The results of the column

experiments show that woodchips completely remove the 20 mg L-1

of nitrate in the influent at design

conditions. A 55 % reduction of phosphate from the initial 1 mg L-1

was observed through the

woodchip reactor after 100 days of operation, while the steel media completely removed any

phosphate from the woodchip reactor effluent. Removal rates similar to those observed for the

woodchip bioreactor were found by Wang et al. (2010), who used two shortcut nitrification and

denitrification (SND) batch bioreactors in sequence to treat swine wastewater and found that mean

removal of dissolved phosphorus around 52 %. While these technologies sound promising for

phosphorus removal, their effectiveness to treat tile drainage outflow at field scale is yet to be

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investigated.

7.4.3 Vegetative filter strips/buffers Buffer strips can perform a multitude of functions such as promote channel stability, filter sediment

and nutrients, purify water from bacteria and pathogens, and support terrestrial and stream habitat

(Barling and Moore, 1994). A theoretical assessment of this BMP in the context of Ontario suggests

that it may not be very effective to treat tile outflow (Hickey and Doran, 2004). According to those

authors, nutrient removal by buffer strips requires contact between runoff water and soil containing

micro-organisms (denitrification) or the roots of plants (plant uptake); however, tile drains concentrate

agricultural runoff, thus reducing the efficacy of nutrient removal by overloading the assimilation

(e.g., plant uptake) and transformation (e.g., denitrification) processes (Hickey and Doran, 2004).

According to the authors, buffer zones are most effective in preventing the deterioration of water

quality in areas where the natural drainage patterns are intact.

This opinion was also corroborated by Osborne and Kovacic (1993) in their study on the effect

riparian vegetated buffer strips in water-quality restoration and stream management. The authors

measured nitrate, dissolved phosphorus, and total phosphorus in the shallow and deep groundwater of

riparian buffers having forest, grass, and tile-drained crop vegetation. The results showed that the total

and dissolved phosphorus concentrations were lower in the cropped buffer due to the rapid transport

of water in tile drains running through the treatment, which flushed phosphorus more rapidly to the

stream channel. In order to address the bypass effect of tile drains through the rootzone, Jaynes and

Isenhart (2014) re-routed a fraction of field tile drainage as subsurface flow through a riparian buffer

for increasing nitrate removal. The authors attributed the strong decrease in nitrate concentration

within the shallow groundwater across the buffer to plant uptake, microbial immobilization, or

denitrification in the buffer.

Another approach proposed in the literature to treat tile drainage outflow is to irrigate riparian

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buffers with drainage water. Hoffmann et al. (2009) reviewed a few studies assessing phosphorus

retention by riparian wetlands and floodplains receiving agricultural tile drainage water and observed

that inter-annual variations may be high, with retention of dissolved reactive phosphorus in one year

and no retention or net loss in the following year. Maybe this is because, in general, sedimentation of

particulate phosphorus is the main trapping mechanism for this nutrient in buffers; as a result,

retention of total phosphorus generally exceeds that of dissolved reactive phosphorus (Hoffmann et

al., 2009). This has implications for tile outflow treatment efficacy, depending on the phosphorus

forms in the effluent. If tile drainage does not carry much sediment, buffers strips will likely be less

effective. However, if conditions such as soil cracking exist to promote export of sediment-rich tile

outflow, buffer strips may be a feasible BMP.

The efficacy of vegetative buffers to treat runoff in cold climates has also been debated due to

higher loss of phosphorus. High Olsen-P in the soil surface and decaying of plant residues have been

suggested to drive the high loss of reactive phosphorus from buffers strips in Finland (Uusi-Kämppä,

2005). In Manitoba, studies in established vegetative filter strips indicated that these features had no

impact on phosphorus concentration in 32 % of the cases sampled or increased phosphorus

concentration in runoff in 18 % of the cases sampled (Sheppard et al., 2006). High available

phosphorus in the soil of the vegetative filter strips along the water flow paths was likely driving these

higher concentrations, indicating that the trend of runoff to flow through rather small portions of

vegetative filter strips in flat-land regions like Manitoba may result in an insufficient capacity of these

features to retain the runoff phosphorus in the longer term. The poor performance of vegetative filter

strips is believed to be largely due to the fact that the majority of runoff and nutrient loading to

surface waters in this region occurs during spring snowmelt when the ground is frozen and nearly

impermeable, and when the vegetation is unable to take up water and nutrients and may actually

release nutrients (Lobb, 2012). Where vegetative filter strips are effective in taking up nutrients, their

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ability to do so will diminish over time, as they become saturated. Removal of nutrient-rich sediment

and vegetation may increase the longer term effectiveness of the management practice.

8. CONSIDERATIONS FOR ADAPTING TILE DRAINAGE BMPS FOR MANITOBA

CONDITIONS Nutrient loss from nonpoint sources to water resources is not caused by any single factor; rather, it

is caused by a combination of factors, including tillage, drainage, crop selection, soil organic matter

levels, hydrology, and temperature and precipitation patterns (Dinnes et al., 2002). Consequently, a

single BMP will not likely address the loss of nutrients from agricultural land. This synergism among

different BMPs has been observed in Ontario in a tile-drainage context (Van Esbroeck et al., 2016). It

is important, though, to consider the conditions under which BMPs are most effective in order to

achieve the expected reduction in nutrient export. For that, specific BMPs should be fit to the

physiographic and hydrological conditions, as well as the agronomic practices of the region. In the

case of Manitoba, these aspects include its cold climate, soil types, crops grown, and farming systems.

These aspects are briefly discussed below based on the topics and issues previously outlined in the

review.

8.1.PHYSIOGRAPHIC CONDITIONS (CLIMATE, RELIEF, SURFICIAL GEOLOGY, SOILS, ETC.) AND THE

INFLUENCE OF SNOWMELT, FROZEN SOILS, CRACKING SOILS, SHALLOW WATER TABLES, ETC.

One of the major parameters influencing nutrient loss and BMP effectiveness is precipitation, since

nutrient movement from agricultural land is inherently linked to water movement. Therefore, volume

and phase of precipitation are quite important. In Manitoba, the mean annual precipitation is around

550 mm, out of which around 25 % takes place as snow (La Salle Redboine Conservation District,

2007; Tiessen et al., 2010). While growing-season rainfall events larger than 10 mm account for 52 %

of the total precipitation in the Canadian Prairies (Akinremi et al., 1999), most of the total annual

runoff takes place during the spring snowmelt (Shook and Pomeroy, 2010). The implication of this

aspect for BMP effectiveness depends on the type of BMP. For hydrological BMPs that aim to reduce

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or treat the volume of water leaving agricultural fields by surface runoff or tile drainage, it means that

the BMP should consider timing, volume, and effluent characteristics. These aspects have been

highlighted in section 7.4.2, where the performance of biofilters is influenced by the volume and

temperature of the water being routed through the filter during the spring. Another example is

controlled drainage, whose operation should be timed to prevent excessive tile drainage outflow

during spring and fall. For nutrient-management BMPs, the effect of precipitation and runoff vary. For

example, manure application timing should not coincide with precipitation events in tile-drained land,

while the effectiveness of cover crops is indirectly linked to precipitation. For the latter BMP, the

effect of other variables such as temperature becomes more important since cover crop establishment

depends on this variable.

In fact, the interaction between precipitation and temperature becomes very important in cold

regions such as Manitoba not only because it influences runoff timing but also due to the effect it has

on soils. As discussed in sections 4.2 and 6.2.1, frozen soils encourage surface runoff and can induce

soil cracking, which is exacerbated by the type and content of clay in Manitoba soils. Temperature is

also linked to enhanced nutrient release from crop residue in Canadian Prairies as a result of rupture

of plant cells caused by freeze-thaw cycles (Liu et al., 2014). Temperature also influences nutrient-

management BMPs such as the timing component of the 4R framework, since manure should not be

spread on frozen soils. This restricts the time window within which this BMP can be implemented in

Manitoba.

The likelihood of soils in Manitoba to develop cracks or channels can also influence BMP

applications in the Province. It has been discussed that manure injection is environmentally friendly

practice. Although injection can disrupt preferential flow, nutrients from injections at high application

rates can still reach tile drains through preferential flow promoted by soil cracking. Furthermore, rates

of manure injection that exceed the soil’s water holding capacity can trigger tile drain outflow. Also,

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the efficacy of BMPs such as tillage can be offset due to soil cracking, which requires frequent re-

working of the fields (section 7.3.1).

The efficiency of BMPs can also be affected by another physiographic feature of Manitoba,

namely, its flat topography, which is unfavorable for application of naturally-occurring water retention

areas such as wetlands for nutrient treatment. The flat topography results in large surface areas if a

larger treatment volume is required for such structures. Application of water retention BMPs thus

require earth movement for increased storage capacity, which increases costs and hazards (e.g.

increased depths) for operation of farm machinery and personnel. While not favorable for natural

water retention structures, the flat topography does favour ponding (thus the justification for adoption

of tile drainage in the province). Surface ponding, associated with clay-rich lacustrine deposits or

glacial tills as the underlying geology (Manitoba Mineral Resources, 2016) that restrict downward

movement of water, promote the recharge of the shallow water tables usually observed in Manitoba.

As discussed in section 7.1, shallow water tables can enhance phosphorus solubility, resulting in

higher concentrations in tile drainage water. This aspect has implications for BMPs such as controlled

drainage, which can itself enhance phosphorus solubility and compound the problem.

All these facets of Manitoba’s physiography are well-known. However, the knowledge is not

complete about some of them, such as implication of surface geology to groundwater recharge

(Harris, 2015b). The interaction between Manitoba’s physiographic features, tile drainage, and BMPs

for treating its water outflow still needs further investigations in order to minimize potential impacts

and enhance efficiency. Investigations of interactions between these features and agronomic practices

typical of the Province are also necessary within a tile-drainage context.

8.2.AGRONOMIC PRACTICES

The literature discussed in the previous sections suggests that the potential for nutrient movement

through tile drainage systems is similar, overall, between manure and synthetic fertilizer, and that

77

nutrient management is a critical factor to avoid this movement. Consequently, agronomic practices

pertaining to nutrient management are likely those that affect nutrient loss the most. Fertilizer

application in the Prairies is often done in the fall after harvest due to the fields often being drier than

at the spring seeding time (Chen et al., 2005). The short growing season in Manitoba (Friesen and

Wall, 1986; Thompson and Clark, 1989) may also contribute to this practice due to the narrow time

window for other spring operations. Fall-applications of fertilizer, while justified from a farm

management and economics (i.e., fertilizers are usually less expensive in the fall) standpoint,

enhances nutrient export. As discussed in section 7.2.5, spring applications of fertilizer are suggested

as a BMP to reduce nutrient loss from farmland in general, which applies to tile-drained lands as well.

Similar to fertilizer application, where cropland is tilled, it is often practiced in the fall (Chen et al.,

2005). This practice has been suggested as a BMP to disrupt preferential flow paths towards tile

drains but its efficacy is time-sensitive since freeze-thaw cycles can re-establish these flow paths, thus

negating the tillage effect (section 7.3.1). In order to address that, spring tillage should be encouraged

if this practice is to be used as a BMP to avoid nutrient transport through tile drains, although spring

tillage has agronomic implications (e.g., loss of seedbed quality).

Investigations of nutrient export from different cropping systems seem to suggest that perennial

cropping systems tend to export more nutrients through tile drainage than annual crops (section 6.2.1).

However, the use of perennial forages will likely be limited in tile-drained land in Manitoba, unless it

is adopted as a cover-crop BMP for nutrient management (section 7.3.3). It also has implications to

nutrient release through break down of plant material by freeze/thaw cycles (sections 7.3.3 and 8.1).

Regarding the effect of individual crops within a cropping system, the results were inconclusive due

to the similarity between crop-years and the carry-over effect from one year to the next. The effect of

annual crops may be better assessed when comparing annual cropping systems with different nutrient

requirements. For example, a more nutrient-intensive cropping system including potato, canola and

78

corn in rotation will likely have a greater potential for nutrient export than a cropping system

containing canola and wheat in rotation. However, tile drainage itself could support intensification of

cropping systems by promoting field conditions more favorable for cash crops in areas where they

were not possible before tile drainage was implemented. Finally, adoption of cover crops seems not to

be widespread in Manitoba at larger scales. Thus, it does not significantly impact nutrient export

under current conditions. However, it has potential to be used as a BMP for nutrient capture and

should receive further consideration to reduce nutrient export from tile drained land.

The impact of the agronomic practices discussed above on mitigation of nutrient export through tile

drainage systems have not been assessed at field-scale yet. The recent, fast adoption of tile drainage

has not allowed for a thorough assessment of the interaction between this practice and other

agronomic practices prevalent in Manitoba. However, the rapid and widespread expansion warrants

research in this field in order to optimize agronomic benefits of tile drainage and minimize its

negative impacts.

9. KNOWLEDGE GAPS AND RESEARCH NEEDS/PRIORITIES Hydrology and nutrient transport in cold climates represent a major gap in the scientific

understanding regarding tile drainage (King et al., 2015b). Research needs on tile drainage have been

outlined below taking into consideration key physiographic features and agronomic practices relevant

to Manitoba. The sequence does not imply any ranking of priority; rather it follows the order of

sections of this review.

Custom design of tile drainage systems for different soil classes: Recent investigations

carried out for sandy loam soils indicated that the design proposed by the Province of

Manitoba may not be the most appropriate for that soil. Chances are that the design is not

optimum for other soils, as well. Thus, investigations of design criteria for other soil types

are required. A modelling framework such as the one used in previous studies (i.e.,

79

DRAINMOD) is probably adequate. However, this approach requires results from multi-

year field monitoring for model calibration and validation, which highlights the need of an

increased effort for monitoring of tile drainage systems. The essential variables for a

hydrological assessment of the system include drainage outflow volume and water table

elevation. However, other variables important for hydrological, environmental, and

agronomic assessments include soil moisture and soil fertility variation, crop yields, crop

nutrient uptake and removal, as well as water quality data (i.e., nutrient concentrations).

Water budgets for tile drained fields with different soils: Results conducted in Manitoba so

far did not include monitoring of surface runoff. As discussed in this review, this variable is

a key component for calculating the total nutrient export from tile drained land and,

therefore, it should be monitored over several years of contrasting precipitation patterns in

order to cast some light on the effect of tile drainage on the overall nutrient dynamics of

tile-drained fields.

Impact of the 4R nutrient management framework on nutrient export through tile drained

land: Studies would aid in the establishment of optimal rates and timing of application for

the most commonly used fertilizers in Manitoba.

Nutrient export from manured and synthetically-fertilized fields: This is a special case of

the need mentioned above and was emphasized as a separate issue since potential

differences between these two nutrient sources are being discussed in Manitoba and

prompted this literature review. While the results discussed here suggest no difference

between the sources, field investigations are important to factor in the effect of

physiography and agronomic practices. Comparisons should include different types, rates,

timing, and application methods for both manure and synthetic fertilizers, as well as

different soils, crops, farm operations (e.g., tillage) and design criteria of the tile drainage

80

system.

Pathogen export through tile drainage systems.

Transport of nutrients and pathogens to tile drains through preferential flow paths: Long-

term, continuous monitoring is required to assess its influence under Manitoba conditions.

Monitoring of critical periods such as end-of-season (due to soil desiccation) and early

snowmelt (due to freezing) could be studied using field techniques or desktop methods

(e.g., hydrograph separation).

Effect of tillage on disrupting preferential flow paths: This practice could be assessed to

identify how effective different tillage forms (e.g., implements) and depths are in disrupting

preferential flow paths. The longevity of the effect could also be assessed in order to define

the impact of freeze-thaw cycles on persistence/re-enacting of preferential flow.

Nutrient and pathogen abatement from tile drainage outflow by water retention structures:

Water retention has been listed as one of the key actions in the Manitoba Surface Water

Management Strategy (Manitoba Conservation and Water Stewardship, 2014). Investigating

the potential nutrient capture, removal and reuse from these structures is a way to expand

the functions of such structures and increase potential for adoption.

Efficacy of biofilters to treat tile drainage outflow: The main question under this item is

whether these structures will be effective to treat (capture, removal and reuse) the nutrient

concentrations in tile outflow taking into consideration the conditions prevailing during the

drainage season (i.e., low temperature and high volumes, which affects flow rate and

residence time).

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10. DISCLAIMER Any data, analyses of data, project results and conclusions conveyed in this report are those of the

project researchers and not of the governments of Canada or Manitoba.

11. ACKNOWLEDGEMENTS This project is supported by the Manitoba Livestock Manure Management Initiative (MLMMI).

MLMMI is funded by the governments of Canada and Manitoba through Growing Forward 2, a

federal-provincial-territorial initiative.

12. REFERENCES Acharya SN. 2014. AAC Meadowview alfalfa. Canadian Journal of Plant Science, 94: 187-189. DOI:

10.4141/cjps2013-2491.

Akinremi OO, McGinn SM, Cutforth HW. 1999. Precipitation trends on the Canadian Prairies. Journal of

Climate, 12: 2996-3003. DOI: 10.1175/1520-0442(1999)012<2996:PTOTCP>2.0.CO;2.

Aksoy H, Kavvas ML. 2005. A review of hillslope and watershed scale erosion and sediment transport models.

Catena, 64: 247-271. DOI: http://dx.doi.org/10.1016/j.catena.2005.08.008.

ASABE. 2005. Manure Production and Characteristics. ASABE.

ASABE. 2013. Operation of Controlled Drainage Systems in Humid Regions. ASABE, pp: 1.

ASABE. 2015. Design and Construction of Subsurface Drainage Systems on Agricultural Lands in Humid

Areas. ASABE, pp: 1.

Asawa GL. 2006. Irrigation and Water Resources Engineering. New Delhi, India: New Age International (P)

Ltd. Publishers.

Ayars JE, Christen EW, Hornbuckle JW. 2006. Controlled drainage for improved water management in arid

regions irrigated agriculture. Agricultural Water Management, 86: 128-139. DOI: http://dx.doi.org/

10.1016/j.agwat.2006.07.004.

Baiyeri K, Tenkouano A. 2007. Manure placement influenced growth and dry matter yield of a plantain hybrid.

African Crop Science Conference Proceedings, 8: 385-390.

Ball Coelho BR, Roy RC, Topp E, Lapen DR. 2007. Tile water quality following liquid swine manure

application into standing corn. Journal of Environmental Quality, 36: 580-587. DOI: 10.2134/jeq2006.0306.

Barling RD, Moore ID. 1994. Role of buffer strips in management of waterway pollution: A review.

Environmental Management, 18: 543-558. DOI: 10.1007/bf02400858.

Beasley DB, Huggins LF, Monke EJ. 1980. ANSWERS: A model for watershed planning. Transactions of the

American Society of Agricultural Engineers, 23: 938-944. DOI: 10.13031/2013.34692.

Beauchemin S, Simard RR, Cluis D. 1998. Forms and concentration of phosphorus in drainage water of

twenty-seven tile-drained soils. Journal of Environmental Quality, 27: 721-728. DOI: 10.2134/jeq1998.

00472425002700030033x.

Beaulac MN, Reckhow KH. 1982. An examination of land use - nutrient export relationships. Journal of the

American Water Resources Association, 18: 1013-1024. DOI: 10.1111/j.1752-1688.1982.tb00109.x.

Bechmann ME, Kleinman PJA, Sharpley AN, Saporito LS. 2005. Freeze–Thaw Effects on Phosphorus Loss in

Runoff from Manured and Catch-Cropped Soils. Journal of Environmental Quality, 34: 2301-2309. DOI:

82

10.2134/jeq2004.0415.

Beegle DB, Carton OT, Bailey JS. 2000. Nutrient management planning: Justification, theory, practice. Journal

of Environmental Quality, 29. DOI: 10.2134/jeq2000.00472425002900010009x.

Bergström L, Brink N. 1986. Effects of differentiated applications of fertilizer N on leaching losses and

distribution of inorganic N in the soil. Plant and Soil, 93: 333-345. DOI: 10.1007/bf02374284.

Bernstein L. 1974. Crop Growth and Salinity. In: Drainage for Agriculture, Van Schilfgaarde J (ed.) American

Society of Agronomy, pp: 39-54.

Beven KJ. 2011. Rainfall-Runoff Modelling: The Primer. West Sussex, UK: Wiley.

Bianchi TS, Allison MA, Cai WJ. 2013. Biogeochemical Dynamics at Major River-Coastal Interfaces:

Linkages with Global Change: Cambridge University Press.

Bicudo JR, Goyal SM. 2003. Pathogens and manure management systems: A review. Environmental

Technology, 24: 115-130. DOI: 10.1080/09593330309385542.

Bjorneberg DL, Kanwar RS, Melvin SW. 1996a. Seasonal changes in flow and nitrate-N loss from subsurface

drains. Transactions of the American Society of Agricultural Engineers, 39: 961-967. DOI: 10.13031/

2013.27582.

Bjorneberg DL, Kanwar Rs, Melvin SW. 1996b. Tillage and crop rotation effects on subsurface drainage

response to rainfall. Transactions of the American Society of Agricultural Engineers, 39: 2147-2154. DOI:

10.13031/2013.27718.

Blackshaw RE, Molnar LJ, Moyer JR. 2010. Suitability of legume cover crop-winter wheat intercrops on the

semi-arid Canadian Prairies. Canadian Journal of Plant Science, 90: 479-488. DOI: 10.4141/CJPS10006.

Blair GJ, Freney JR, Park JK. 1990. Effect of sulfur, silicon, and trace metal interactions in determining the

dynamics of phosphorus in agricultural systems. In: Phosphorus requirements for sustainable agriculture in

Asia and Oceania, pp: 269-280, Manila, Philippines, 6-10 March 1989.

Blann KL, Anderson JL, Sands GR, Vondracek B. 2009. Effects of agricultural drainage on aquatic ecosystems:

A review. Critical Reviews in Environmental Science and Technology, 39: 909-1001. DOI:

10.1080/10643380801977966.

Blowes DW, Robertson WD, Ptacek CJ, Merkley C. 1994. Removal of agricultural nitrate from tile-drainage

effluent water using in-line bioreactors. Journal of Contaminant Hydrology, 15: 207-221. DOI: http://dx.

doi.org/10.1016/0169-7722(94)90025-6.

Bolton EF, Aylesworth JW, Hore FR. 1970. Nutrient losses through tile drains under three cropping systems

and two fertility levels on a Brookston clay soil. Canadian Journal of Soil Science, 50: 275-279. DOI:

10.4141/cjss70-038.

Bower SS. 2006. Wet prairie: an environmental history of wetlands, flooding and drainage in agricultural

Manitoba, 1810-1980. PhD Thesis. Vancouver, B.C.: The University of British Columbia, Department of

Geography.

Bower SS. 2007. Watersheds: Conceptualizing Manitoba's drained landscape, 1895–1950. Environmental

History, 12: 796-819. DOI: 10.1093/envhis/12.4.796.

Boydaş MG, Turgut N. 2007. Effect of tillage implements and operating speeds on soil physical properties and

wheat emergence. Turkish Journal of Agriculture and Forestry, 31: 399-412.

Brown KW, Duble RL, Thomas JC. 1977. Influence of management and season on fate of n applied to golf

greens. Agronomy Journal, 69: 667-671. DOI: 10.2134/agronj1977.00021962006900040036x.

Brunet NN, Westbrook CJ. 2012. Wetland drainage in the Canadian prairies: Nutrient, salt and bacteria

characteristics. Agriculture, Ecosystems & Environment, 146: 1-12. DOI: http://dx.doi.org/10.1016/j.agee.

2011.09.010.

83

Bullock DG. 1992. Crop rotation. Critical Reviews in Plant Sciences, 11: 309-326. DOI: 10.1080/

07352689209382349.

Cade-Menun BJ, Bell G, Baker-Ismail S, Fouli Y, Hodder K, McMartin DW, Perez-Valdivia C, Wu K. 2013.

Nutrient loss from Saskatchewan cropland and pasture in spring snowmelt runoff. Canadian Journal of Soil

Science, 93: 445-458. DOI: 10.4141/cjss2012-042.

Calhoun FG, Baker DB, Slater BK. 2002. Soils, water quality, and watershed size. Journal of Environmental

Quality, 31: 47-53. DOI: 10.2134/jeq2002.4700.

Campbell CA, Zentner RP, Janzen HH, Bowren KE. 1990. Crop Rotation Studies on the Canadian Prairies:

Research Branch, Agriculture Canada.

Carpenter SR, Caraco NF, Correll DL, Howarth RW, Sharpley AN, Smith VH. 1998. Nonpoint pollution of

surface waters with phosphorus and nitrogen. Ecological Applications, 8: 559-568. DOI: 10.1890/1051-

0761(1998)008[0559:NPOSWW]2.0.CO;2.

Chahinian N, Moussa R, Andrieux P, Voltz M. 2005. Comparison of infiltration models to simulate flood

events at the field scale. Journal of Hydrology, 306: 191-214. DOI: http://dx.doi.org/10.1016/

j.jhydrol.2004.09.009.

Chambers BJ, Smith KA, Pain BF. 2000. Strategies to encourage better use of nitrogen in animal manures. Soil

Use and Management, 16: 157-166. DOI: 10.1111/j.1475-2743.2000.tb00220.x.

Chen Y, Tessier S, Cavers C, Xu X, Monero F. 2005. A survey of crop residue burning practices in Manitoba.

Applied Engineering in Agriculture, 21: 317-323. DOI: 10.13031/2013.18446.

Chien S, Sale P, Hammond L. 1990. Comparison of the effectiveness of phosphorus fertilizer products. In:

Phosphorus requirements for sustainable agriculture in Asia and Oceania, pp: 143-156, Manila,

Philippines, 6-10 March 1989.

Chow VT. 1964. Handbook of Applied Hydrology: A Compendium of Water-Resources Technology: McGraw-

Hill.

Christianson L, Bhandari A, Helmers M, Kult K, Sutphin T, Wolf R. 2012a. Performance evaluation of four

field-scale agricultural drainage denitrification bioreactors in iowa. Transactions of the American Society of

Agricultural Engineers, 55: 2163-2174. DOI: 10.13031/2013.42508.

Christianson LE, Bhandari A, Helmers MH, Clair MS. 2009. Denitrifying bioreactors for treatment of tile

drainage. In: World Environmental and Water Resources Congress 2009, pp: 1-10, Kansas City, MO.

Christianson LE, Bhandari A, Helmers MJ. 2011. Pilot-scale evaluation of denitrification drainage bioreactors:

Reactor geometry and performance. Journal of Environmental Engineering, 137: 213-220. DOI:

doi:10.1061/(ASCE)EE.1943-7870.0000316.

Christianson LE, Bhandari A, Helmers MJ. 2012b. A practice-oriented review of woodchip bioreactors for

subsurface agricultural drainage. Applied Engineering in Agriculture, 28: 861-874. DOI: 10.13031/

2013.42479.

Chun JA, Cooke RA, Eheart JW, Cho J. 2010. Estimation of flow and transport parameters for woodchip-based

bioreactors: II. field-scale bioreactor. Biosystems Engineering, 105: 95-102. DOI: http://dx.doi.org/10.1016/

j.biosystemseng.2009.09.018.

Coelho BRB, Roy RC, Bruin AJ, More A, White P. 2009. Zonejection: Conservation tillage manure nutrient

delivery system. Agronomy Journal, 101: 215-225. DOI: 10.2134/agronj2008.0001x.

Colhoun J. 1973. Effects of environmental factors on plant disease. Annual Review of Phytopathology, 11: 343-

364. DOI: doi:10.1146/annurev.py.11.090173.002015.

Collins RE. 1961. Flow of Fluids Through Porous Materials. New York, NY: Reinhold Publishing Corp.

Conley DJ, Paerl HW, Howarth RW, Boesch DF, Seitzinger SP, Havens KE, Lancelot C, Likens GE. 2009.

84

Controlling eutrophication: nitrogen and phosphorus. Science, 323: 1014-1015.

Cook MJ, Baker JL. 2001. Bacteria and nutrient transport to tile lines shortly after application of large volumes

of liquid swine manure. Transactions of the American Society of Agricultural Engineers, 44: 495. DOI:

10.13031/2013.6109.

Cooke RA, Bell NL. 2014. Protocol and interactive routine forthe design of subsurface bioreactors. Applied

Engineering in Agriculture, 30: 761-771. DOI: 10.13031/aea.30.9900.

Cordeiro MRC, Krahn V, Sri Ranjan R, Sager S. 2015a. Water table contribution and diurnal water

redistribution within the corn root zone. Canadian Biosystems Engineering 57: 1.39-31.48. DOI:

10.7451/CBE.2015.56.1.39.

Cordeiro MRC, Krahn V, Sri Ranjan R, Sager S. 2015b. Water table contribution and diurnal water

redistribution within the corn root zone. Canadian Biosystems Engineering, 57: 1.39-31.48. DOI:

10.7451/cbe.2015.57.1.39.

Cordeiro MRC, Sri Ranjan R. 2012. Corn yield response to drainage and subirrigation in the Canadian Prairies.

Transactions of the American Society of Agricultural Engineers, 55: 1771-1780. DOI: doi: 10.13031/

2013.42369.

Cordeiro MRC, Sri Ranjan R. 2015. DRAINMOD simulation of corn yield under different tile drain spacing in

the Canadian Prairies. Transactions of the American Society of Agricultural Engineers, 58: 1481-1491. DOI:

10.13031/trans.58.10768.

Cordeiro MRC, Sri Ranjan R, Ferguson IJ, Cicek N. 2014. Nitrate, phosphorus, and salt export through

subsurface drainage from corn fields in the Canadian Prairies. Transactions of the American Society of

Agricultural Engineers, 57: 43-50. DOI: 10.13031/trans.57.10370.

Corey AT. 1994. Mechanics of Immiscible Fluids in Porous Media: Water Resources Publications.

Crumpton W, Stenback G, Miller B, Helmers M. 2006. Potential benefits of wetland filters for tile drainage

systems: Impact on nitrate loads to Mississippi River subbasins. Project completion report. In: Ecology,

Evolution and Organismal Biology Reports, pp: 34.

Culley JLB, Bolton EF, Bernyk V. 1983. Suspended solids and phosphorus loads from a clay soil: I. Plot

studies. Journal of Environmental Quality, 12: 493-498. DOI: 10.2134/jeq1983.00472425001200040011x.

Dadfar H, Allaire SE, De Jong R, van Bochove E, Denault JT, Thériault G, Dechmi F. 2010. Development of a

method for estimating the likelihood of crack flow in Canadian agricultural soils at the landscape scale.

Canadian Journal of Soil Science, 90: 129-149. DOI: 10.4141/CJSS09066.

Daniel TC, Sharpley AN, Lemunyon JL. 1998. Agricultural phosphorus and eutrophication: A symposium

overview. Journal of Environmental Quality, 27: 251-257. DOI: 10.2134/jeq1998.

00472425002700020002x.

David MB, Drinkwater LE, McIsaac GF. 2010. Sources of Nitrate Yields in the Mississippi River Basin.

Journal of Environmental Quality, 39: 1657-1667. DOI: 10.2134/jeq2010.0115.

Dayton EA, Whitacre SD, Holloman CH. 2014. Demonstrating the relationship between soil phosphorus

measures and phosphorus solubility: Implications for Ohio phosphorus risk assessment tools. Journal of

Great Lakes Research, 40: 473-478. DOI: http://dx.doi.org/10.1016/j.jglr.2014.04.001.

Dean DM, Foran ME. 1992. The effect of farm liquid waste application on tile drainage. Journal of Soil and

Water Conservation, 47: 368-369.

Dell CJ, Meisinger JJ, Beegle DB. 2011. Subsurface application of manures slurries for conservation tillage

and pasture soils and their impact on the nitrogen balance. Journal of Environmental Quality, 40: 352-361.

DOI: 10.2134/jeq2010.0069.

Dietrich Sr WJ. 2005. Slurry fertilizer applicator. US Patent Classification

85

Dingman SL. 2008. Physical Hydrology: Waveland Press.

Dinnes DL, Karlen DL, Jaynes DB, Kaspar TC, Hatfield JL, Colvin TS, Cambardella CA. 2002. Nitrogen

management strategies to reduce nitrate leaching in tile-drained Midwestern soils. Agronomy Journal, 94:

153-171. DOI: 10.2134/agronj2002.1530.

Doty CW, Parsons JE, Nassehzadeh-Tabrizi A, Skaggs RW, Badr AW. 1984. Stream water levels affect field

water tables and corn yields. Transactions of the American Society of Agricultural Engineers, 27: 1300-

1306. DOI: 10.13031/2013.32963.

Dring C, Devlin JF, Boag G, Sunohara MD, Fitzgibbon J, Topp E, Lapen DR. 2015. Incentives and

disincentives identified by producers and drainage contractors/experts on the adoption of controlled tile

drainage in eastern Ontario, Canada. Water Quality Research Journal of Canada, 51: 1-16. DOI:

10.2166/wqrjc.2015.047.

Drury CF, Tan CS, Reynolds WD, Welacky TW, Oloya TO, Gaynor JD. 2009. Managing tile drainage,

subirrigation, and nitrogen fertilization to enhance crop yields and reduce nitrate loss. Journal of

Environmental Quality, 38. DOI: 10.2134/jeq2008.0036.

Drury CF, Tan CS, Welacky TW, Reynolds WD, Zhang TQ, Oloya TO, McLaughlin NB, Gaynor JD. 2014.

Reducing nitrate loss in tile drainage water with cover crops and water-table management systems. Journal

of Environmental Quality, 43: 587-598. DOI: 10.2134/jeq2012.0495.

Duhovnik J, Benedicic J, Bernik R. 2004. Analysis and design parameters for inclined rotors used for manure

dispersal on broadcast spreaders for solid manure. Transactions of the American Society of Agricultural

Engineers, 47: 1389-1404. DOI: 10.13031/2013.17604.

Eastman M, Gollamudi A, Stämpfli N, Madramootoo CA, Sarangi A. 2010. Comparative evaluation of

phosphorus losses from subsurface and naturally drained agricultural fields in the Pike River watershed of

Quebec, Canada. Agricultural Water Management, 97: 596-604. DOI: http://dx.doi.org/10.1016/

j.agwat.2009.11.010.

Easton ZM, Petrovic AM. 2004. Fertilizer source effect on ground and surface water quality in drainage from

turfgrass. Journal of Environmental Quality, 33: 645-655. DOI: 10.2134/jeq2004.6450.

Eghball B, Power JF. 1994. Beef cattle feedlot manure management. Journal of Soil and Water Conservation,

49: 113-122.

Eghball B, Power JF. 1999. Phosphorus- and nitrogen-based manure and compost applications corn production

and soil phosphorus. Soil Science Society of America Journal, 63: 895-901. DOI: 10.2136/sssaj1999.

634895x.

Ehlers w. 1975. Observations on earthworm channels and infiltration on tilled and untilled loess soil. Soil

Science, 119: 242-249.

Errebhi M, Rosen CJ, Gupta SC, Birong DE. 1998. Potato yield response and nitrate leaching as influenced by

nitrogen management. Agronomy Journal, 90: 10-15. DOI: 10.2134/agronj1998.00021962009000010003x.

Evans RO, Westerman PW, Overcash MR. 1984. Subsurface drainage water quality from land application of

swine lagoon effluent. Transactions of the American Society of Agricultural Engineers, 27: 473-480. DOI:

10.13031/2013.32813.

Everts CJ, Kanwar RS. 1990. Estimating preferential flow to a subsurface drain with tracers. Transactions of

the American Society of Agricultural Engineers, 33: 451-457. DOI: 10.13031/2013.31350.

Fageria NK. 2008. The Use of Nutrients in Crop Plants. Boca Raton, FL.: CRC Press.

Fang X, Pomeroy JW. 2007. Snowmelt runoff sensitivity analysis to drought on the Canadian prairies.

Hydrological Processes, 21: 2594-2609. DOI: 10.1002/hyp.6796.

Farming for Tomorrow. (2011). Tile drainage research begins in Manitoba. Retrieved Januray 27, 2016 http://

86

www.farmingfortomorrow.ca/index.php/previous-issues-2011

Fetter CW. 1999. Contaminant Hydrogeology: Prentice Hall.

Fleming RA, Babcock B, Wang E. 1998a. Resource or Waste? The Economics of Swine Manure Storage and

Management. Review of Agricultural Economics, 20: 96-113. DOI: 10.2307/1349536.

Fleming RJ, MacAlpine M, Tiffin C. 1998b. Nitrate levels in soil, tile drainage water and shallow groundwater

under a variety of farm management systems. Canadian Society for Engineering in Agricultural, Food, and

Biological Systems, pp: 98-101.

Fouss JL, Evans RO, Ayars JE, Christen EW. 2007. Water Table Control Systems. In: Design and Operation of

Farm Irrigation Systems, Hoffman GJ (ed.) American Society of Agricultural and Biological Engineers, pp:

684-724.

Fouss JL, Reeve RC. 1987. Advances in Drainage Technology: 1955-85. In: Farm Drainage in the United

States: History, Status, and Prospects, Pavelis GA (ed.) U.S. Department of Agriculture, Economic

Research Service.

Fox GA, Marvin MM, Guzman JA, Hoang CK, Malone RW, Kanwar RS, Shipitalo MJ. 2012. E. Coli transport

through surface-connected biopores identified from smoke injection tests. Transactions of the American

Society of Agricultural Engineers, 55: 2185-2194. DOI: 10.13031/2013.42511.

Fraser H, Fleming RJ. 2001. Environmental benefits of tile drainage. University of Guelph.

Frey SK, Topp E, Ball BR, Edwards M, Gottschall N, Sunohara M, Zoski E, Lapen DR. 2013. Tile drainage

management influences on surface-water and groundwater quality following liquid manure application.

Journal of Environmental Quality, 42: 881-892. DOI: 10.2134/jeq2012.0261.

Friesen GH, Wall DA. 1986. Tolerance of early-maturing soybean cultivars to metribuzin. Canadian Journal of

Plant Science, 66: 125-130. DOI: 10.4141/cjps86-016.

Gasser JKR. 1987. The future of animal manures as fertilizer or waste. In: Animal Manure on Grassland and

Fodder Crops. Fertilizer or Waste? Proceedings of an International Symposium of the European Grassland

Federation, Wageningen, The Netherlands, 31 August–3 September 1987, Meer HG, Unwin RJ, Dijk TA,

Ennik GC (eds.) Springer Netherlands, pp: 259-278.

Gast RG, Nelson WW, Randall GW. 1978. Nitrate accumulation in soils and loss in tile drainage following

nitrogen applications to continuous corn. Journal of Environmental Quality, 7: 258-261. DOI: 10.2134/

jeq1978.00472425000700020021x.

Gayle GA, Skaggs RW. 1978. Surface storage on bedded cultivated lands. Transactions of the American

Society of Agricultural Engineers, 21: 101-104. DOI: 10.13031/2013.35258.

Gaynor JD, Findlay WI. 1995. Soil and phosphorus loss from conservation and conventional tillage in corn

production. Journal of Environmental Quality, 24: 734-741. DOI: 10.2134/jeq1995.

00472425002400040026x.

Gburek WJ, Sharpley AN, Heathwaite L, Folmar GJ. 2000. Phosphorus management at the watershed scale: A

modification of the phosphorus index. Journal of Environmental Quality, 29: 130-144. DOI:

10.2134/jeq2000.00472425002900010017x.

Girard JE, Girard J. 2014. Principles of Environmental Chemistry: Jones & Bartlett Learning.

Givi J, Prasher SO, Patel RM. 2004. Evaluation of pedotransfer functions in predicting the soil water contents

at field capacity and wilting point. Agricultural Water Management, 70: 83-96. DOI: http://dx.doi.org/

10.1016/j.agwat.2004.06.009.

Goswami NN, Kamath MB, Santoso D. 1990. Phosphorus requirements and management of maize, sorghum,

and wheat. In: Phosphorus requirements for sustainable agriculture in Asia and Oceania, pp: 349-359,

Manila, Philippines, 6-10 March 1989.

87

Grass LB, MacKenzie AJ, Willardson LS, Region USARSW. 1975. Inspecting and cleaning subsurface drain

systems. Washington, DC: U.S. Dept. of Agriculture.

Gray D, Prowse T. 1993. Snow and Floating Ice. In: Handbook of Hydrology, Maidment DR (ed.) McGraw-

Hill, pp: 7.1-7.58.

Gray DM, Toth B, Zhao L, Pomeroy JW, Granger RJ. 2001. Estimating areal snowmelt infiltration into frozen

soils. Hydrological Processes, 15: 3095-3111. DOI: 10.1002/hyp.320.

Guan TY, Holley RA. 2003. Pathogen survival in swine manure environments and transmission of human

enteric illness—a review. Journal of Environmental Quality, 32: 383-392. DOI: 10.2134/jeq2003.3830.

Harris B. (2015a). RM of Portage researching tile drainage issues. Retrieved February 01, 2016 http://www.

portageonline.com/local/44691-rm-of-portage-researches-tile-drainage-issues

Harris B. (2015b). RM addresses manure on tile-drained land. Retrieved February 01, 2016

http://www.portageonline.com/local/45297-rm-addresses-manure-on-tile-drained-land

Hart MR, Quin BF, Nguyen ML. 2004. Phosphorus runoff from agricultural land and direct fertilizer effects.

Journal of Environmental Quality, 33: 1954-1972. DOI: 10.2134/jeq2004.1954.

Hartz TK, Johnstone PR. 2006. Relationship between soil phosphorus availability and phosphorus loss

potential in runoff and drainage. Communications in Soil Science and Plant Analysis, 37: 1525-1536. DOI:

10.1080/00103620600710058.

Hassanpour B, Parsinejad M, Yazdani MR, Dalivand FS, Kossari H. 2011. Evaluation of modified

DRAINMOD in predicting groundwater table fluctuations and yield of canola in paddy fields under snowy

conditions (case study: Rasht Iran). Irrigation and Drainage, 60: 660-667. DOI: 10.1002/ird.614.

He ZL, Wilson MJ, Campbell CO, Edwards AC, Chapman SJ. 1995. Distribution of phosphorus in soil

aggregate fractions and its significance with regard to phosphorus transport in agricultural runoff. Water, Air,

and Soil Pollution, 83: 69-84. DOI: 10.1007/bf00482594.

Heathwaite AL, Griffiths P, Parkinson RJ. 1998. Nitrogen and phosphorus in runoff from grassland with buffer

strips following application of fertilizers and manures. Soil Use and Management, 14: 142-148. DOI:

10.1111/j.1475-2743.1998.tb00140.x.

Heckman JR, Sims JT, Beegle DB, Coale FJ, Herbert SJ, Bruulsema TW, Bamka WJ. 2003. Nutrient removal

by corn grain harvest. Agronomy Journal, 95: 587-591. DOI: 10.2134/agronj2003.5870.

Heckrath G, Brookes PC, Poulton PR, Goulding KWT. 1995. Phosphorus leaching from soils containing

different phosphorus concentrations in the broadbalk experiment. Journal of Environmental Quality, 24:

904-910. DOI: 10.2134/jeq1995.00472425002400050018x.

Hergert GW, Klausner SD, Bouldin DR, Zwerman PJ. 1981. Effects of dairy manure on phosphorus

concentrations and losses in tile effluent. Journal of Environmental Quality, 10: 345-349. DOI: 10.2134/

jeq1981.00472425001000030018x.

Hickey MBC, Doran B. 2004. A review of the efficiency of buffer strips for the maintenance and enhancement

of riparian ecosystems. Water Quality Research Journal of Canada, 39: 311-317.

Hiler EA. 1969. Quantitative evaluation of crop-drainage requirements. Transactions of the American Society

of Agricultural Engineers, 12: 499-505. DOI: 10.13031/2013.38876.

Hillel D. 1998. Environmental Soil Physics: Fundamentals, Applications, and Environmental Considerations.

London, UK: Elsevier Science.

Hodgkinson RA, Chambers BJ, Withers PJA, Cross R. 2002. Phosphorus losses to surface waters following

organic manure applications to a drained clay soil. Agricultural Water Management, 57: 155-173. DOI:

http://dx.doi.org/10.1016/S0378-3774(02)00057-4.

Hoffmann CC, Kjaergaard C, Uusi-Kämppä J, Hansen HCB, Kronvang B. 2009. Phosphorus retention in

88

riparian buffers: Review of their efficiency. Journal of Environmental Quality, 38: 1942-1955. DOI:

10.2134/jeq2008.0087.

Hoorman JJ, Shipitalo MJ. 2006. Subsurface drainage and liquid manure. Journal of Soil and Water

Conservation, 61: 94A-97A.

Howarth RW, Sharpley A, Walker D. 2002. Sources of nutrient pollution to coastal waters in the United States:

Implications for achieving coastal water quality goals. Estuaries, 25: 656-676. DOI: 10.1007/bf02804898.

International Fertilizer Industry Association. 2009. The global “4R” nutrient stewardship framework:

Developing fertilizer best management practices for delivering economic, social, and environmental

benefits. International Fertilizer Industry Association.

Irwin RW. 1967. Drainage problems with organic soil. Transactions of the American Society of Agricultural

Engineers, 10: 427-428. DOI: 10.13031/2013.39690.

Jacobsen SM, J.D. W, Weber R, Willey PH. 2010. ND drain computer software for lateral effect evaluation of a

subsurface drain. ASABE.

Jamieson A, Madramootoo C, Enright P. 2003. Phosphorus losses in surface and subsurface runoff from a

snowmelt event on an agricultural field in Quebec. Canadian Biosystems Engineering, 45: 1.1-1.7.

Jamieson R, Gordon R, Sharples K, Stratton G, Madani A. 2002. Movement and persistence of fecal bacteria in

agricultural soils and subsurface drainage water: A review. Canadian Biosystems Engineering, 44: 1.1-1.9.

Jaynes DB. 2013. Nitrate loss in subsurface drainage and corn yield as affected by timing of sidedress nitrogen.

Agricultural Water Management, 130: 52-60. DOI: http://dx.doi.org/10.1016/j.agwat.2013.08.010.

Jaynes DB, Colvin TS, Karlen DL, Cambardella CA, Meek DW. 2001. Nitrate loss in subsurface drainage as

affected by nitrogen fertilizer rate. Journal of Environmental Quality, 30. DOI: 10.2134/jeq2001.3041305x.

Jaynes DB, Isenhart TM. 2014. Reconnecting tile drainage to riparian buffer hydrology for enhanced nitrate

removal. Journal of Environmental Quality, 43: 631-638. DOI: 10.2134/jeq2013.08.0331.

Jiang Y, Somers G. 2008. Modeling effects of nitrate from non-point sources on groundwater quality in an

agricultural watershed in Prince Edward Island, Canada. Hydrogeology Journal, 17: 707-724. DOI:

10.1007/s10040-008-0390-2.

Jin CX, Sands GR. 2003. The long–term field–scale hydrology of subsurface drainage systems in a cold

climate. Transactions of the American Society of Agricultural Engineers, 46: 1011–1021. DOI: 10.13031

/2013.13963.

John MS, Donald LR. 1978. Nonpoint sources: State-of-the-art overview. Transactions of the ASABE, 21: 565-

573. DOI: 10.13031/2013.35329.

Johnson AI. 1967. Specific yield: compilation of specific yields for various materials. In: Water Supply Paper.

Johnson MH, Jr, Thomas DL, McLendon BD. 1993. Controlled-drainage/subirrigation system automation

based on soil water potential. Transactions of the ASABE, 36: 751-759. DOI: 10.13031/2013.28394.

Joshi B, Maulé C. 1999. Vadose zone flow and transport in cracked heavy textured soil. Journal of Irrigation

and Drainage Engineering, 125: 122-130. DOI: doi:10.1061/(ASCE)0733-9437(1999)125:3(122).

Joy DM, Lee H, Reaume CM, Whiteley HR, Zelin S. 1998. Microbial contamination of subsurface tile

drainage water from field applications of liquid manure. Canadian Agricultural Engineering, 40: 153-160.

Kahimba F, Ranjan RS, Froese J, Entz M, Nason R. 2008. Cover crop effects on infiltration, soil temperature,

and soil moisture distribution in the Canadian Prairies. Applied Engineering in Agriculture, 24: 321.

Kandel HJ, Brodshaug JA, Steele DD, Ransom JK, DeSutter TM, Sands GR. 2013. Subsurface drainage effects

on soil penetration resistance and water table depth on a clay soil in the Red River of the North Valley,

USA. Agricultural Engineering International: CIGR Journal, 15: 1-10.

89

Kaspar TC, Jaynes DB, Parkin TB, Moorman TB, Singer JW. 2012. Effectiveness of oat and rye cover crops in

reducing nitrate losses in drainage water. Agricultural Water Management, 110: 25-33. DOI: http://dx.

doi.org/10.1016/j.agwat.2012.03.010.

Kay P, Blackwell PA, Boxall ABA. 2004. Fate of veterinary antibiotics in a macroporous tile drained clay soil.

Environmental Toxicology and Chemistry, 23: 1136-1144. DOI: 10.1897/03-374.

Kellogg RL, Lander CH, Moffitt DC, Gollehon N. 2000. Manure nutrients relative to the capacity of cropland

and pastureland to assimilate nutrients: Spatial and temporal trends for the united states. Proceedings of the

Water Environment Federation, 2000: 140. DOI: 10.2175/193864700784994812.

Kengni L, Vachaud G, Thony JL, Laty R, Garino B, Casabianca H, Jame P, Viscogliosi R. 1994. Field

measurements of water and nitrogen losses under irrigated maize. Journal of Hydrology, 162: 23-46. DOI:

http://dx.doi.org/10.1016/0022-1694(94)90003-5.

King KW, Williams MR, Fausey NR. 2015a. Contributions of systematic tile drainage to watershed-scale

phosphorus transport. Journal of Environmental Quality, 44: 486-494. DOI: 10.2134/jeq2014.04.0149.

King KW, Williams MR, Macrae ML, Fausey NR, Frankenberger J, Smith DR, Kleinman PJA, Brown LC.

2015b. Phosphorus transport in agricultural subsurface drainage: A review. Journal of Environmental

Quality, 44: 467-485. DOI: 10.2134/jeq2014.04.0163.

King LG, Willardson LS. 2007. Drainage systems. In: Design and Operation of Farm Irrigation Systems,

Hoffman GJ (ed.) ASABE, pp: 289-319.

Kinley RD, Gordon RJ, Stratton GW, Patterson GT, Hoyle J. 2007. Phosphorus losses through agricultural tile

drainage in Nova Scotia, Canada. Journal of Environmental Quality, 36: 469-477. DOI: 10.2134/

jeq2006.0138.

Kirkby EA, Johnston AE. 2008. Soil and fertilizer phosphorus in relation to crop nutrition. In: The

Ecophysiology of Plant-Phosphorus Interactions, White PJ, Hammond JP (eds.) Springer Netherlands, pp:

177-223.

Kleinman PJA, Sharpley AN. 2003. Effect of broadcast manure on runoff phosphorus concentrations over

successive rainfall events. Journal of Environmental Quality, 32: 1072-1081. DOI: 10.2134/jeq2003.1072.

Kleinman PJA, Sharpley AN, Moyer BG, Elwinger GF. 2002. Effect of mineral and manure phosphorus

sources on runoff phosphorus. Journal of Environmental Quality, 31: 2026-2033. DOI: 10.2134/

jeq2002.2026.

Kleinman PJA, Sharpley AN, Saporito LS, Buda AR, Bryant RB. 2009. Application of manure to no-till soils:

phosphorus losses by sub-surface and surface pathways. Nutrient Cycling in Agroecosystems, 84: 215-227.

DOI: 10.1007/s10705-008-9238-3.

Klute A. 1982. Tillage effects on the hydraulic properties of soil: A review. In: Predicting Tillage Effects on

Soil Physical Properties and Processes, Unger PW, Van Doren Jr DM (eds.) American Society of

Agronomy and Soil Science Society of America, pp: 29-43.

Koiter AJ, Lobb DA, Owens PN, Petticrew EL, Tiessen KHD, Li S. 2013. Investigating the role of connectivity

and scale in assessing the sources of sediment in an agricultural watershed in the Canadian prairies using

sediment source fingerprinting. Journal of Soils and Sediments, 13: 1676-1691. DOI: 10.1007/s11368-013-

0762-7.

Kovacic DA, David MB, Gentry LE, Starks KM, Cooke RA. 2000. Effectiveness of constructed wetlands in

reducing nitrogen and phosphorus export from agricultural tile drainage. Journal of Environmental Quality,

29: 1262-1274. DOI: 10.2134/jeq2000.00472425002900040033x.

Kumaragamage D, Flaten DN, Akinremi OO, Sawka CA, Ige D, Zvomuya F. 2012. Impact of manure

phosphorus fractions on phosphorus loss from manured soils after incubation. Journal of Environmental

Quality, 41: 845-854. DOI: 10.2134/jeq2011.0415.

90

La Salle Redboine Conservation District. 2007. La Salle River Watershed - State of the Watershed Report. pp:

296.

Lamer M. 1957. The World Fertilizer Economy: Stanford University Press.

Lapen DR, Topp E, Edwards M, Sabourin L, Curnoe W, Gottschall N, Bolton P, Rahman S, Ball-Coelho B,

Payne M, Kleywegt S, McLaughlin N. 2008. Effect of liquid municipal biosolid application method on tile

and ground water quality. Journal of Environmental Quality, 37: 925-936. DOI: 10.2134/jeq2006.0486.

Larney FJ, Kladivko EJ, Monke EJ. 1988. Subsurface drain spacing effects on soil moisture regime of clermont

silt loam. Transactions of the American Society of Agricultural Engineers, 31: 1128-1134.

Lazarus WF, Mulla DJ, Wall D. 2014. A spreadsheet planning tool for assisting a state agency with cost-

effective watershed scale surface water nitrogen planning. Journal of Soil and Water Conservation, 69:

45A-50A. DOI: 10.2489/jswc.69.2.45A.

Lemaire G, Gastal F. 1997. N uptake and distribution in plant canopies. In: Diagnosis of the nitrogen status in

crops, Lemaire G (ed.) Springer, pp: 3-43.

Liu K, Elliott JA, Lobb DA, Flaten DN, Yarotski J. 2013a. Critical factors affecting field-scale losses of

nitrogen and phosphorus in spring snowmelt runoff in the canadian prairies. Journal of Environmental

Quality, 42: 484-496. DOI: 10.2134/jeq2012.0385.

Liu K, Elliott JA, Lobb DA, Flaten DN, Yarotski J. 2013b. Critical Factors Affecting Field-Scale Losses of

Nitrogen and Phosphorus in Spring Snowmelt Runoff in the Canadian Prairies. Journal of Environmental

Quality, 42: 484-496. DOI: 10.2134/jeq2012.0385.

Liu K, Elliott JA, Lobb DA, Flaten DN, Yarotski J. 2014. Nutrient and sediment losses in snowmelt runoff

from perennial forage and annual cropland in the Canadian Prairies. Journal of Environmental Quality, 43:

1644-1655. DOI: 10.2134/jeq2014.01.0040.

Lobb D. 2012. Determining the effective use of riparian areas to filter sediments and phosphorus. University of

Manitoba. Winnipeg, MB.

Lobb D. 2015. Getting the most out of your land for forages and grasses. Canadian Forage and Grassland

Association, CFGA. Annual Meeting. November 18, 2015. Saskatoon, SK., pp: 32 pages.

Logan TJ, Eckert DJ, Beak DG. 1994. Tillage, crop and climatic effects of runoff and tile drainage losses of

nitrate and four herbicides. Soil and Tillage Research, 30: 75-103. DOI: http://dx.doi.org/10.1016/0167-

1987(94)90151-1.

Luo W, Skaggs RW, Chescheir GM. 2000. DRAINMOD modifications for cold conditions. Transactions of the

American Society of Agricultural Engineers, 43: 1569-1582. DOI: 10.13031/2013.3057.

Lyseng R. (2004). Tile drainage pays dividends. Retrieved January 26th, 2016, from GVIC Communications

http://www.producer.com/2005/12/tile-drainage-pays-dividends/

Ma L, Malone RW, Heilman P, Jaynes DB, Ahuja LR, Saseendran SA, Kanwar RS, Ascough Ii JC. 2007.

RZWQM simulated effects of crop rotation, tillage, and controlled drainage on crop yield and nitrate-N loss

in drain flow. Geoderma, 140: 260-271. DOI: http://dx.doi.org/10.1016/j.geoderma.2007.04.010.

Madramootoo CA, Wiyo KA, Enright P. 1992. Nutrient losses through tile drains from two potato fields.

Applied Engineering in Agriculture, 8: 639-646. DOI: 10.13031/2013.26136.

MAFRD. 2015. Soil Management Guide. Retrieved 2016/02/16, from https://www.gov.mb.ca/agriculture/

environment/soil-management/soil-management-guide/drainage-management.html

Maguire RO, Kleinman PJA, Dell CJ, Beegle DB, Brandt RC, McGrath JM, Ketterings QM. 2011. Manure

application technology in reduced tillage and forage systems: A review. Journal of Environmental Quality,

40: 292-301. DOI: 10.2134/jeq2009.0228.

Manitoba Agricultural Services Corporation. 2016. Manitoba Management Plus Program: Seeding dates and

91

yield information. Retrieved 2016/02/29, from http://www.mmpp.com/mmpp.nsf/

mmpp_seeding_dates.html

Manitoba Agricultural Sustainability Initiative. 2007. Soil Fertility Guide. Manitoba Agriculture, Food and

Rural Development.

Manitoba Conservation and Water Stewardship. 2014. Manitoba's Surface Water Management Strategy.

Manitoba Conservation and Water Stewardship, pp: 28.

Manitoba Mineral Resources. 2016. The Quaternary Landscape of Manitoba. Retrieved May 21, 2016, from

http://www.manitoba.ca/iem/geo/gis/sgcmsdesc.html

Marshall TJ, Holmes JW, Rose CW. 1996. Soil Physics. New York, NY: Cambridge University Press.

MAWMA. 2015. Manitoba Agricultural Water Management Association - Advocacy. Retrieved May 30, 2016,

from http://www.mawma.org/

McCullough GK, Page SJ, Hesslein RH, Stainton MP, Kling HJ, Salki AG, Barber DG. 2012. Hydrological

forcing of a recent trophic surge in Lake Winnipeg. Journal of Great Lakes Research, 38, Supplement 3: 95-

105. DOI: http://dx.doi.org/10.1016/j.jglr.2011.12.012.

McDonnell JJ. 2003. Where does water go when it rains? Moving beyond the variable source area concept of

rainfall-runoff response. Hydrological Processes, 17: 1869-1875. DOI: 10.1002/hyp.5132.

McDonnell JJ. 2013. Are all runoff processes the same? Hydrological Processes, 27: 4103-4111. DOI:

10.1002/hyp.10076.

Mikkelsen RL. 2011. The “4R” nutrient stewardship framework for horticulture. HortTechnology, 21: 658-662.

Milburn P, Richards JE, Gartley C, Pollock T, O'Neill H, Bailey H. 1990. Nitrate leaching from systematically

tiled potato fields in New Brunswick, canada. Journal of Environmental Quality, 19. DOI: 10.2134/jeq1990.

00472425001900030016x.

Mishra SK, Tyagi JV, Singh VP. 2003. Comparison of infiltration models. Hydrological Processes, 17: 2629-

2652. DOI: 10.1002/hyp.1257.

Molnar JL, Kubiszewski I. 2012. Managing natural wealth: Research and implementation of ecosystem

services in the United States and Canada. Ecosystem Services, 2: 45-55. DOI: http://dx.doi.org/

10.1016/j.ecoser.2012.09.005.

Mongillo JF, Zierdt-Warshaw L. 2000. Encyclopedia of Environmental Science: Oryx Press.

Mooleki SP, Schoenau JJ, Hultgreen G, Wen G, Charles JL. 2002. Effect of rate, frequency and method of

liquid swine manure application on soil nitrogen availability, cropperformance and N use efficiency in east-

central Saskatchewan. Canadian Journal of Soil Science, 82: 457-467. DOI: 10.4141/S01-057.

Moore PA, Daniel TC, Sharpley AN, Wood CW. 1995. Poultry manure management: Environmentally sound

options. Journal of Soil and Water Conservation, 50: 321-327.

Nguyen HQ, Kanwar RS, Hoover NL, Dixon P, Hobbs J, Pederson C, Soupir ML. 2013. Long-term effects of

poultry manure application on nitrate leaching in tile drain water. 56. DOI: 10.13031/2013.42593.

Nichols DJ, Daniel TC, Edwards DR. 1994. Nutrient runoff from pasture after incorporation of poultry litter or

inorganic fertilizer. Soil Science Society of America Journal, 58: 1224-1228. DOI: 10.2136/sssaj1994.

03615995005800040032x.

Nikièma P, Buckley KE, Enns JM, Qiang H, Akinremi OO. 2013. Effects of liquid hog manure on soil

available nitrogen status, nitrogen leaching losses and wheat yield on a sandy loam soil of western Canada.

Canadian Journal of Soil Science, 93: 573-584. DOI: 10.4141/cjss2012-070.

Nolte BH, Fausey NR, Skaggs RW. 1983. Time available for field work on two Ohio soils. Transactions of the

American Society of Agricultural Engineers, 26: 445-451. DOI: 10.13031/2013.33956.

92

Olson RA, Dreier AF. 1956. Fertilizer placement for small grains in relation to crop stand and nutrient

efficiency in Nebraska. Soil Science Society of America Journal, 20: 19-24. DOI: 10.2136/sssaj1956.

03615995002000010005x.

Osborne LL, Kovacic DA. 1993. Riparian vegetated buffer strips in water-quality restoration and stream

management. Freshwater Biology, 29: 243-258. DOI: 10.1111/j.1365-2427.1993.tb00761.x.

Osei E, Gassman PW, Hauck LM, Jones R, Beran L, Dyke PT, Goss DW, Flowers JD, McFarland AMS, Saleh

A. 2003. Environmental benefits and economic costs of manure incorporation on dairy waste application

fields. Journal of Environmental Management, 68: 1-11. DOI: http://dx.doi.org/10.1016/S0301-

4797(02)00226-8.

Øygarden L, Kværner J, Jenssen PD. 1997. Soil erosion via preferential flow to drainage systems in clay soils.

Geoderma, 76: 65-86. DOI: http://dx.doi.org/10.1016/S0016-7061(96)00099-7.

Ozkan HE, He B, Holmes RG. 1991. Estimating effects of subsurface drainage and machinery size on returns.

Transactions of the American Society of Agricultural Engineers, 34: 1935-1940. DOI: 10.13031/

2013.31819.

Pachepsky YA, Sadeghi AM, Bradford SA, Shelton DR, Guber AK, Dao T. 2006. Transport and fate of manure-

borne pathogens: Modeling perspective. Agricultural Water Management, 86: 81-92. DOI: http://dx.doi.

org/10.1016/j.agwat.2006.06.010.

Pappas EA, Kanwar RS, Baker JL, Lorimor JC, Mickelson S. 2008. Fecal indicator bacteria in subsurface drain

water following swine manure application. Transactions of the American Society of Agricultural Engineers,

51: 1567-1573. DOI: 10.13031/2013.25313.

Partheeban C, Kjaersgaard J, Hay C, Trooien T. 2014. Demonstrating the nitrogen-removal effectiveness of

denitrifying bioreactors for improved drainage water management in South Dakota. ASABE, pp: 1.

Patni NK, Masse L, Jui PY. 1996. Tile effluent quality and chemical losses under conventional and no tillage

— part 1: Flow and nitrate. Transactions of the American Society of Agricultural Engineers, 39: 1665-1672.

DOI: 10.13031/2013.27683.

Patni NK, Toxopeus R, Tennant AD, Hore FR. 1984. Bacterial quality of tile drainage water from manured and

fertilized cropland. Water Research, 18: 127-132. DOI: http://dx.doi.org/10.1016/0043-1354(84)90059-9.

Petrolia DR, Gowda PH. 2006. Missing the boat: Midwest farm drainage and Gulf of Mexico hypoxia. Applied

Economic Perspectives and Policy, 28: 240-253. DOI: 10.1111/j.1467-9353.2006.00284.x.

Phillips PA, Culley JLB, Hore FR, Patni NK. 1981. Pollution potential and corn yields from selected rates and

timing of liquid manure applications. Transactions of the American Society of Agricultural Engineers, 24.

DOI: 10.13031/2013.34213.

Pomeroy JW, Gray DM, Brown T, Hedstrom NR, Quinton WL, Granger RJ, Carey SK. 2007. The cold regions

hydrological model: a platform for basing process representation and model structure on physical evidence.

Hydrological Processes, 21: 2650-2667. DOI: 10.1002/hyp.6787.

Potter P, Ramankutty N, Bennett EM, Donner SD. 2010. Characterizing the spatial patterns of global fertilizer

application and manure production. Earth Interactions, 14: 1-22. DOI: 10.1175/2009EI288.1.

Power JF, Schepers JS. 1989. Effects of agriculture on groundwater nitrate contamination of groundwater in

North America. Agriculture, Ecosystems & Environment, 26: 165-187. DOI: http://dx.doi.org/10.1016/0167-

8809(89)90012-1.

Rajsic P, Weersink A. 2008. Do farmers waste fertilizer? A comparison of ex post optimal nitrogen rates and ex

ante recommendations by model, site and year. Agricultural Systems, 97: 56-67. DOI: http://dx.doi.org/

10.1016/j.agsy.2007.12.001.

Randall GW, Goss MJ. 2008. Nitrate losses to surface water through subsurface, tile drainage. In: Nitrogen in

the Environment, Hatfield JL, Follett RF (eds.) Elsevier Science, pp: 145-176.

93

Randall GW, Hoeft RG. 1988. Placement methods for improved efficiency of p and k fertilizers: A review.

Journal of Production Agriculture, 1: 70-79. DOI: 10.2134/jpa1988.0070.

Randall GW, Iragavarapu TK, Schmitt MA. 2000. Nutrient losses in subsurface drainage water from dairy

manure and urea applied for corn. Journal of Environmental Quality, 29: 1244-1252. DOI:

10.2134/jeq2000.00472425002900040031x.

Randall GW, Mulla DJ. 2001. Nitrate nitrogen in surface waters as influenced by climatic conditions and

agricultural practices. Journal of Environmental Quality, 30. DOI: 10.2134/jeq2001.302337x.

Raney W, Zingg A. 1957. Principles of tillage. USDA. Yearbook of Agriculture. Washington: 277-281.

Rasmussen K. 2002. Influence of liquid manure application method on weed control in spring cereals. Weed

Research, 42: 287-298. DOI: 10.1046/j.1365-3180.2002.00286.x.

Rasouli S, Whalen JK, Madramootoo CA. 2014. Review: Reducing residual soil nitrogen losses from

agroecosystems for surface water protection in Quebec and Ontario, Canada: Best management practices,

policies and perspectives. Canadian Journal of Soil Science, 94: 109-127. DOI: 10.4141/cjss2013-015.

Reddy KR, Khaleel R, Overcash MR. 1981. Behavior and transport of microbial pathogens and indicator

organisms in soils treated with organic wastes. Journal of Environmental Quality, 10: 255-266. DOI:

10.2134/jeq1981.00472425001000030001x.

Rees RM, Roelcke M, Li SX, Wang XQ, Li SQ, Stockdale EA, McTaggart IP, Smith KA, Richter J. 1997. The

effect of fertilizer placement on nitrogen uptake and yield of wheat and maize in Chinese loess soils.

Nutrient Cycling in Agroecosystems, 47: 81-91. DOI: 10.1007/bf01985721.

Reid LM, Nicol RW, Ouellet T, Savard M, Miller JD, Young JC, Stewart DW, Schaafsma AW. 1999. Interaction

of Fusarium graminearum and F. Moniliforme in maize ears: Disease progress, fungal biomass, and

mycotoxin accumulation. Phytopathology, 89: 1028-1037. DOI: 10.1094/PHYTO.1999.89.11.1028.

Reidy B, Ruettimann L, Menzi H. 2005. DYNAMO: An empirical model for the calculation of ammonia

emissions of farms and regions. In: Emissions from European Agriculture, Kuczyński; T, Dämmgen; U,

Webb; J, Myczko A (eds.) Wageningen Academic Publishers, pp: 125-128.

Richardson CJ, Flanagan NE, Ho M, Pahl JW. 2011. Integrated stream and wetland restoration: A watershed

approach to improved water quality on the landscape. Ecological Engineering, 37: 25-39. DOI:

http://dx.doi.org/10.1016/j.ecoleng.2010.09.005.

Riddle MU, Bergstrőm L. 2013. Phosphorus Leaching from Two Soils with Catch Crops Exposed to Freeze–

Thaw Cycles. Agronomy Journal, 105: 803-811. DOI: 10.2134/agronj2012.0052.

Robertson WD, Ptacek CJ, Brown SJ. 2009. Rates of nitrate and perchlorate removal in a 5-year-old wood

particle reactor treating agricultural drainage. Ground Water Monitoring & Remediation, 29: 87-94. DOI:

10.1111/j.1745-6592.2009.01231.x.

Rowe RC, Nielsen IW. 1981. Pink rot. In: Compendium of Potato Diseases, Hooker WJ (ed.) American

Phytopathological Society, pp: 39-40.

Royer TV, David MB, Gentry LE. 2006. Timing of riverine export of nitrate and phosphorus from agricultural

watersheds in Illinois:  Implications for reducing nutrient loading to the Mississippi River. Environmental

Science & Technology, 40: 4126-4131. DOI: 10.1021/es052573n.

Salo MW, Hua G, Schmit C, Hay C. 2015. Nutrient removal from agricultural subsurface drainage using

denitrification bioreactors and phosphate adsorbents. ASABE, pp: 1.

Sarrantonio M, Gallandt E. 2003. The role of cover crops in north american cropping systems. Journal of Crop

Production, 8: 53-74. DOI: 10.1300/J144v08n01_04.

Satchithanantham S, Sri Ranjan R. 2015. Evaluation of DRAINMOD for potato crop under cold conditions in

the Canadian Prairies. Transactions of the American Society of Agricultural Engineers, 58: 307-317. DOI:

94

10.13031/trans.58.10300.

Satchithanantham S, Sri Ranjan R, Bullock P. 2014. Protecting water quality using controlled drainage as an

agricultural BMP for potato production. Transactions of the American Society of Agricultural Engineers, 57:

815-826. DOI: 10.13031/trans.57.10385.

Satchithanantham S, Sri Ranjan R, Shewfelt B. 2012. Effect of water table management and irrigation on

potato yield. Transactions of the American Society of Agricultural Engineers, 55: 2175-2184. DOI:

10.13031/2013.42509.

Schafer CE. 2011. Groundwater control system with automatic water flow regulator. US Patent Classification

405/39; 405/96.

Schilling KE, Helmers M. 2008. Effects of subsurface drainage tiles on streamflow in Iowa agricultural

watersheds: Exploratory hydrograph analysis. Hydrological Processes, 22: 4497-4506. DOI: 10.1002/

hyp.7052.

Schilling KE, Libra RD. 2000. The relationship of nitrate concentrations in streams to row crop land use in

Iowa. Journal of Environmental Quality, 29. DOI: 10.2134/jeq2000.00472425002900060016x.

Schindler DW. 1977. Evolution of phosphorus limitation in lakes. Science, 195: 260-262.

Schindler DW, Hecky RE, McCullough GK. 2012. The rapid eutrophication of Lake Winnipeg: Greening under

global change. Journal of Great Lakes Research, 38: 6-13. DOI: 10.1016/j.jglr.2012.04.003.

Schmidt K, Sherman I. 1987. Effect of irrigation on groundwater quality in California. Journal of Irrigation

and Drainage Engineering, 113: 16-29. DOI: doi:10.1061/(ASCE)0733-9437(1987)113:1(16).

Schmitt MA, Evans SD, Randall GW. 1995. Effect of liquid manure application methods on soil nitrogen and

corn grain yields. Journal of Production Agriculture, 8: 186-189. DOI: 10.2134/jpa1995.0186.

Schmitt MA, Russelle MP, Randall GW, Sheaffer CC, Greub LJ, Clayton PD. 1999. Effect of rate, timing, and

placement of liquid dairy manure on reed canarygrass yield. Journal of Production Agriculture, 12: 239-

243. DOI: 10.2134/jpa1999.0239.

Schroeder PD, Radcliffe DE, Cabrera ML. 2004. Rainfall timing and poultry litter application rate effects on

phosphorus loss in surface runoff. Journal of Environmental Quality, 33. DOI: 10.2134/jeq2004.2201.

Scott CA, Geohring LD, Walter MF. 1998. Water quality impacts of tile drains in shallow, sloping, structured

soils as affected by manure application. Applied Engineering in Agriculture, 14: 599-603. DOI: 10.13031/

2013.19428.

Selezneva OI, Jiang YJ, Larson G, Puzin T. 2008. Long-Term Pavement Performance Computed Parameter:

Frost Penetration. Infrastructure Research and Development, Federal Highway Administration.

Sharpley A, Moyer B. 2000. Phosphorus forms in manure and compost and their release during simulated

rainfall. Journal of Environmental Quality, 29: 1462-1469. DOI: 10.2134/jeq2000.

00472425002900050012x.

Sharpley AN. 1997. Rainfall frequency and nitrogen and phosphorus runoff from soil amended with poultry

litter. Journal of Environmental Quality, 26. DOI: 10.2134/jeq1997.00472425002600040026x.

Sharpley AN, Chapra SC, Wedepohl R, Sims JT, Daniel TC, Reddy KR. 1994. Managing agricultural

phosphorus for protection of surface waters: Issues and options. Journal of Environmental Quality, 23: 437-

451. DOI: 10.2134/jeq1994.00472425002300030006x.

Sharpley AN, Kleinman PJA, McDowell RW, Gitau M, Bryant RB. 2002. Modeling phosphorus transport in

agricultural watersheds: Processes and possibilities. Journal of Soil and Water Conservation, 57: 425-439.

Sharpley AN, Smith SJ, Jones OR, Berg WA, Coleman GA. 1992. The transport of bioavailable phosphorus in

agricultural runoff. Journal of Environmental Quality, 21: 30-35. DOI: 10.2134/jeq1992.

00472425002100010003x.

95

Sheppard SC, Sheppard MI, Long J, Sanipelli B, Tait J. 2006. Runoff phosphorus retention in vegetated field

margins on flat landscapes. Canadian Journal of Soil Science, 86: 871-884. DOI: 10.4141/S05-072.

Sheriff G. 2005. Efficient waste? Why farmers over-apply nutrients and the implications for policy design.

Applied Economic Perspectives and Policy, 27: 542-557. DOI: 10.1111/j.1467-9353.2005.00263.x.

Shipitalo MJ, Gibbs F. 2000. Potential of earthworm burrows to transmit injected animal wastes to tile drains.

Soil Science Society of America Journal, 64: 2103-2109. DOI: 10.2136/sssaj2000.6462103x.

Shook K, Pomeroy J. 2012. Changes in the hydrological character of rainfall on the Canadian prairies.

Hydrological Processes, 26: 1752-1766. DOI: 10.1002/hyp.9383.

Shook KR, Pomeroy JW. 2010. Hydrological effects of the temporal variability of the multiscaling of snowfall

on the Canadian prairies. Hydrol. Earth Syst. Sci., 14: 1195-1203. DOI: 10.5194/hess-14-1195-2010.

Simard R, Beauchemin S, Haygarth P. 2000. Potential for preferential pathways of phosphorus transport.

Journal of Environmental Quality, 29: 97-105.

Sims JT, Bergström L, Bowman BT, Oenema O. 2005. Nutrient management for intensive animal agriculture:

policies and practices for sustainability. Soil Use and Management, 21: 141-151. DOI: 10.1111/j.1475-

2743.2005.tb00118.x.

Šimůnek J, Jarvis NJ, van Genuchten MT, Gärdenäs A. 2003. Review and comparison of models for describing

non-equilibrium and preferential flow and transport in the vadose zone. Journal of Hydrology, 272: 14-35.

DOI: http://dx.doi.org/10.1016/S0022-1694(02)00252-4.

Singh AK, Madramootoo CA, Smith DL. 2014. Impact of different water management scenarios on corn water

use efficiency. Transactions of the American Society of Agricultural Engineers, 57: 1319-1328. DOI:

10.13031/trans.57.10005.

Skaggs RW, Brevé MA, Gilliam JW. 1994. Hydrologic and water quality impacts of agricultural drainage.

Critical Reviews in Environmental Science and Technology, 24: 1-32. DOI: 10.1080/10643389409388459.

Skaggs RW, Miller DE, Brooks RH. 1980. Soil water Part I - Properties. In: Design and operation of farm

irrigation systems, Jensen ME (ed.) American Society of Agricultural Engineers, pp: 75-123.

Skaggs RW, Nassehzadeh-Tabriz A. 1986. Design drainage rates for estimating drain spacings in North

Carolina. Transactions of the ASAE, 29: 1631-1640. DOI: 10.13031/2013.30364.

Skaggs RW, Youssef MA, Chescheir GM. 2012. DRAINMOD: Model use, calibration, and validation.

Transactions of the American Society of Agricultural Engineers, 55: 1509-1522.

Smedema LK, Vlotman WF, Rycroft D. 2004. Modern Land Drainage: Planning, Design and Management of

Agricultural Drainage Systems: Taylor & Francis.

Smith DR, Owens PR, Leytem AB, Warnemuende EA. 2007. Nutrient losses from manure and fertilizer

applications as impacted by time to first runoff event. Environmental Pollution, 147: 131-137. DOI:

http://dx.doi.org/10.1016/j.envpol.2006.08.021.

Snapp SS, Swinton SM, Labarta R, Mutch D, Black JR, Leep R, Nyiraneza J, O'Neil K. 2005. Evaluating cover

crops for benefits, costs and performance within cropping system niches. Agronomy Journal, 97: 322-332.

DOI: 10.2134/agronj2005.0322.

Sommer SG, Jacobsen OH. 1999. Infiltration of slurry liquid and volatilization of ammonia from surface

applied pig slurry as affected by soil water content. The Journal of Agricultural Science, 132: 297-303.

DOI: doi:null.

Spence C. 2010. A paradigm shift in hydrology: Storage thresholds across scales influence catchment runoff

generation. Geography Compass, 4: 819-833. DOI: 10.1111/j.1749-8198.2010.00341.x.

Stacey EF, Jeffrey SS, Gary RS, Adam SB. 2010. Controlled drainage to improve edge-of-field water quality in

southwest Minnesota, USA. ASABE, pp: 1-10.

96

Stampfli N, Madramootoo CA. 2006. Dissolved phosphorus losses in tile drainage under subirrigation. Water

Quality Research Journal of Canada, 41: 63-71.

Stewart WM, Dibb DW, Johnston AE, Smyth TJ. 2005. The contribution of commercial fertilizer nutrients to

food production. Agronomy Journal, 97. DOI: 10.2134/agronj2005.0001.

Stillman JS, Haws NW, Govindaraju RS, Suresh C. Rao P. 2006. A semi-analytical model for transient flow to

a subsurface tile drain. Journal of Hydrology, 317: 49-62. DOI: http://dx.doi.org/10.1016/j.jhydrol.

2005.04.028.

Stonehouse DP. 1995. Profitability of soil and water conservation in Canada: A review. Journal of Soil and

Water Conservation, 50: 215-219.

Stratton GW, Madani A, Gordon RJ, Sharples K, Coulter T, and AT. 2004. Bacterial quality of surface and

subsurface drainage water from manured fields. In: Eighth International Draiange Symposium,

(Sacramento, California USA), Publication Date 21 March 2004., pp: 197-205, St. Joseph, Mich., 21-24

March 2004.

Strock JS, Porter PM, Russelle MP. 2004a. Cover cropping to reduce nitrate loss through subsurface drainage

in the northern U.S. Corn belt. Journal of Environmental Quality, 33: 1010-1016. DOI: 10.2134/jeq2004.

1010.

Strock JS, Porter PM, Russelle MP. 2004b. Cover cropping to reduce nitrate loss through subsurface drainage

in the northern U.S. Corn Belt. Journal of Environmental Quality, 33. DOI: 10.2134/jeq2004.1010.

Stumborg C, Schoenau JJ, Malhi SS. 2007. Nitrogen balance and accumulation pattern in three contrasting

prairie soils receiving repeated applications of liquid swine and solid cattle manure. Nutrient Cycling in

Agroecosystems, 78: 15-25. DOI: 10.1007/s10705-006-9071-5.

Tan CS, Drury CF, Gaynor JD, Ng HYF. 1999. Effect of controlled drainage and subirrigation on subsurface

tile drainage nitrate loss and crop yield at the farm scale. Canadian Water Resources Journal, 24: 177-186.

DOI: 10.4296/cwrj2403177.

Tan CS, Drury CF, Gaynor JD, Welacky TW, Reynolds WD. 2002. Effect of tillage and water table control on

evapotranspiration, surface runoff, tile drainage and soil water content under maize on a clay loam soil.

Agricultural Water Management, 54: 173-188. DOI: http://dx.doi.org/10.1016/S0378-3774(01)00178-0.

Tan CS, Drury CF, Soultani M, van Wesenbeeck IJ, Ng HYF, Gaynor JD, Welacky TW. 1998. Effect of

controlled drainage and tillage on soil structure and tile drainage nitrate loss at the field scale. Water Science

and Technology, 38: 103-110. DOI: http://dx.doi.org/10.1016/S0273-1223(98)00503-4.

Tan CS, Zhang TQ, Drury CF, Reynolds WD, Oloya T, Gaynor JD. 2007. Water quality and crop production

improvement using a wetland-reservoir and draining/subsurface irrigation system. Canadian Water

Resources Journal, 32: 129-136. DOI: 10.4296/cwrj3202129.

Tan CS, Zhang TQ, Drury CF, Welacky TW, Reynolds WD. 2009. Impacts of liquid and solid manures under

controlled drainage with sub-irrigation recycling systems on water quality and crop production. ASABE, pp:

1-11.

Tan CS, Zhang TQ, Zheng ZM. 2015. Impact of subsurface drainage management and organic manure and

chemical fertilizer on nutrient loss. ASABE, pp: 1.

The Prairie Provinces’ Committee on Livestock Development and Manure Management. 2004. Tri-Provincial

Manure Application and Use Guidelines - Manitoba Version. Alberta Agriculture, Food and Rural

Development/Saskatchewan Agriculture, Food and Rural Revitalization/Manitoba Agriculture, Food and

Rural Initiatives.

Thompson DJ, Clark KW. 1989. Influence of nitrogen fertilization and mechanical stubble removal on seed

production of kentucky bluegrass in Manitoba. Canadian Journal of Plant Science, 69: 939-943. DOI:

10.4141/cjps89-114.

97

Tiessen KHD, Elliott JA, Stainton M, Yarotski J, Flaten DN, Lobb DA. 2011. The effectiveness of small-scale

headwater storage dams and reservoirs on stream water quality and quantity in the Canadian Prairies.

Journal of Soil and Water Conservation, 66: 158-171. DOI: 10.2489/jswc.66.3.158.

Tiessen KHD, Elliott JA, Yarotski J, Lobb DA, Flaten DN, Glozier NE. 2010. Conventional and conservation

tillage: influence on seasonal runoff, sediment, and nutrient losses in the Canadian Prairies. Journal of

Environmental Quality, 39: 964-980. DOI: 10.2134/jeq2009.0219.

Tilman D. 1999. Global environmental impacts of agricultural expansion: The need for sustainable and

efficient practices. Proceedings of the National Academy of Sciences, 96: 5995-6000. DOI: 10.1073/

pnas.96.11.5995.

Tilman D, Cassman KG, Matson PA, Naylor R, Polasky S. 2002. Agricultural sustainability and intensive

production practices. Nature, 418: 671-677.

Tomar JS, Soper RJ. 1981. Fate of tagged urea n in the field with different methods of N and organic matter

placement. Agronomy Journal, 73: 991-995. DOI: 10.2134/agronj1981.00021962007300060020x.

Turtola E, Jaakkola A. 1995. Loss of phosphorus by surface runoff and leaching from a heavy clay soil under

barley and grass ley in Finland. Acta Agriculturae Scandinavica, Section B — Soil & Plant Science, 45:

159-165. DOI: 10.1080/09064719509413099.

Unc A, Goss MJ. 2004. Transport of bacteria from manure and protection of water resources. Applied Soil

Ecology, 25: 1-18. DOI: http://dx.doi.org/10.1016/j.apsoil.2003.08.007.

USBR. 1993. Drainage Manual 3rd. edition ed. Denver, CO.: United States Department of Interior, Bureau of

Reclamation.

USDA-NRCS. 2001. Water management (Draiange). In: National Engineering Handbook, Part 630, U.S.

Department of Agriculture, Soil Conservation Service, pp: 14.11-14.192.

Uusi-Kämppä J. 2005. Phosphorus purification in buffer zones in cold climates. Ecological Engineering, 24:

491-502. DOI: http://dx.doi.org/10.1016/j.ecoleng.2005.01.013.

Vadas PA, Kleinman PJA, Sharpley AN. 2004. A simple method to predict dissolved phosphorus in runoff from

surface-applied manures. Journal of Environmental Quality, 33: 749-756. DOI: 10.2134/jeq2004.7490.

Valero CS, Madramootoo CA, Stämpfli N. 2007. Water table management impacts on phosphorus loads in tile

drainage. Agricultural Water Management, 89: 71-80. DOI: http://dx.doi.org/10.1016/j.agwat.2006.12.007.

van der Kamp G, Hayashi M, Gall�n D. 2003. Comparing the hydrology of grassed and cultivated catchments

in the semi-arid Canadian prairies. Hydrological Processes, 17: 559-575. DOI: 10.1002/hyp.1157.

van Driel PW, Robertson WD, Merkley LC. 2006. Denitrification of agricultural drainage using wood-based

reactors. Transactions of the American Society of Agricultural Engineers, 49: 565-573. DOI: 10.13031/

2013.20391.

van Es HM, Schindelbeck RR, Jokela WE. 2004. Effect of manure application timing, crop, and soil type on

phosphorus leaching. Journal of Environmental Quality, 33: 1070-1080. DOI: 10.2134/jeq2004.1070.

van Es HM, Sogbedji JM, Schindelbeck RR. 2006. Effect of manure application timing, crop, and soil type on

nitrate leaching. Journal of Environmental Quality, 35: 670-679. DOI: 10.2134/jeq2005.0143.

Van Esbroeck CJ, Macrae ML, Brunke RI, McKague K. 2016. Annual and seasonal phosphorus export in

surface runoff and tile drainage from agricultural fields with cold temperate climates. Journal of Great

Lakes Research, In Press. DOI: http://dx.doi.org/10.1016/j.jglr.2015.12.014.

Van Schilfgaarde J. 1974. Introduction. In: Drainage for Agriculture, Van Schilfgaarde J (ed.) American

Society of Agronomy, pp: 3-4.

Vigier B, Reid LM, Seifert KA, Stewart DW, Hamilton RI. 1997. Distribution and prediction of Fusarium

species associated with maize ear rot in Ontario. Canadian Journal of Plant Pathology, 19: 60-65. DOI:

98

10.1080/07060669709500574.

Volf CA, Ontkean GR, Bennett DR, Chanasyk DS, Miller JJ. 2007. Phosphorus losses in simulated rainfall

runoff from manured soils of Alberta. Journal of Environmental Quality, 36: 730-741. DOI:

10.2134/jeq2006.0269.

Walter MF, Steenhuis TS, Haith DA. 1979. Nonpoint source pollution control by soil and water conservation

practices. Transactions of the ASABE, 22: 834-840. DOI: 10.13031/2013.35110.

Wang L, Zhu J, Miller C. 2010. Shortcut nitrification and denitrification of swine wastewater in a sequencing

batch reactor system. Transactions of the American Society of Agricultural Engineers, 53: 813-818. DOI:

10.13031/2013.30064.

Watson JT, Reed SC, Kadlec RH, Knight RL, Whitehouse AE. 1989. Performance expectations and loading

rates for constructed wetlands. In: Constructed Wetlands for Wastewater Treatment: Municipal, Industrial

and Agricultural., Hammer DA (ed.) Lewis Publishers, pp: 319-351.

Wesseling J. 1974. Crop Growth and Wet Soils. In: Drainage for Agriculture, Van Schilfgaarde J (ed.)

American Society of Agronomy.

Willardson LS. 1974. Envelope Materials. In: Drainage for Agriculture, Van Schilfgaarde J (ed.) American

Society of Agronomy.

Wolfe ML. 2000. Hydrology. In: Agricultural Nonpoint Source Pollution, CRC Press.

Woltemade CJ. 2000. Ability of restored wetlands to reduce nitrogen and phosphorus concentrations in

agricultural drainage water. Journal of Soil and Water Conservation, 55: 303-309.

Xu XM, Monger W, Ritieni A, Nicholson P. 2007. Effect of temperature and duration of wetness during initial

infection periods on disease development, fungal biomass and mycotoxin concentrations on wheat

inoculated with single, or combinations of, Fusarium species. Plant Pathology, 56: 943-956. DOI:

10.1111/j.1365-3059.2007.01650.x.

Young TC, Ligon JT. 1972. Water table and soil moisture probabilities with tile drainage. Transactions of the

ASABE, 15: 448-451. DOI: 10.13031/2013.37924.

Zhang TQ, Tan CS, Zheng ZM, Drury CF. 2015. Tile drainage phosphorus loss with long-term consistent

cropping systems and fertilization. Journal of Environmental Quality, 44: 503-511. DOI: 10.2134/jeq2014.

04.0188.

Zhao L, Gray DM. 1999. Estimating snowmelt infiltration into frozen soils. Hydrological Processes, 13: 1827-

1842. DOI: 10.1002/(SICI)1099-1085(199909)13:12/13<1827::AID-HYP896>3.0.CO;2-D.

Zhao SL, Gupta SC, Huggins DR, Moncrief JF. 2001. Tillage and nutrient source effects on surface and

subsurface water quality at corn planting. Journal of Environmental Quality, 30: 998-1008. DOI: 10.2134/

jeq2001.303998x.