Chemical Fractions and Prediction for Long-term Releases ...

Chemical Fractions and Prediction for Long-term Releases of

Phosphorus in Typical Canadian Agricultural Soils

by

Aruna Kanthi Withana Herath

A Thesis

presented to

The University of Guelph

In partial fulfillment of the requirements

for the degree of

Doctor of Philosophy

in

Environmental Sciences

Guelph, Ontario, Canada

© Aruna Withana Herath, April, 2013

ABSTRACT

CHEMICAL FRACTIONS AND PREDICTION FOR LONG-TERM RELEASES OF

PHOSPHORUS IN TYPICAL CANADIAN AGRICULTURAL SOILS

Aruna Withana Herath Co-Advisors: Dr. Gary Parkin University of Guelph, 2013 Dr. T. Q. Zhang

Phosphorus (P) pollution has been identified as the most significant agriculture-related

threat to water quality impairment in Canada. One approach to reduce P pollution is to

identify soils with high P loss potential and develop management strategies to minimize

that risk. This thesis contributes towards greater understanding of short- and long- term

P dynamics in soils to which different P sources had been applied (Chapters 3 and 4)

and improvement in the P measurements for determining long-term P loss potential

(Chapter 5). Chapter 3 evaluated immediate and residual effects of swine manure and

fertilizer on soil P. Soils were sampled from Brookston clay loam in south-western

Ontario, Canada which were treated with liquid (LM), solid (SM), composted (MC)

manure and fertilizer, only in the corn phase. Soils were analyzed using a modified

Hedley’s fractionation. All P sources influenced soil labile and moderately labile P in the

year of application, while MC and SM showed significant residual impacts in the

following year. Residual effects of MC and SM are beneficial for crops; however, there

may be a P loss potential through leaching/runoff.

Chapter 4 considered long-term effects of dairy manure slurry (DMS) and ammonium

nitrate (AN) on soil P. Soils were sampled from south coastal region of BC, Canada,

which were treated with DMS or AN at 50 or 100 kg NH4-N ha-1, and analyzed using a

modified Hedley’s fractionation. DMS significantly increased labile and moderately

stable P in surface soil, indicating short- and long-term impacts on P availability and

loss potential.

Chapter 5 analyzed a new test to predict long-term soil P loss potential. Soils were

collected from four agro-ecological areas across Canada, and analyzed using Mehlich-

3, Olsen, Resin strips (RMS), FeO-strips, and new procedures: various combinations of

NaOH with and without EDTA, with four shaking periods. Statistically significant linear

and quadratic relationships between the RMS and NaOH with EDTA-P indicated that

the latter provide an efficient basis for predicting long-term soil P loss potential. A highly

significant relationship between RMS-P and 0.025M NaOH with EDTA-P indicates this

extractant was effective for measuring Total Releasable P.

iv

Gratefully Dedicated

To

My loving

Amma and Thaththa

Who laid the foundation many years ago,

And

Whose values and vision have remained instrumental

up to

This moment of Accomplishment

v

Acknowledgements

My special and lasting gratitude towards my advisory committee; Dr. Gary Parkin

(Advisor), Dr. Tiequan Zhang (Co-advisor), Dr. Michael Goss (Former advisor,

committee) and Dr. Ivan O’ Halloran (Committee), for their continuous support,

generosity and guidance throughout the period of my graduate studies and for their

invaluable contribution to bring my thesis to a successful completion.

I am very grateful to Dr. Tiequan Zhang for arranging funds for my research through

Agriculture and Agri-Food Canada, and also providing me with all the laboratory

facilities in his lab at Harrow experiment station. I would also like to thank Dr. Chantal

Hamel, Lead Researcher of the GAPS project, for arranging funds for my research. I

wish to thank Mary Ann Reeb and Brian Hohner from the Greenhouse and Processing

Crops Research Center, Agriculture and Agri-Food Canada, Harrow for technical

assistance. I also want to thank Ranee Pararajasingham and Glen Wilson at school of

Environmental Science for sharing some lab facilities.

I am very grateful to my loving parents for supporting me spiritually throughout my life.

To them I dedicate this thesis. Also, I wish to thank my loving sisters and brothers who

always kept me away from family responsibilities and encouraged me to concentrate on

my studies. I am heartily thankful them and their families for their unfailing love and

support throughout my life.

I wish to give warm thanks to all of my friends for their friendly encouragement during

this period and for their support during times of frustration. To my dearest friends and

relatives, there are too many of you to list, but you are present in my heart. Thanks all.

vi

My warmest thank to my loved ones, Isuri, Thilini, and Sachin who are the true joy of my

life. They truly comforted me with unconditional love and support. I hope you always

know how proud I am of you, how grateful I am for you, and how much I love you.

More importantly, it is my greatest pleasure to acknowledge the most influential person

in my life, my husband, Dammika who was always ready to review my rough drafts and

provided me with important feedback. He gave his moral support and encouragement

throughout my graduate studies and always provides me love and laughter to my

everyday life. I cannot thank you enough.

Lastly, I offer my regards and blessings to all of those who supported me in any respect

during the completion of my graduate studies.

vii

Table of Contents

Acknowledgements ......................................................................................................... v

Table of contents ............................................................................................................ vii

List of Tables ................................................................................................................... x

List of figures .................................................................................................................. xii

Chapter 1: General Introduction ...................................................................................... 1

1.1 Aims and Objectives .................................................................................................. 6

Chapter 2: General Literature Review ............................................................................. 8

2.1 Present concerns with soil phosphorus ………………….……….….….... .................. 8

2.2 Soil phosphorus forms and their availability ...............................…............. .............. 9

2.3 Soil phosphorus dynamics and transformations ……………….…..…..... ................ 11

2.3.1 Physical-chemical and chemical processes ……………………………. ................ 12

2.3.2 Biological processes ………………………………………….….…….....…. ............ 15

2.3.2.1 Mineralization …………………………..……………………….….…..…… ........... 16

2.3.2.2 Immobilization…………………………………………………….…...…..….. ......... 17

2.3.3 Effects of cropping ……………………………………………….…....….…. ............. 18

2.3.3.1 Effects of crop residue on soil phosphorus transformations .….….. .................. 18

2.3.3.2 Transformations of phosphorus in the rhizosphere ….…………...…. ................ 19

2.3.4 Effects of phosphate fertilizer application …………………….…...…… ................. 19

2.3.5 Effects of livestock manure application ………………………………… ................. 22

2.4 Soil phosphorus mobility ……………………….………………………….….. ............. 25

2.5 Soil phosphorus transport pathways ……………………………………… ................. 27

2.5.1 Factors affecting soil phosphorus loss ……………….…………….…… ................ 28

2.6 Analysis of soil phosphorus…………………………………………………… ............. 32

2.7 Soil phosphorus testing ………………….………………………………..….. .............. 36

2.8 Research Issue ……………………………………………………………….... ............. 45

Chapter 3: Immediate and residual effects of different forms of swine manure on soil

phosphorus fractions in a clay loam soil under corn-soybean rotation .......................... 48

viii

3.1 Abstract …………………………………………………………………………. ............. 48

3.2 Introduction ............................................................................................................. 49

3.3 Methodology ............................................................................................................ 53

3.3.1 Site descriptions……. ........................................................................................... 53

3.3.2 Treatments, soil sampling and Analysis ............................................................... 53

3.3.3 Hedley sequential phosphorus fractionation ......................................................... 55

3.3.4 Statistical analysis ………………………………………………………….... ............ 57

3.4 Results and discussion …………………………………… ......................................... 57

3.4.1 Labile and moderately labile inorganic P (H2O-P, Bicarb-Pi and NaOH-1-Pi)…… 58

3.4.2 Labile and moderately labile organic P (Bicarb-Po and NaOH-1-Po)…….…. ...... 64

3.4.3 Stable Phosphorus fraction (HCl-Pi +NaOH-2-P +Residual-P)………………….. . 68

3.4.4 Total inorganic P, Total organic P and Total P……………………..…. .................. 72

3.5 Conclusion …………………………………………………………………….. ............... 77

Chapter 4: Phosphorus fractions in a sandy loam soil following long-term applications of

dairy manure slurry and inorganic fertilizer .................................................................... 91

4.1 Abstract ………………………………………………………………….…..….. ............. 91

4.2 Introduction ……………………………………………………………………..………. .. 92

4.3 Materials and Methods ............................................................................................ 93

4.3.1 Site descriptions ................................................................................................... 93

4.3.2 Treatments and soil sampling............................................................................... 94

4.3.3 Soil phosphorus fractionation ............................................................................... 97

4.3.4 Statistical Analysis ……………………………………………………….…. .............. 97

4.4 Results and Discussion……….………………………………..………....………….. ... 98

4.4.1 Labile phosphorus fraction (H2O-P and Bicarb-P) .............................................. 102

4.4.2 Moderately labile phosphorus fraction (NaOH-1-P) ............................................ 105

4.4.3 Moderately stable inorganic phosphorus fraction (HCl-Pi)……… ....................... 107

4.4.4 Stable phosphorus fraction ................................................................................. 109

4.4.5 Total inorganic, total organic and total P ............................................................ 110

4.5 Conclusions …………………………………………………………….…….……. ...... 112

ix

Chapter 5: Development of a soil phosphorus test for predicting long-term soil

phosphorus losses ...................................................................................................... 120

5.1 Abstract ………………………………………………………………………… ............ 120

5.2 Introduction ……………………………………………………………………. ............. 121

5.3 Materials and Methods ………………………………………………………. ............. 128

5.3.1 Site and experiment descriptions …………………………………… .................... 128

5.3.2 Analysis of basic soil properties.………………………………….….. ................... 129

5.3.3 Soil Test Phosphorus………………………………………………………. ............. 133

5.3.3.1 Agronomic phosphorus tests………………………………………...…. .............. 137

5.3.3.2 Environmental phosphorus tests………………………………….... ................... 137

5.3.3.2.1 Anion resin membrane strips (RMS)... …………………………………..…..... 137

5.3.3.2.2 Iron oxide impregnated filter paper strips (FeO-strips).………….. ................ 138

5.3.3.3 Newly proposed extraction methods ……...……………………….... ................ 139

5.3.4 Statistical Analysis …………………………………………………………. ............. 140

5.4 Results and discussion ………………………………………………….……. ............ 141

5.4.1 Comparison of soil P extractability of existing agronomic and environmental soil P

tests…... . .................................................................................................................... 141

5.4.2 Comparison of phosphorus extracted by newly proposed mehods .................... 147

5.4.3 Correlations between amounts of P extracted by existing (agronomic and

environmental) soil P extraction methods and the cumulative amounts of P extracted by

resin membrane strips; “Total Releasable P” .............................................................. 157

5.4.4 Correlations between P extracted by newly proposed methods and the cumulative

amounts of P extracted by resin membrane strips; “Total Releasable Phosphorus” ... 158

5.4.5 Relationship between soil test phosphorus methods ……………………………. 161

5.5 Conclusion ………………………………………………………………………….…. .. 163

Chapter 6: Summary, Conclusions and Future Studies............................................... 181

6.1Summary and Conclusions .................................................................................... 181

6.1 Future studies ....................................................................................................... 187

Chapter 7: References ............................................................................................... 188

x

List of Tables

Table 3.1 Physical and chemical compositions of the different forms of swine manure

materials applied to corn phase in 2006 on a Brookston clay loam at Woodslee, Ontario,

Canada .......................................................................................................................... 79

Table 3.2 Physical and chemical characteristics of the Brookston clay loam soil used in

field plots at Eugene Whalen Research Farm, Woodslee, Ontario, Canada ................. 80

Table 3.3 Distribution of Phosphorus fractions (mg P Kg-1) for different phosphorus

sources in 2006 and 2007 ............................................................................................. 81

Table 3.4 Phosphorus fractions for different phosphorus sources in both cropping

phases (% of Total-P) .................................................................................................... 82

Table 3.5 The significance levels of Analysis of Variance (ANOVA) for main effects .... 83

Table 3.6 Treatment mean differences between 2006 and 2007, and their significant

levels of the Hedley P fractions ..................................................................................... 83

Table 4.1 Physical and chemical characteristics of the silty loam soil used in field plots

at Agassiz, British Columbia, Canada ........................................................................... 94

Table 4.2 Annual fertilizer and dairy manure slurry application rates (Four applications

per year) ........................................................................................................................ 96

Table 4.3 Chemical and physical composition of the dairy manure slurry applied to a tall

fescue sward in a multi-year study in south coastal British Columbia ........................... 96

Table 4.4 The significance levels of Analysis of Variance (ANOVA) for the main effects

(treatments, depth and the interaction and R2 values of the model ............................. 100

xi

Table 4.5 Statistical significance of Depth comparisons from ANOVA (averaged over

treatments) ................................................................................................................. 100

Table 4.6 Mean values of Hedley P fractions (mg P kg-1 soil) of top 0-15 cm soil depth

for all treatments ......................................................................................................... 101

Table 5.1 Different treatments used in field experiments located in four different agro-

ecological areas across Canada ................................................................................. 131

Table 5.2 Physical and chemical characteristics of the soils of field experiments located

in four different agro-ecological areas across Canada ................................................ 132

Table 5.3 Typical soil phosphorus tests used in Canada ............................................ 136

Table 5.4a Descriptive statistics for P (mg P kg-1) extracted by different extractants for

all four sites. RMS = Resin Membrane Strips, Total P (TP) = (Pi + Po), ENa = EDTA +

NaOH, Na = NaOH ...................................................................................................... 145

Table 5.4b Descriptive statistics for P (mg P kg-1) extracted by different extractants for

all four sites. RMS = Resin Membrane Strips, Total P (TP) = (Pi + Po), ENa = EDTA +

NaOH, Na = NaOH ...................................................................................................... 146

Table 5.5 Pearson correlation coefficients (r) between amounts of P extracted by P

extraction methods and Total Releasable P (RMS-P) for the whole soil collection (n=57),

and for different locations (The correlations with * are significant at P < 0.05) ............ 160

xii

List of Figures

Figure 3.1 Modified Hedley sequential fractionation procedure for soil phosphorus (Pi,

Po and Pt refer to inorganic, organic and total P, respectively) ..................................... 56

Figure 3.2 The amounts of water extractable P fraction for soils treated with inorganic

fertilizer P (TSP), Liquid Swine Manure (LM), Swine Manure Compost (MC), Solid

Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and soybean

phase in 2007 (red) in Brookston clay loam soil ............................................................ 84

Figure 3.3 The amounts of NaHCO3 extractable Pi and Po fractions for soils treated with

inorganic fertilizer P (TSP), Liquid Swine Manure (LM), Swine Manure Compost (MC),

Solid Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and

soybean phase in 2007 (red) in Brookston clay loam soil ............................................. 85

Figure 3.4 The amounts of NaOH-1 extractable -Pi and -Po fractions for soils treated

with inorganic fertilizer P (TSP), Liquid Swine Manure (LM), Swine Manure Compost

(MC), Solid Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and

soybean phase in 2007(red) in Brookston clay loam soil .............................................. 86

Figure 3.5 The amounts of HCl extractable -Pi fraction for soils treated with inorganic

fertilizer P (TSP), Liquid Swine Manure (LM), Swine Manure Compost (MC), Solid

Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and soybean

phase in 2007 (red) in Brookston clay loam soil ............................................................ 87

Figure 3.6 The amounts of NaOH-2 extractable -Pi and -Po fractions for soils treated


(MC), Solid Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and


Figure 3.7 The amounts of Residual -P fraction and Total Pt fraction for soils treated


xiii

(MC), Solid Swine Manure (SM) and the CK (CK) for corn phase in 2006 (blue) and


Figure 3.8 The amounts of Total-Pi and Total-Po fractions for soils treated with

inorganic fertilizer P (TSP), Liquid Swine Manure (LM), Swine Manure Compost (MC),

Solid Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and


Figure 4.1a Water extractable Pi (A) and Po (B) fractions in the soil profile of 0-60 cm

soil depth ..................................................................................................................... 113

Figure 4.1b Bicarb-Pi (C) and -Po (D) fractions in the soil profile of 0-60 cm depth .... 114

Figure 4.1c Moderately labile-Pi (E) and moderately labile-Po (F) fractions in the soil

profile of 0-60 cm depth ............................................................................................... 115

Figure 4.1d HCl-Pi (G) and residual-P (H) fractions in the soil profile of 0-60 cm depth

.................................................................................................................................... 116

Figure 4.1e NaOH-2-Pi (I) and NaOH-2-Po (J) fractions in the soil profile of 0-60 cm

depth ........................................................................................................................... 117

Figure 4.1f Total Pi (K) and total Po (L) fractions in the soil profile of 0-60 cm depth .. 118

Figure 4.1g Total soil Pt (M) in the soil profile of 0-60 cm depth ................................. 119

Figure 5.1 Major soil types in the typical agro-ecological systems of Canada ............. 130

Figure 5.2 Means of soil P (mg P kg-1) extracted using agronomic and environmental

soil P extracting methods for soil samples collected from various field plots across four

sites in Canada ........................................................................................................... 144

xiv

Figure 5.3 Means of soil P extracted using various concentrations of NaOH with EDTA

(a) and NaOH without EDTA (b) for four different shaking periods for soil samples

collected from various field plots across four sites in Canada ..................................... 152


(a) and without EDTA (b) for four different shaking periods for soil samples collected

from Harrow field experimental plots ........................................................................... 153



from Agassiz field experimental plots .......................................................................... 154



from Swift Current field experimental plots .................................................................. 155



from Indian Head field experimental plots ................................................................... 156

Figure 5.8 Linear (left) and Non-linear (right) relationships between soil P extracted by

three soil test P methods (FeO-strips, Olsen and Mehlich-3) and cumulative amount of

P extracted by resin membrane strip method (Total Releasable P) ............................ 166

Figure 5.9 Linear (left) and non-linear (right) relationships between soil P (mg P kg-1)

extracted by 0.025M NaOH + EDTA, and cumulative amount of P extracted by resin

membrane strip method (Total Releasable P) ............................................................. 168




xv





extracted by 0.05M NaOH and cumulative amount of P extracted by resin membrane

strip method (Total Releasable P) ............................................................................... 174









strip method (Total Releasable P) … .......................................................................... 180

1

Chapter 1: General Introduction

Phosphorus (P) is an essential plant macronutrient, making up about 0.2 to 0.4% of a

plant’s dry weight (Brady and Weil, 2007). It is a component of key molecules, such as

nucleic acids, phospholipids and adenosine triphosphate (ATP). The functions of P

within the plant include energy transfer and energy storage. Consequently, plants

cannot grow without a reliable supply of this nutrient. However, P does not occur in the

soil as abundantly as other macronutrients such as nitrogen (N) and potassium (K). The

soil solution invariably contains too little P to meet the requirements of actively growing

plants. Accordingly, there has to be a readily available supply of P in the soil to

replenish that in solution as it is taken up by a crop. To meet these requirements,

external P-inputs (inorganic fertilizers and manures) must be added to soils if they are to

produce enough food to sustain humankind both now and in the future.

When the soluble sources of P, such as those in inorganic P fertilizers and organic

amendments, including manure, are added to soils, they react with various soil

components and are changed into forms that are unavailable to plants. Eventually these

P compounds become highly insoluble ‘fixed P’ compounds. These P fixation reactions

in soil may allow only a small fraction (10 to 15%) of the P applied in fertilizers and

manures to be taken up by plants in the year of application (Tisdale et al. 1993; Subba

Rao et al. 1995; Brady and Weil, 2007). These stable forms of organic and inorganic

soil P develop into plant available phosphate forms at a rate too slow to meet crop P

requirements. Thus, in production agriculture, farmers typically apply P fertilizers to soils

in excess of P removed in the crop harvest to optimize crop growth (Brady and Weil,

2007). Over the years, continual application of inorganic fertilizer or manure at levels

2

exceeding crop P needs can lead to accumulation of large amounts of insoluble and

chemically stable forms of soil P. This buildup of the level of soil P can certainly improve

soil fertility status. However, the application of P in excess of that being removed by

crops often results in an increased concentration of soil test P that is associated with an

elevated risk of P losses from the land to neighboring aquatic environments.

Phosphorus is often the most limiting nutrient to the growth of vegetation in freshwater

bodies. Therefore, very small increases in P concentration can result in excessive

growth of aquatic vegetation, leading to eutrophication. Eutrophication impedes water

use for fisheries and other industries because of the increased growth of undesirable

algae and aquatic weeds. The masses of dead algae and other organic materials in the

aquatic bodies are continuously decomposed, utilizing the dissolved oxygen in water,

thereby depleting the available gaseous oxygen (Sharpley et al. 2000). As a

consequence, eutrophic water becomes inhospitable to aquatic flora and fauna with

lower dissolved oxygen levels and increased turbidity leading to loss of aquatic

biodiversity (Hansen et al. 2002). A nutrient-rich aquatic environment may also lead to

periodic surface blooms of cyanobacteria (blue green algae) and contamination of

drinking water sources (Sharpley et al. 2003). The highly toxic and volatile chemicals

released by these organisms can cause serious health hazards to animals and humans.

In addition, poor taste and foul odor of eutrophic freshwater decrease the recreational

amenity values for such activities as swimming and boating. Thus, minimizing P inputs

to the freshwater bodies through proper P management is of prime importance in

reducing eutrophication of freshwater bodies.

3

Eutrophication can be accelerated by any human activities that increase P entering into

surface waters through point sources and non-point sources. Point- sources of P, such

as effluent discharge from wastewater treatment plants, sewage treatment plants and

industrial factories, are relatively easy to identify and regulate. Accordingly, during the

past several decades many developed countries have successfully reduced the mass of

P entering into surface water bodies from point sources. However, there has been little

success in regulating and managing non-point source pollution of P such as runoff

water, eroded sediments and subsurface movement from nearby agricultural lands.

Currently, these non-point sources of P are often the main cause of excessive P in

aquatic environments that lead to eutrophication.

The rate of pollution from non-point source P is higher mostly with intensive agriculture,

which involves heavy P application through inorganic phosphate fertilizers and organic

amendments, such as animal manure. Presently, one of the major concerns is high P

loading in agricultural soils through excessive rates of livestock manure application. This

is often an issue in the areas of intensive livestock production where soils are found to

be enriched with P due to the overloading of manure onto a limited area of cropland.

High transportation costs and other logistic difficulties prevent livestock manure from

being transported to farms in distant locations, where it could supplement or even

replace mineral fertilizer requirements.

Buildup of soil P may be made even greater by inadvertent P inputs when manure with

a small N: P ratio is applied to provide the required amount of N to crops (see Chapter

2). Excess P from N-based manure application increases the soil test P concentrations

far beyond the crop requirements. This accumulation of excess P in the soil is often

4

accompanied by increases in the degree of P saturation (Simard et al. 1995; Zheng et

al. 2001). Both of these factors can increase the risk of P transport from agricultural

lands to nearby freshwater bodies through runoff, erosion and tile drainage.

Over the last three decades, agricultural researchers have been exploring

environmentally sound P management practices in agricultural lands to protect surface

water quality. To manage P for economically optimal crop production and for

environmental protection, it is necessary to develop ways of maximizing the efficiency of

P utilization by crops and minimizing P losses from soil to water bodies. One way to

minimize P losses from crop lands is to ensure that the soils are not over- enriched with

P. Thus, balancing P inputs and outputs is one of the main challenges to make modern

farming systems both economically and environmentally sustainable. To meet such a

challenge one must be able to identify sites, soils and management systems that are

vulnerable to P losses, so that appropriate remedial measures can be targeted

effectively. To achieve this, environmental assessments are required at the farm level;

for example, soil tests are needed to assess the potential of a soil to release P to runoff

rather than for its availability to plants. To ensure that soil P levels are not allowed to

exceed those that represent a threat to the environment, it is imperative to identify a

critical level of P in the soil. The upper limit of manure that could be applied with

minimum environmental pollution could be determined by the pre-application level of P

in the topsoil. Therefore, testing of soil P to assess its availability is currently the best

management tool available to ensure that crops are provided with adequate, but not

excessive supplies of the nutrient.

5

Recently, there has been an increased interest in using existing agronomic soil tests as

an indicator of potential environmental risk of P loss leading to eutrophication. The use

of locally-calibrated routine soil tests as environmental risk indicators has a significant

practical advantage due to the widespread use of such tests and their well- established

agronomic applications together with the large data base they provide on soil P levels.

Details of extraction methods commonly adopted in such tests are included in Chapter

two. However, for environmentally sound predictions, the suitability of an agronomic soil

test P depends on the chemical nature of the extractants, procedure, soil type and

management and the climatic conditions. Therefore it is difficult to assign universally

acceptable routine soil test P values for environmentally sustainable management

systems.

In addition to the use of agronomic tests, several new soil P tests have been developed

by researchers that might better estimate the potential of soil P losses to aquatic

environments. Some of the most promising methods are distilled water extraction and

‘Ion Sink’ methods (more details are included in Chapter two). These methods are

designed to extract all or a representative amount of P fraction in soils that could be

easily lost through surface runoff, erosion and sub-surface leaching. However, these

methods have some limitations for their use as environmental soil P tests and there are

improvements needed that are addressed in this thesis.

6

1.1 Aims and Objectives

Taking account of all the agronomic and environmental soil P test methods, it is evident

from the literature that their overall effectiveness in predicting P losses from agricultural

soils is limited. Therefore, better soil testing methods that are more precise and

accurate have to be developed to have reliability in identification of soils where P loss to

the environment will be of concern. The overall goal of this thesis is to develop an

environmentally sound Soil Test Phosphorus (STP) method for more relevant

measurements of soil P status, to identify the forms of P present and the prediction of

long-term P loss potential. The test must be applicable to agricultural lands under

different soil-crop management practices, such as different sources of P (mineral

fertilizer or organic amendments, including manure) and in different cropping systems.

To accomplish this, the more specific objectives of this dissertation are:

1. Determine the immediate and residual effects of swine manure and manure

compost on soil P fractions in a clay loam soil under corn-soybean rotation

2. Determine the short- term and long-term effects of soil P inputs from both dairy

manure slurry and mineral fertilizer on soil P forms and their distribution in a silty

to sandy loam soil profile

3. Develop a new soil P test for prediction of long-term soil P loss potential

These three objectives contribute to increase the knowledge regarding changes in soil P

forms and their behavior in soils with different agricultural management practices that

7

collectively advance the ultimate goals of protecting surface water quality and

preserving farm productivity.

This dissertation is organized in the form of three separate journal articles; the first

paper (Chapter 3) is entitled “Immediate and residual effects of swine manure and

manure compost on soil P fractions in a clay loam soil under corn-soybean crop

rotation”. The second paper (Chapter 4) is entitled “Phosphorus fractions in a sandy

loam soil following long-term application of dairy manure slurry and inorganic fertilizer”.

The third paper (Chapter 5) is entitled “Development of a soil P test for prediction of

long-term soil P loss”. Each paper includes an abstract, introduction, methodology,

results and discussion, conclusions, tables and figures. These papers are preceded by

a Thesis Abstract, General Introduction, General Literature Review and are followed by

Thesis Summary including general conclusions and future research requirements and

References.

8

Chapter 2: General literature review

2.1 Present concerns with soil phosphorus

The native P levels in soils are small enough to limit crop production. Therefore both

inorganic fertilizer and organic sources are applied to provide P and correct deficiencies

in the soil. When these sources are properly managed, P additions can increase crop

production to an optimum level with minimum losses to the environment. However,

inadequate management of these sources can result in increasing the soil test P

concentrations far beyond crop requirements, thereby increasing the risk of P losses to

the aquatic environment. Phosphorus, in particular, has been found to be the limiting

nutrient in freshwater ecosystems and its presence, even in relatively small amounts,

greatly accelerates the potential for water quality degradation through the process of

eutrophication. Eutrophication results in increased growth rates of algae and aquatic

macrophytes which can cause recreational problems, decrease water quality, clarity and

depth of light transmission, result in unpleasant odor, anoxia, as well as toxicity to fish,

other life forms in the aquatic environment, livestock, and even humans.

To manage P for economic crop production and for environmental protection, we have

to understand the nature and the availability of the different forms of the P found in soils,

the manner in which the various forms interact within the soil and in the environment,

and also the effects of different management practices (such as cropping and fertilizer

and manure application) on these forms and their transformations within the soil P

fractions. Essential details of these topics will be briefly discussed in the following

literature review.

9

2.2 Soil Phosphorus forms and their availability

The amount of total P (mass percentage) in a soil can range from 0.02% to 0.5%, with

an average of approximately 0.05% (Barber, 1995). However, only a minute part of the

total P (usually <1%) is available for plant growth at any given time and this P fraction is

referred to as “Available Phosphorus”. The bulk of the soil P exists as groups of

compounds: organic P, Ca-bound inorganic P, and Fe- or Al- bound inorganic P. All

these groups of compounds slowly contribute P to the soil solution, but most of the P in

each group is of very low solubility and not readily available for plant uptake.

Soil P content varies with parent material, extent of pedogenesis, soil texture, and

management factors, such as rates and types of P applied and soil cultivation

(Sharpley, 2000). Even in the same soil profile, P content of surface horizons is usually

greater than that of the subsoil (Sharpley, 2000). This is due to the sorption of added P

by soil constituents in surface soils, more organic materials in surface soil layers,

greater micro-biological activity and cycling of P from roots to above-ground plant

biomass. Especially, in minimum tillage or zero-tillage systems, buildup of excess P in

the top 5 to 12 cm of soil is very likely, given that fertilizers and manures are not

incorporated into the soil or they are incorporated only to shallow depths with such

tillage practices.

Soil P can be categorized into two major groups; inorganic and organic P. The inorganic

P forms originate from weathering of primary minerals, and from the addition of

inorganic P fertilizer and organic amendments, such as livestock manure. Organic P

forms originate from organic amendments, plant residues and products of soil microbes.

10

Inorganic P forms are associated with amorphous and crystalline sesquioxides and

calcareous compounds. The organic P is associated with labile phospholipids and fulvic

acids and more resistant humic acids. Generally, in most agricultural soils, the inorganic

P content is greater than the organic P content. For example, Sharply (2000) has found

that, 50 to 75% of the P found in most agricultural soils exists as inorganic P, although

this fraction can vary from 10 to 90%. The main exception being peat (Histosols), where

essentially all the P occurs in organic forms (Stevenson, 1986). Generally, prairie

grassland soils, forest soils and certain tropical soils also contain relatively high

amounts of P in organic forms (Stevenson, 1986). However, both inorganic and organic

P contents vary with soil type and other environmental and management practices such

as cropping, inorganic fertilizer and organic manure application.

The knowledge of the specific nature of most of the organic-bound P in soils is quite

limited. However, three broad groups of organic P compounds are known to exist in

soils; inositol phosphates, phospholipids and nucleic acids. These organic P

compounds are mineralized in soils by a reaction which is catalyzed by the enzyme

phosphatase. Inositol phosphates are the most abundant of the known organic P

compounds, comprises up to10 to 50 % of the total organic P (Brady, 2007), and

represent a series of phosphate esters ranging from monophosphate up to

hexaphosphate. Most of inositol phosphates in soils are products of microbial activity

and degradation of plant residues.

Nucleic acids occur in all living cells and are produced during the decomposition of

residues by soil microorganisms. Two distinct forms of nucleic acids, DNA and RNA are

released into the soil in greater quantities relative to inositol phosphates. However, DNA

11

and RNA are mineralized in most soils much more rapidly and incorporated into the

microbial biomass. Therefore nucleic acids represent only a small portion of total

organic P in soils, approximately 2.5% or less (Condron et al.1985).

Phospholipids are organic compounds, insoluble in water, but are readily utilized and

synthesized by soil microorganisms. Thus, the free phospholipids content in soils is also

small, representing only 5% or less of the total organic P. The most common

phospholipids are derivatives of glycerol dominated by choline phospho-glyceride,

followed by ethanolamine phospho-glyceride (Dalal, 1977). Other soil organic P

compounds can be sugar phosphates (Anderson and Malcolm, 1974) and phosphate

proteins that contribute to trace amounts of soil organic P.

Phosphorus enters the soil solution through dissolution of primary and secondary

minerals; desorption of P from clays, oxides, and minerals; and biological conversion of

P in organic materials to inorganic forms (mineralization).The next section will discuss

the processes that control the forms of soil P and their availability to plants.

2.3 Soil Phosphorus dynamics and transformations

The P cycle in soil is a dynamic system involving the soil, plants and micro-organisms.

In the closed natural system, essentially all the P consumed by plants is returned to the

soil either as plant or animal residues. However, in agricultural systems, some P is

removed from the soil with harvested parts of crops and only a portion of this P returns

directly to the soil.

Phosphorus transformations in soils involve complex mineralogical, chemical, physical-

chemical and biological processes (Frossard et al. 2000). Chemical processes include

12

precipitation and dissolution, while physical-chemical processes include adsorption and

desorption, and biological processes involve immobilization and mineralization. The

physical-chemical and chemical processes that influence soil P forms and their

availability are discussed in the next section.

2.3.1 Physical-chemical and chemical Processes

Chemical weathering and solubilization of soil mineral phosphates release P in the form

of orthophosphates into the soil solution, where they exist in very small concentrations.

Apatite [(Ca10 (PO4)6 (X)2, where X represents F-, Cl-, OH- or CO32-)] is the most

common primary P mineral in soils with high pH. In soils with low pH, Variscite

(AlPO4.2H2O) and Strengite (FePO4.2H2O) are the probable P minerals (Savant and

Racz, 1973; Lindsay, 1979).

When phosphate fertilizers are added to soils, initially P in water soluble compounds

goes into the soil solution as phosphate ions. Once in the soil solution, P can be taken

up by plant roots, assimilated by biological organisms, sorbed to sesquioxides and

calcium carbonates, precipitate as an insoluble compounds, or be lost in surface or

subsurface runoff. The major processes for removal of inorganic P from soil solution are

generally adsorption to soil surfaces and conversion to secondary P minerals. This

phenomenon is often referred to as ‘P fixation’ which involves both P sorption and P

precipitation mechanisms. Phosphorus sorption covers surface adsorption and

absorption that is subsequent to the penetration of P into the retaining component.

Generally, clay content approximates the reactive surface area responsible for P

sorption of a soil (Sharpley et al. 1984a; Hedley et al. 1995).

13

Phosphorus fixation in soil consists of two different patterns; an initial rapid adsorption

process followed by a very much slower reaction process. When soluble P is added to a

soil, initially the phosphate ions dissolve in the soil solution. However, a rapid reaction

then removes P from solution. This rapid reaction involves an exchange of P with

anions on the surfaces of Fe- and Al-oxides (Rajan and Fox, 1975; Rao and Sridharan,

1984). It has been found that a number of soil minerals and soil colloids are involved in

adsorbing phosphates from the soil solution. In calcareous soils, it appears that a

surface coating of phosphate can be formed on calcium carbonate. In neutral and acidic

soils, P is more likely to be adsorbed on hydrated Fe oxides, Al oxides or on the edges

of clay minerals.

The adsorption capacity of the soil is mainly determined by the nature and the amount

of soil components, i.e. the type of surfaces which may be in contact with P in the soil

solution and the number of sites available for reaction with added P. In soils with

significant contents of Fe and Al oxides, where the oxides are less crystalline, the P

adsorption capacity is greater owing to the larger surface area. For instance, hydrous

metal oxides of Fe and Al, which are most abundant in weathered soils, have a greater

capacity to adsorb very large amounts of P. Therefore, highly weathered soils are

characterized by small values of total available P because a large proportion of

inorganic phosphates are removed as relatively insoluble Al and Fe forms from the soil

solution, leading in turn to greater P requirements in these soils.

Slower reactions are considered to involve slow sorption and precipitation. In the slow

reaction process, phosphate is removed from solution over a long period of time and

there is continual gradual reduction in the solubility of total P. Therefore, the freshly

14

fixed P may be slightly soluble and some value to plants. However, with time, the

solubility of fixed P tends to decrease to extremely low levels. Both sorption and

precipitation mechanisms occur in all types of soils at all pH levels. However, the type

and the relative proportion of the fixed P compounds depend mainly on the nature of the

clay particles and the pH of the soil. Generally, Ca controls these reactions in neutral or

calcareous environments, while Al and Fe are the dominant controlling cations in acidic

environments.

Precipitation involves transformation of soluble P into relatively less soluble Fe, Al, Ca

and Mg phosphates, which control P concentration in the soil solution. Addition of

fertilizer P leads to the formation of strongly concentrated P solutions and often a low

pH in the vicinity of the fertilizer granule. Sample et al. (1980) have reported that P

concentration could be as high as 1.5 M to 12 M and pH can be as low as 1 depending

on the source of P. This acidification effect may cause degradation of clay mineral

structures, dissolution of CaCO3 and subsequent precipitation of amorphous Al-

phosphates and Ca-phosphates (Freeman and Rowell, 1981). Generally, precipitation

reactions are favoured by the very large P concentrations existing in close proximity to

granules, droplets and bands of fertilizer P. However, adsorption seems to be the

dominant mechanism with small P concentrations in the soil solution developing over a

short period of time. Therefore adsorption reactions are most important at the periphery

of the soil-fertilizer reaction zone, where P concentration is much smaller.

Local soil conditions that lead to precipitation of Fe-, Al- and Ca-phosphates usually

change with time. However, in the long run, the formation of various Al- and Fe-

phosphates may depend mostly on average soil properties. The initial compounds

15

precipitated are likely to be meta-stable and will usually change with time into more

stable and less soluble compounds, or re-dissolve into the soil solution. However, these

more stable phosphate products are unlikely to govern P concentration in soil solution

(Ryden and Pratt, 1980).

Soil P transformations are generally affected by soil biological reactions, and the next

section will discuss the biological processes which influence soil P forms, their

transformations and availability.

2.3.2 Biological Processes

Generally, P mineralization and immobilization processes occur in soils simultaneously.

Accordingly, the maintenance of soluble phosphate in the soil solution will depend to

some extent on the magnitude of the two opposing processes. The quantities and rates

of P mineralization and immobilization are determined by many soil factors, such as soil

temperature, moisture, aeration and soil pH, intensity of cultivation, total organic C and

P fertilizer applications. For example, additions of fertilizer P can lead to increase in soil

organic P through net immobilization.

The P content of decomposing organic residues plays a key role in regulating the

quantity of soluble P in the soil at any one time. The C/P ratio of the decomposing

residues regulates the predominance of P mineralization over immobilization. Generally,

net immobilization of soluble P is most likely to occur when the C/P ratios of added crop

residues are 300 or more and net mineralization of organic P occurs when the ratio is

200 or less (Stevenson, 1986). When residues are added to soil, net immobilization

occurs during the early stages of decomposition, followed by net P mineralization as the

16

C/P ratio of the residue decreases. When the C/P ratio is in between 200 and 300, no

gain or loss of inorganic P occurs.

2.3.2.1 Mineralization

The initial sources of soil organic P are plant and animal residues, which decompose

under the action of microorganisms to produce other organic compounds and release

inorganic P. A wide range of soil microorganisms are capable of mineralizing organic P.

However, some organic P is resistant to microbial degradation and is most likely

associated with humic acids. According to Charter and Mattingly, (1979), the calculated

rates of inorganic P released by mineralization of soil organic P accounted for up to

67% of the total P uptake by crops in an unfertilized soils in eastern Canada. Barber,

(1979) found that in cropland to which no fertilizer was added, there was a more-or-less

constant content of inorganic P, although levels of available P gradually declined. The

net loss of P from the system through removal of harvested crop was primarily

accounted for by a decrease in organic forms.

The quantity of P mineralized during a crop growing season varies widely among soils

due to soil characteristics and environmental conditions. Typically, large quantities of

organic P are mineralized in the tropics where distinct wet and dry seasons and warmer

soil temperatures enhance microbial activity. Research has found that in tropical high-

temperature environments, organic P mineralization tends to be greater, ranging from

67 to 157 kg P ha-1 yr-1 and can apparently supply a large part of the crop P requirement

(Anderson, 1980). However, the slow-continual release of P would also be important in

soils with a high capacity to fix inorganic P as well. In the temperate zone, organic P

17

mineralization rates are relatively slow due to cool temperatures and less microbial

activity. Larson et al. (1972) reported that an annual mineralization rate for temperate

soils was ≈ 10 kg P ha-1. Stewart and Tiessen, (1987) found that the amounts of organic

P mineralized in temperate dry-land soils range from 5 to 20 kg P ha-1 yr-1. Since

inadequate levels of P are mineralized during a given cultivation and growing season in

temperate soils, regular additions of fertilizer P are necessary to maintain optimum

levels of P for plant growth.

2.3.2.2 Immobilization

Mineral P from fertilizer can be immobilized to organic P by soil microorganisms.

Research has shown that continuous applications of fertilizer P could lead to an

increase in soil organic P (Zhang and MacKenzie, 1997) and such increases take place

through net immobilization of P in crop residues and conversion of available inorganic P

into microbial biomass and organic P (McLaughlin et al. 1977).

Soil P transformations are affected by different soil factors such as pH, texture, organic

matter content, the amount of CaCO3, Fe- and Al- oxides, as well as soil temperature

and moisture content. However, in a given agricultural soil, cropping and fertilizer

applications that alter the status of soil organic matter and P concentrations in the soil

solution, are the most important factors influencing transformations in the soil (Beck and

Sanchez, 1994; Zhang and MacKenzie, 1997; Zheng et al. 2002). In the next section,

the effects of cropping on soil P forms and their transformations are discussed.

18

2.3.3 Effects of cropping

Generally, long-term cropping without fertilizer or manure additions leads to a reduction

in soil P content. Several studies have explored the relative changes in soil organic and

inorganic P fractions due to long-term cultivation. Hedley et al. (1982) found that

continuous cropping without addition of inorganic P fertilizer resulted in the depletion of

soil organic P. Adepetu and Corey (1977) observed that 25% of the organic P content of

some Nigerian surface soils was mineralized during the first two cropping periods after

initial cultivation. Similar results were reported by McKenzie et al. (1992a) in a study

focused on the effects of different cropping systems (continuous wheat, fallow-wheat

and fallow-wheat-wheat) on soil P fractions in the soil, indicating that the mineralization

of organic P occurred in a considerable degree during cultivation to supply crop P

needs.

Crop production modifies soil P transformations through addition of crop residues and

release of exudates into the rhizosphere. The effects of crop residue on soil P

transformations are discussed in the next section.

2.3.3.1 Effects of crop residue on soil Phosphorus transformations

Crop residues are a potential source of plant nutrients, which may be released into the

soil during their decomposition and are then available for uptake by following crops. The

transformations of residual P in soil into plant available P forms and its contribution to

crop P nutrition depend on inherent soil properties, cropping system, fertilizer

management and history, and climatic conditions. The influence of rhizosphere

characteristics on soil P transformations is discussed in next section.

19

2.3.3.2 Transformations of Phosphorus in the rhizosphere

The rhizosphere is the unique volume of soil that is directly affected by the activity of

plant roots and where roots, soil and the soil biota all interact. Most of the interactions

are beneficial for plants as they improve soil fertility. When the plant roots absorb

nutrients from the soil solution, they release exudates into the rhizosphere. Therefore,

the chemical, physical and biological properties of rhizosphere soil are markedly

different from the bulk soil in many respects such as; (1) the presence of a larger

number of microorganisms, (2) a larger amount of organic C as a result of root exudates

and materials sloughed off from root surfaces, (3) the presence of small-molecular-

weight organic acids secreted by plant roots, and (4) lower pH due to differential uptake

of cations over anions by plant roots (depending on N source). Because of this unique

nature, the rhizosphere is a key site for P transformation with a significant mobilization

of P from the non-exchangeable inorganic and organic P fractions.

2.3.4 Effects of phosphate fertilizer application

Phosphorus fertilizers are added to soil to improve productivity. The level of P addition

varies with both soil and plant type (Pierzynski and Logan, 1993). For example, Tiessen

(2008) has found that crop requirements for fertilizer P varied from <1 kg P ha-1 in

relatively un-weathered soils of arid environments to 200 or 300 kg P ha-1 in oxide rich

tropical or volcanic soils. When fertilizer P is added to the soil, soluble P goes into soil

solution before reacting with the soil and initially increases solution P. But subsequently

solution P decreases due to both biological and chemical processes of P fixation. The

duration of elevated solution P levels depends on the application rate of fertilizer P, the

20

method of placement, the quantity of P removed by the crop together with the soil

properties that influence P availability. From the soil solution, P is either taken up by the

crop, becomes weakly (physical) or strongly (chemical) adsorbed onto Al, Fe and Ca

surfaces, or incorporated into organic (microbial) P (McLaughlin et al. 1988; Syers and

Curtin, 1989).

Most of the P (as high as 90 % or more) applied to soil as fertilizer, is not taken up by

the crop, but is retained in insoluble or fixed forms. While a portion of the residual P can

be used by subsequent crops, further additions of fertilizer are often required in order to

maintain high crop yields. Therefore, over the years, repeated applications of fertilizer P

in amounts exceeding crop uptake inevitably result in an accumulation of P in the soil.

The extent of P accumulation depends on both fertilizer application rate and years of

application. For example, Barber (1979) found P accumulation when P application rates

exceeded 22 kg ha-1 yr-1 in a rotation-fertility experiment over a 25-year period.

Carpenter et al. (1998) also reported that the estimated value for the average annual

rate of soil P accumulation resulting from fertilizer applications was about 22 kg P ha-1 in

the USA and Europe.

Several studies have explored the influence of applied fertilizer P on the different P

fractions in soil. MacKenzie et al. (1992) and Schmidt et al. (1996) reported that

continuous application of fertilizer P for many years has increased inorganic and organic

P fractions in soil. Based on a study of eight agriculturally important soil series in the US

comparing the relative amounts and distribution of P forms in virgin soil profiles with

those of similar soils that had been cultivated and fertilized for at least 15 years,

Sharpley and Smith (1983) found an average decrease of 43% in soil organic P content

21

of cultivated surface horizons, although fertilizer application increased the total P by

25% compared to their virgin analogues. The authors asserted that as a result of

fertilizer application and organic P mineralization during cultivation, the inorganic P

content increased by 118% and plant available (Bray-1-P) inorganic P content

increased by 85% for all surface soil horizons. However, they observed that the content

of all P forms in the sub-soils was relatively unaffected by cultivation. And the change in

content of P forms with cultivation and fertilizer P application was closely correlated with

P content of the virgin soils and the amount of fertilizer P added during cultivation.

The sink with which excessive P is mostly associated depends on soil type, cropping

system and climatic conditions. For example, in a field study conducted on a Hawaiian

Ultisol, Linquist et al. (1997) reported that one year after fertilizer application almost

40% of the applied triple super phosphate fertilizer was in the hot HCl and H2SO4

fractions. O’Halloran (1993) reported that fertilizer application increased the resin

extractable inorganic P, NaOH extractable inorganic P and residual P and total P

fractions of a sandy loam soil, however, for a clay soil, only the resin P fraction was

affected. Beck and Sanchez (1994) found that in a tropical Ultisol of the Amazon basin,

with 18 years of continuous cultivation, the NaOH extractable inorganic P fraction

served as a primary sink for fertilizer P added to soil. Similar results were reported by

Richards et al. (1995); Motavalli and Miles (2002), for different cropping systems with

continuous fertilizer application of soils for more than 9 years. While the assertion of

Beck and Sanchez (1994) has been supported by studies on temperate soils by

Schmidt et al. (1996) and Zhang and Mackenzie (1997a, 1997b), the later authors found

involvement of three other P fractions as a P sink, namely NaOH-extractable organic P

22

(NaOH-Po), HCl-extractable P (Ca-P) and the residual P fractions. Zhang and

MacKenzie (1997) have reported that application of fertilizer P to a Hapludalf soil has

increased the labile and moderately labile inorganic P fractions. In a Chicot sandy clay

loam soil, NaOH extractable inorganic P fraction was the major sink for P from the

excessively added fertilizer P (Zhang and MacKenzie, 1997). Wager et al. (1986) found

that, after one year, most fertilizer P added to calcareous Canadian soils entered the

easily extractable inorganic P fractions. In subsequent years (2-8) these fractions were

depleted due to crop P uptake and transformation to more stable forms of P.

2.3.5 Effects of livestock manure application

In production agriculture, livestock manures are used as a soil amendment to improve

soil quality and productivity. The availability of P from manure is somewhat different

from that of mineral fertilizer. For example, Lucero et al. (1995) found that 3 – 4.5 kg P

ha-1 from poultry manure and 16.5 kg P ha-1 from mineral fertilizer were required to

increase Mehlich-3- P by 1 mg kg-1. Reddy et al. (1999) showed very similar results,

with 5.6 kg P ha-1 from poultry manure and 17.9 kg P ha-1 from mineral fertilizer needed

to increase Olsen extractable P by 1 mg kg-1. Some research suggests that manure P

may be equally available or more available to plants than fertilizer P (Gale et al. 2000;

Meek et al. 1979; Abbott and Tucker, 1973). Elias-Azar et al. (1980) found that P from

fresh and composted dairy manure was as available as P from KH2PO4 in alkaline

sandy soils. However, manure P has not always been found to be more plant available

than fertilizer P. For example, Sharpley and Sisak (1997) found that the availability of P

from poultry manure leachate was somewhat less than that from KH2PO4.

23

Researchers have also found that, when manure and inorganic fertilizer are applied

together, a synergistic effect occurs whereby available P is increased more than the

sum of the increase from either applied separately (Copeland and Merkle, 1941; Dalton

et al. 1952; Toor and Bahl, 1997; Reddy et al. 1999). This synergistic effect may be

explained by the fact that several anions of organic acids have been found to prevent P

fixation and are able to replace P bound to the soil resulting in greater concentrations of

available P (Nagarajah et al. 1970; Kafkafi et al. 1988). However, In a study conducted

by O’Halloran and Sigrist, (1993) to find the effects of incubating monocalcium

phosphate monohydrate (MCPM) with liquid hog manure (LHM) on P availability, the

results indicated that the additions of both MCPM and LHM had the same effect on soil

P fractions regardless of whether the materials had been incubated together or added to

the soil separately.

The application of livestock manure to soil significantly influences the chemical, physical

and biological properties of soil (Sommerfeldt and Chang, 1985) thereby providing

several potential benefits for crop growth. Improving soil physical properties such as

bulk density, soil compaction and aeration or porosity can enhance root growth (Egrinya

et al. 2001). Manure application can also enhance water retention and water holding

capacity, especially in sandy soils, thus aiding nutrient uptake through both diffusion

and mass flow. There may even be manure impacts on soil color and subsequently on

soil temperature, thus aiding root growth particularly in the spring. The effects of manure

on increasing soil organic carbon and improving tilth in favour of crop emergence and

growth have been reported for years by many researchers (Hoyt and Rice, 1977; Meek

et al. 1982; Sommerfeldt et al. 1988). Manure also has a residual nutritional effect as

24

mineralization of its organic fraction often takes longer than a cropping season (Cusick

et al. 2006; Zhang et al. 2006). It has been reported that manuring tends to move soil

pH towards neutrality, whether in acidic soils (Whalen et al. 2000) or alkaline soils

(Chang et al. 1990), thus improving nutrient availability, especially for P and other

micronutrients by maintaining the soil pH in the ideal range for most field crops.

Most of these effects of manure application are attributable to an increase in soil organic

matter, because animal manure, especially solid and slurry manure, contains a large

amount of organic matter. The ability of organic matter to promote formation of water-

stable aggregates in the soil has a substantial effect on soil structure and physical

characteristics. For example, increasing the stability of aggregates to excess water

could increase infiltration, porosity and water holding capacity of the soil and thereby

decrease soil compaction and erosion. Organic matter increases the cation exchange

capacity of the soil and serves as a buffer against rapid change in pH. It also serves as

an energy source for soil micro-organisms. Consequently, an increase in the organic

matter content of a soil from manure application improves overall soil fertility and

productivity for plant growth.

The rate of manure application is usually based on the plant-available N content of the

manure and the recommended N rate for the crop to be grown. Generally the manure

N:P ratio is often smaller than the N:P uptake ratio of most crops (Simard et al. 1995;

Gburek et al. 2000; Whalen, 2001). Some manure sources particularly ruminant and

poultry manure are reported to have a high P content relative to N content (Leaver

1984; Nguyen and Goh 1992; Greatz and Nair 1995; Mikkelsen, 2000). Therefore,

manure applied to meet crop N requirements often causes accumulation of P in soils

25

exceeding the optimum concentration needed for crop production (Mozaffari and Sims,

1994, Simard et al. 1995; Sims et al. 2000; Whalen and Chang, 2001). This buildup of P

in the soil could be initially beneficial to crop growth, especially in soils that are deficient

in P. However, this accumulation is accompanied by increases in soil-test P levels

(Simard et al. 1995, Zheng et al. 2001), that can pose environmental problems, through

increases in soluble/dissolved P in runoff water (Pote et l. 1996; Andraski and Bundy,

2003; Fang et al. 2002; Schroeder et al. 2004; Sims et al. 1998) and downward

movement of P to ground water (Haygarth et al. 1998).

The next section will discuss movement of P in soil, including different transport

pathways and loss potential to the environment.

2.4 Soil Phosphorus mobility

Phosphorus is the least mobile macro nutrient in soil and this poor mobility of P is

mostly due to the strong reactivity of phosphate ions with the soil components. The

mobility of phosphate ions in soil depends on the nature of mineral surfaces present,

since phosphate anions are strongly adsorbed by mineral constituents such as clays

and sesquioxides (Parfitt, 1978).

Mobility of phosphate in soil is also dependent on its form in a soil. Generally, organic P

anions can move to a greater depth than inorganic P in soil (Chardon et al. 1997). For

example, an accumulation of inositol phosphate in the B horizon of soil was found by

Halstead and McKercher (1975). Several studies also demonstrated an appreciable

downward movement of P following application of manure which resulted in elevated P

levels at 60 -120 cm depths. Chater and Mattingly (1979) have found that application of

26

inorganic fertilizer P resulted in much less downward movement of P compared to the

organic P in manure.

The addition of decomposable organic matter enhances the mobility of P in soil because

the concentrations of PO4-P in soil solutions have increased as the residue loading

rates were increased (McDowell et al. 1980). The mechanisms that may most likely be

involved in the decomposition of organic residues and its contribution to soil P mobility

are explained by Stevenson (1994) as:

(1) Formation of chelate complexes with Ca, Fe and Al and the subsequent release of

phosphate to water soluble forms

(2) Competition between humate and phosphate ions for adsorbing surfaces prevents

fixation of phosphate

(3) Formation of protective coatings over colloidal sesquioxides with reduction in

phosphate adsorption and

(4) Formation of phospho-humic complexes through bridging with Fe and/or Al.

Accordingly, management practices that increase the loading of crop residue may

increase P mobility in a soil profile.

The next section discusses the different pathways and the mechanisms of soil P

transportation to neighboring water bodies and which forms are transported through

each path.

27

2.5 Soil Phosphorus transport pathways

Phosphorus can be transported from agricultural fields to adjacent surface water bodies

through surface (runoff and erosion) and/or subsurface (leaching and preferential flow)

pathways. Under typical agricultural practices, P remains concentrated in the surface

horizon which leads to greater P loss through surface transport and less P loss through

subsurface leaching. Studies have reported that the greatest potential for P losses from

soil is through erosion (Tiessen et al. 1983) and surface runoff (Khasawneh et al. 1980).

Phosphorus losses through leaching are much smaller than losses by soil erosion and

surface runoff because P becomes a part of very insoluble and less mobile compounds

in the soil. It is reported that 10% of P export from land occurs by leaching and ground

water transport while 90% is transported by overland flow through erosion and surface

runoff. Although the proportion of P leaching is much smaller than overland flow loss, it

has a greater effect on eutrophication because it is soluble P and therefore readily

available to aquatic biota. However, subsurface runoff (i.e. tile drainage) has been

reported recently as a major soil P loss pathway (up to 95% of total soil P loss) in tile

drained lands that are increasingly a common practice in North America (Zhang et al.

2009: Tan and Zhang, 2011).

Phosphorus can be lost from agricultural lands in two different forms; particulate-bound

(sediment- bound) form and dissolved (soluble) form (Sims et al. 1998). Particulate

bound P forms include P associated with soil particles and organic materials eroded

during flow events and constitute about 75 to 90% of total P (TP) transported in surface

runoff from most cultivated lands (Sharpley et al. 1993). However, surface runoff from

28

grass, forest, or non-cultivated soils carries little sediment and is therefore generally

dominated by the dissolved form of P. The dissolved form comes from the release of P

from soil and plant materials. This release occurs when rainfall or irrigation water

interacts with a thin layer of surface soil (2.5 - 5 cm) and plant materials before leaving

the field as surface runoff (Sharpley, 1985). According to Sharpley et al. (1993),

dissolved P contributes 10 to 40% of the total (TP) transported from most cultivated

soils to water bodies through runoff and leaching. Most dissolved P is readily

bioavailable and thus increases algal growth in surface waters. Though sediment bound

P is not readily bioavailable, it can be a long-term source of P for aquatic biota (Ekholm,

1994; Sharpley, 1993).

The amount of P loss and its delivery rate from soils to surface water is influenced by

various factors and these factors will be discussed in detailed in the next section.

2.5.1 Factors affecting soil Phosphorus loss

The most influential factors that affect soil P loss and its delivery rate to water bodies

include the type of P source added (Kleinman et al. 2002), soil P levels (Sharpley, 1995;

Heckrath et al. 1995; Pote et al. 1996; Ige et al. 2005), soil physical and chemical

characteristics (Nearing et al. 1993), the amount and intensity of rainfall (Edwards and

Daniel, 1993), and the field slope and proximity to surface waters. Generally, highly

water soluble P sources increase the risk of dissolved P loss. For example, Anderson

and Magdoff (2005) have found that repeated application of P sources that are high in

soluble organic P (such as animal manures) to agricultural soils could release a

substantial amount of organic P to ground water. McDowell et al. (2001) have reported

29

that losses of P are related to soil P concentration and therefore strongly influenced by

P additions as fertilizers and manures. According to Pote et al. (1999b) and Sharpley

(1995), soils with high soil test P levels can contribute significant amounts of P to runoff

in the forms of dissolved P or particulate bound P. Because, when the binding sites of

soil particles are highly saturated in P, its capacity to retain additional P decreases and

the potential for losses of dissolved P increases through desorption and dissolution.

Several studies have found that particular soil physical and chemical properties such as

coarse texture, small concentration of Fe and Al oxides and large P concentration can

result in limited P adsorption and thereby could increase the potential for dissolved P

losses (Sharpley et al. 1996; Kleinman et al. 1999; Pote et al. 1999; Morel et al. 2000;

Pautler and Sims, 2000; Schoumans and Groenendijk, 2000). Gaynor and Findlay

(1995) showed that ortho-phosphate concentration and losses in drain flow from clay-

loam soil averaged 0.24 mg P L-1 and 0.38 kg P ha-1 yr-1 which represented

approximately 3% of the total P fertilizer applied. Fox and Kamprath (1971) concluded

that P leaching is primarily governed by the concentrations of Fe and Al sesquioxides.

Similarly, a substantial decrease in P leaching was reported for finer-textured soils and

soils which contain greater amounts of reactive Al. Accordingly, the potential for P to

leach in organic soils is greater than such potential in mineral soils which contain

greater amounts sesquioxides. Thus, it is evident that P sorption capacity of soils plays

a key role in reducing the potential for P losses through leaching.

For soils with large organic matter content and soils receiving large quantities of P

fertilizer, Sharpley et al. (1993), Eghball et al. (1996) and Elliott et al. (2002) have found

that phosphate can be leached considerably when soil phosphate sorption capacity is

30

saturated by fertilizer application. In a study conducted in an organic soil during one

study period, Miller (1979) reported that dissolved ortho-phosphate concentration and

loss from subsurface drainage water were as high as 18.2 mg P L-1 and 36.8 kg P ha-1

respectively. Accordingly, P leaching can be a significant problem in poorly drained soils

with high organic matter levels (Sharpley et al. 1994), soils with a long history of manure

application (Breeuwsma et al. 1995; Heckrath et al. 1995) and also in agro-ecosystems

characterized by soils excessively rich in P with small P sorption capacity (Cully et al.

1983; Gaynor and Findlay, 1995; Beauchemin et al. 1998; Sims et al. 1998; Smith et al.

2001b). Especially in flat fields, phosphate may be lost mainly by leaching from soils in

which the phosphate sorption capacity has been saturated by P fertilizer application.

Culley et al. (1983) also showed that more than 50% of the total P losses might be lost

through subsurface drainage water in flat plots of Ontario. Ozanne et al. (1961) found

that in sandy soils, which have small capacities for water retention and phosphate

buffering, the removal of P through leaching was as much as 80% of P applied. Similar

results were found by Humphreys and Pritchett (1971), in that all of the P applied as

inorganic fertilizer to sandy soils (>93% sand) leached to a depth below 50cm.

Generally P concentration of water percolating through the soil profile is small because

of P fixation by P-deficient sub-soils. However, exceptions can occur in sandy, acid

organic or peaty soils with low P fixation or holding capacities as well as in soils where

the preferential flows of water occur rapidly through macro pores and earthworm holes

(Bengston et al. 1992; Sharpley and Syers, 1979; Sims et al. 1998). Gaynor and

Findlay, (1995) found that, with artificially drained fine-textured soils that may not be P

31

saturated, but allow for rapid movement of P through preferential flow paths to

subsurface drains and then to surface waters.

Soil P loss through erosion is more severe in regions with intense rainfall and where the

soil on sloping land is not protected by a permanent cover of vegetation. Heavy

raindrops can detach both mineral and organic particles from the main body of soil. If

the volume of rainfall is sufficient for excess to run off over the surface, then the

detached particles are carried in the flowing water and can eventually reach water

bodies. Whenever the soil particles and organic residues are removed from a field

through soil erosion, the P adsorbed to these particles also moves out from the field.

Accordingly, management practices that prevent soil erosion (such as residue

management, cover crops, contour farming and no-tillage) could minimize surface

losses of P.

The potential of P losses in overland flow is also affected by the method of fertilizer or

manure application (Tabbara, 2003). According to Hansen et al. (2002) and Tabbara

(2003), surface applications of fertilizer or manure are extremely vulnerable to losses of

P into surface waters particularly when runoff occurs during or shortly after application.

This is because, most of the P applied to soils as fertilizers and manures remains within

the top 2.5 – 5 cm of a soil profile. Management practices such as incorporation,

injection or banding of fertilizer P or manure into the seedbed can reduce much of the

potential P runoff and make more P available to a crop.

The magnitude of P loss from soil increased with high rates of fertilizer and manure

application (Kleinman and Sharpley, 2003), and decreased with time after application

32

(Eghball et al. 2002) and with successive rainfall events (Kleinman and Sharpley 2003).

Phillips et al. (1981) have found that P concentration in surface runoff water was greater

following winter application of manure compared with similar applications in spring and

fall. Royer et al. (2003) found that the risk of water contamination would be less if

manure were applied in spring rather than in the fall.

Accordingly, pathways of soil P and forms of P that are lost from fields to neighboring

water bodies are governed by many soil factors, management practices and

environmental factors. Accurate measurements of soil P forms are imperative to make

decisions on environmentally sound P management. However, some difficulties are

reported when identifying these P forms. The next section will discuss laboratory

methods which are used to analyze and characterize soil P forms and their status.

2.6 Analysis of soil Phosphorus

Phosphorus is an extremely chemically reactive nutrient element. Estimating availability

of P is complex because soluble inorganic P can be immobilized by Fe, Al and other

metal oxides in soils (Lyamuremye et al. 1996; Griffin et al. 2003). Due to these different

reactions soil P exists in different inorganic (Pi) and organic (Po) forms, which control

the supply of labile P to the soil solution. These different inorganic and organic forms

and their quantities in soil can be estimated by extraction with acids and alkalis that

dissolve specific complexes binding P (Olsen and Sommers, 1982; Tiessen and Moir,

1993). The success of any extractant to estimate available P depends on its suitability

of the chemical used in relation to particular soil properties. However, it is unlikely that

33

an extractant could be found that would measure exclusively a single fraction of soil

inorganic P, although some components of extractants are aimed at specific P fractions.

There are different techniques for isolating specific soil P compounds, such as high-

performance liquid chromatography (HPLC), nuclear magnetic resonance spectroscopy

(NMR) and X-ray absorption near edge structure spectroscopy (XANES). However,

there are some financial and practical limitations with these methods. Again most P

analysis methods are not able to characterize all of the P in the soil and may also give

different results. As an alternative method, sequential fractionation approach has proven

useful for assessing soil P status in a wide range of soils. Sequential chemical

extraction procedures have been widely used to study the nature of soil P forms and to

quantify the availability of inorganic and organic P fractions (Chang and Jackson, 1957;

Bowman and Cole, 1978; Hedley et al. 1982b; Cross and Schlesinger, 1995). Based on

the differential solubility of the various inorganic P forms in various extracts, soil P can

be fractionated into different P fractions (Hedley et al. 1982b). The underlying

assumption in these approaches is that readily available soil P is removed first with mild

extractants, while less available or plant-unavailable P can only be extracted with

stronger acids and alkali. Residual P that remains after extracting the soil with stronger

acids and alkali represents recalcitrant inorganic and organic P forms.

In a sequential fractionation procedure, P fractions are functionally defined by the

extractants removing them from the soil. Resin-Pi represents inorganic P either from the

soil solution or weakly adsorbed on (oxy) hydroxides or carbonates (Mattingly, 1975)

and is the most biologically available P in soil (Amer et al. 1955; Sibbesen, 1977).

Alternatively, distilled water extraction (H2O-P) is used to measure a form of P

34

correlated to soluble P loss in runoff or leachate. Phosphorus extractable by NaHCO3

consists of weakly adsorbed inorganic P compounds (Hedley et al. 1982) and easily

hydrolysable organic P compounds such as ribonucleic acids and glycerol-phosphate

(Bowman and Cole, 1978). NaHCO3 extractable inorganic P fraction is considered to be

largely available to plants (Olsen et al. 1954). NaOH-Pi and NaOH-Po compounds are

held more strongly by chemisorptions to Fe- and Al- compounds of soil (Ryden et al.

1977; McLaughlin et al. 1977). NaOH-Pi is associated with amorphous and crystalline Al

and Fe (oxy) hydroxides and clay minerals. NaOH- Po is extracted from organic

compounds. NaOH-P is considered as moderately labile soil P and considered to be

slowly available to plants by desorption (Tiessen et al. 1984). Thus, both NaHCO3-P

and NaOH-P are considered to be available P forms and can contribute to plant-

available P (Tiessen et al. 1984). Hydrochloric acid extractable P fraction (HCl- P) can

be defined as stable P, which is an inorganic P in apatite-type minerals or in octa-

calcium P (Williams et al. 1971; Tiessen et al. 1984; Frossard et al. 1995). This P

fraction is generally assumed to be of limited availability to plants (McKenzie et al.

1992a). Residual P that remains after extracting the soil with HCl extractants represents

recalcitrant P forms, which include more chemically stable organic P forms and

relatively insoluble inorganic forms. The residual P can be further fractionated into Pi

and Po with an additional extraction of NaOH, to extract P held by the internal surfaces

of soil aggregates (Beck and Sanchez, 1994; Hedley et al. 1982a; Zheng et al. 2002).

The final residual P is determined after digestion with concentrated sulfuric acid and

hydrogen peroxide (based on the Thomas et al. 1967 method).

35

In general, fractionation methods do not quantify the organic P fraction by means of a

specific extractant. The fraction of total P remaining after inorganic P fraction has been

removed is taken as organic P (Barbanti et al. 1994). Therefore, the organic P fraction

of soils often is quantified as the difference between orthophosphate as detected by the

colorimetric acid molybdate method of Murphy and Riley (1962) before and after acid

digestion (e.g., Chardon et al. 1997). In Hedley’s fractionation procedure, two

measurements of NaHCO3 and NaOH extractions have to be conducted. One is to

determine total P content in the soil extracts after acid digestion of the filtered extracts.

The other one is to measure P contents after soil extract is acidified to precipitate

organic matter and P content in this acidified solution is considered as inorganic P.

Organic P in each extraction is calculated by the difference between inorganic P and

total P.

Many studies have used a sequential extraction technique to study the effects of

cropping systems and fertilizer application and other management practices on soil P

status in short- and long-term field experiments (Hedley et al. 1982a; Tiessen et al.

1984; Condron et al. 1985; Wagar et al. 1986; Schoenau et al. 1989; Beck and

Sanchez, 1994; Richards et al. 1995; Schmidt et al. 1996; Zhang and Mackenzie, 1997;

Zheng et al. 2001, Zheng et al. 2002: Zhang et al. 2004). Some studies have used this

technique to look into the changes of soil P during short-term incubation experiments

(Buehler et al. 2002; Daroub et al. 2000; Hedley et al. 1982a; Iyamuremye et al. 1996;

Qian and Schoenau, 2000). Although this technique is time consuming, the major

advantage of this technique is that soil P can be systematically quantified, thereby

providing a tool for investigating soil P transformations (Hedley et al. 1982a; Tiessen et

36

al. 1984; Condron et al. 1985; Schoenau et al. 1989). Another advantage is that the

sequential fractionation approach has also been proven useful for assessing P forms in

a wide range of soils (Beck and Sanchez, 1994; Nair et al. 1995; Sui et al. 1999).

To assess soil P status either for agronomic or environmental purposes, different soil

tests are used in different regions. The next section reviews the existing soil P tests

(agronomic and environmental), their limitations and identifies the improvements

needed for environmental soil P testing.

2.7 Soil Phosphorus Testing

Traditionally, the purpose of soil P testing is to identify the “optimum” soil test P

concentration that is required for optimal plant growth and crop production. Thus, the

soil test P values serve as a guide in making fertilizer recommendations for

economically optimum rates of P addition. Given this requirement, agronomic (routine)

soil P tests have been designed to measure the amount of P in soil available for crop

uptake. Since plant uptake of P could be influenced by soil characteristics such as type

of minerals present and soil pH, the soil P test suitable for one set of soil properties may

give erroneous results if applied to other soils. For example, the ammonium fluoride and

acid based soil tests of Bray and Kurtz (1945) are recommended for acid soils but not

for calcareous soils due to the dissolution of CaCO3 and precipitation of F by Ca

interfering with the extraction of P. Therefore, a large number of soil P extraction

methods that have been designed to account for various soil types and mechanisms

controlling the chemistry of soil P.

37

The most commonly used routine soil P tests in North America include Bray and Kurtz l,

Mehlich l, Mehlich 3, Kelowna and Olsen P (Bray and Kurtz, 1945; Mehlich, 1953, 1984;

Van Leirop, 1988; Watanabe and Olsen, 1965). Since the extracting agents used in

these tests have different abilities to extract different forms and different amounts of P,

the choice of an ideal method has been based on the regional understanding of soil

properties and the solubility of the forms of P prevalent in the region. As a general rule,

acidic soil extractants (Bray-P-1, Mehlich-3) are designed for acidic or non-calcareous

soils, where Al- and Fe- dominate P reactions, whereas alkaline extractants (Olsen) are

used for basic or calcarious soils, where Ca dominates soil P reactions. The form and

amount of soil P that is extracted by each extractant is determined by its solution pH

and the reaction of the ions present in the extractant. Using any of the chemical

extractants beyond the range of soils for which it was developed can result in the

buffering of acid or base extractants and dissolution of non-available P. Accordingly, the

reactions in chemical tests for soil P may solubilize non-labile P that are more tightly

bound to Al, Fe, or Ca complexes, which may not be plant available (Fixen and Grove,

1990; Mallarino, 1997). When this occurs accurate interpretation of test results becomes

difficult. Therefore, chemical extractants for soil P are not always equally reliable over

all soil types. Use of a given extractant is limited to specific soil types and test

interpretation is usually dependent on the region and crops grown.

The use of locally calibrated, routine soil P tests such as Olsen P or Mehlich-3 P, as

environmental risk indicators has significant practical advantages due to their well-

established data base (Sims et al. 2002). Accordingly, several studies have been

conducted in different regions to analyze the relationship between amounts of P

38

extracted by some environmental soil P tests and P extracted by agronomic soil P tests.

Their results in general demonstrate that the P values obtained by routine agronomic P

tests are well correlated with those of the environmental P tests (Mallarino, 1999; Atia

and Mallarino, 2002; Kleinman and Sharpley, 2002; Maguire and Sims, 2002). However,

the relationships varied with soil properties such as soil type, soil pH, particle-size

composition and mineralogy and management practices, which are known to influence

soil P sorption (Pote et al. 1996; Fernandes and Coutinho, 1997; Nuernberg et al. 1998;

Magdoff et al. 1999; Pautler and Sims, 2000). Therefore across regions with contrasting

soils, only a few generalizations can be made about the relationship between agronomic

and environmental P tests.

The amount of P in soil, sediment, and water that is potentially available for algal uptake

could be defined as bioavailable P (Sharpley et al. 2008) hence bioavailable P is directly

related to the risk of P loss from a given soil or site. It is important to have a close

correlation between routine soil P test values and the levels of bioavailable P to justify

the use of routine soil P tests to assess the bioavailable P. In fact several laboratory

studies have demonstrated that soils with high agronomic soil test P values are more

likely to have higher concentrations of soluble and bioavailable P (Sims, 1998; Sibbesen

and Sharpley, 1997). The above relationship between bioavailable P and the routine

agronomic soil test P has been further supported by Sims et al. (2002), Klatt et al.

(2003) and Andraski and Bundy (2003), who have shown that routine soil P tests can

provide useful estimates of total or bioavailable P concentrations in runoff. The evidence

for correlation between the P values from routine agronomic P tests and dissolved P

levels in runoff and drainage waters, as measures of bioavailable P, have been reported

39

by Daniel et al. (1994); Sharpley et al. (1994); Pote et al. (1996); Kleinman et al. (1999);

Sims et al. (2000); Cox and Hendricks (2000): McDowell et al. (2001), and Wang et al.

(2010, 2012). Mallarino (1999) found that soil P measured by routine soil tests (Bray-P,

Olsen-P and Mehlich-3-P as recommended for soils of the North Central region of US)

was well correlated with dissolved P in runoff water at 10-15 fold higher soil test P levels

relative to the optimum P amount needed for crop production. Many studies have

reported that P losses through erosion, surface runoff and leaching-lateral subsurface

flow are greater when soil test P values are above the agronomical optimum range

(Beauchemin et al. 1998).

Despite the evidence for correlation between routine soil P test and bioavailable P, a

number of researchers have raised concerns over the general suitability of routine soil

tests as a good diagnostic tool to monitor P levels in soils for environmental purposes.

For instance, Fang et al. (2002) found that Bray-1 P was least effective in explaining the

availability of dissolved P in runoff for calcareous soils. Allen et al. (2006) found a

similar relationship for typical USA Midwest soils. In addition, some agronomic soil test

P methods could not predict the potential P loss to surface water due to soil conditions.

For example, Mehlich-3-P has been shown to overestimate potential loss of soil P from

heavily manured soils due to formation of acid soluble Ca-P in soils following manure

applications (Sharpley et al. 2004). Torrent and Delgado, (2001) found that, at the same

soil test P reading with Olsen method, soils with a larger P sorption capacity release

less P than soils with a smaller P sorption capacity. Accordingly, the suitability of

agronomic soil test P for environmental predictions varies, depending on the nature of

the soil properties.

40

While the studies discussed above show promise in describing the relationship between

level of soil P and surface runoff P, as an indicator of potential environmental risk of

excess P in soil, these methods have several limitations. First, routine soil test

extraction methods were designed to estimate the availability of soil P to plants as the

basis for making recommendations for additional fertilizer application. This may not

accurately reflect soil P release to water in surface or subsurface runoff. Second,

although dissolved P is an important water quality variable, it represents only the

dissolved portion of P that is readily available for aquatic plant growth. It does not reflect

fixed soil P that may become available with time. Third, these routine soil extractants

are either more acidic or alkaline compared to the pH value of the soil solution.

Therefore, a portion of P extracted by routine soil extractants is actually of small

availability. For example, acidic extractants such as Bray-P1 or Mehlich-3 extractants

would likely dissolve Ca phosphates that are sparingly water soluble but which may not

be readily available for algal uptake (Self-Davis et al. 2000). Conversely, acidic

extractants may extract less available P relative to other P tests in many CaCO3-

affected soils due to the reactions of acidic extractant with CaCO3, (Atia and Mallarino,

2002). Accordingly, these routine soil P tests may not necessarily correlate well with P

loss potential or resulting potential for algae growth in surface waters.

Another concern with routine soil tests as indicators of potential environmental risk of P

is the typical depth of soil sampled. It is generally recommended that soil samples

should be collected to plow depth, usually 0-20 cm for routine evaluation of soil fertility.

However, the zone of interaction of runoff waters with most soils is normally less than 5

cm and this sampling depth is important when conducting soil tests to estimate the

41

potential for P losses. Consequently, different sampling procedures may be necessary

when using a soil test to estimate the potential for P loss.

As noted above, the relationships between routine soil test P and potential for P losses

are conditioned by a variety of soil properties. Therefore it is difficult to assign

universally acceptable routine soil test P values for environmentally sustainable

management systems. However, the use of mild extractants such as distilled water and

dilute salt solution (e.g. 0.01M CaCl2 and NH4Cl) overcomes the problem of location

dependency of the readings of soil test P, because these extractants have an ionic

strength similar to most soil solutions (Van der Paaw, 1971; Self-Davis et al. 2000;

Racz, 1979; Pote et al. 1995; Sotomayor-Ramirez et al. 2004). Since their use is not

limited to specific soil types, such extractants can better predict the potential for P

losses to surface and ground water compared to agronomic soil P tests. Furthermore,

McDowell and Sharpley (2001) found that water extractable P had the strongest

relationship with P lost via overland flow, while CaCl2 extractable -P had the strongest

relationship with P lost via leaching.

In addition, the extraction of soil P with distilled water provides a rapid (usually 1h

extraction time), low cost and simple method of determining the amount of soil P that

can be released from soil to runoff water. Distilled water extraction method is based on

weak desorption reactions and assumes that extraction with water replicates the

reaction between soil and runoff water. This extraction maintains the soil pH within one

unit of its original value, a desirable attribute since P solubility is highly dependent upon

soil pH (Golterman, 1988; Sharpley, 1993). Recently, distilled water extraction method

has been extensively used to study soil P effects on dissolved reactive P concentrations

42

in runoff. For example, in a field study with tall fescue, Pote et al. (1996) found an

excellent correlation between water extractable P and dissolved reactive P

concentrations in runoff. However, there are some limitations; the amount of P extracted

by distilled water (mainly P in dissolved forms) is very small for most soils, and may not

reflect all forms of labile P. The difficulties related to chemical analysis of small amounts

of soil P also limit the use of water as an extractant. Thus, as an alternative to distilled

water extractant, the ion-sink methods can be used to measure bioavailable P in soil.

Ion sink methods that rely on sorption-desorption reactions could provide different

estimates of soil P available for plant growth compared to tests based on chemical

extractive solutions. Since ion sinks only adsorb P in soil solution onto the sink surface,

they interact minimally with the soil and do not alter the soil conditions. Since the ion

sink methods operate with negligible amount of chemical extraction, they mimic

rhizosphere conditions more closely, acting more or less analogous to the withdrawing

behaviour of a plant root. Accordingly, ion sink methods have an advantage over routine

chemical extractant methods such as Bray-1, Olsen, and Mehlich-3, because they

adsorb available P ions from the in situ labile P fractions in soil (Menon et al. 1989).

Previous studies have indicated that, ion sink methods often provide similar or better

correlations with crop responses to P compared to such correlations with chemical

extractants (Sharpley et al. 1994a). In pot experiments with canola, Qian et al. (1992)

reported that, the resin membranes have provided a better index of P availability than

conventional chemical extraction methods. Similar results were reported by Saggar et

al. (1992) for rye grass. Therefore, ion sink methods have been favorably employed to

estimate plant-available P for soils with large variations in physical and chemical

43

properties (Menon et al. 1990; Sharpley et al. 1994). Especially where fertilizer history is

unknown and frequent changes in fertilizer type may have been made, it is difficult to

choose the appropriate soil test method. However, ion sink methods have provided

accurate estimates, irrespective of management history (Sharpley, 1993; Yang et al.

1991; Qian etal., 1992; Somasiri and Edwards, 1992).

The most promising ion sink methods are: iron-oxide impregnated filter paper strip

(FeO-strip) (Chardon et al. 1996) and Anion Exchange Resin Membrane (AEM)

methods (Abrams and Jarrell, 1992; Qian et al. 1992; Saggar et al. 1992). The FeO-

strip method has a stronger theoretical justification as a reliable estimate of bioavailable

P than do chemical extractants, since the strips act as an “infinite sink” to measure

desorbable soil P, and thus measured the potential of a soil to continue to release P

during a runoff or leaching event (Moore et al. 1998). Research has shown that FeO-

filter paper strip method effectively estimate plant available P in a wide range of soils

and management systems (Menon et al. 1989, 1990; Sharpley, 1991). Several studies

have shown that the amounts of P extracted by standard soil tests (Bray, Mehlich and

Olsen) are correlated with bioavailable P estimated by FeO-strip. For example, Sharpley

(1996) found a consistent increase in FeO-strip P with increasing levels of soil test P

(Mehlich-3) as a result of long term applications of beef, poultry and swine manures.

Barberris et al. (1996) found significant positive correlations between several soil P tests

and FeO-strip P (r = 0.62 to 0.89) for over-fertilized European soils. Pote et al. (1996)

found that FeO-strip P method accurately predicted the quantity of P susceptible to

runoff, better than most agronomic soil P tests. Sharpley (1993) also observed that the

FeO-strip P content of runoff was closely related (r2 = 0.92-0.95) with the growth of

44

several algae species where runoff was the sole source of P and therefore Fe-oxide

strip P was a good indicator of the biological availability of P in runoff waters to algae.

However, there are some limitations with the use of FeO-coated papers. Papers coated

with FeO are not commercially available. This has led to various methods for their

preparation and use (Myers et al. 2005). Another concern is contamination of the FeO-

coated papers with fine soil particles during shaking period (Chardon et al. 1996; Myers

et al. 1995) which can lead to error in estimating desorbable P (Uusitalo and Yli-Halla,

1999). This can however, be minimized by the use of CaCl2 solution as the background

electrolyte which tend to minimize soil dispersion (Myers et al. 2005). But this can lead

to reduction in the amount of P extracted (Koopmans et al. 2001). With all these

disadvantages of the FeO-coated paper method, AEM method was developed to

simplify problems associated with ion-sink extraction of soil P.

The behaviour of AEM resembles the action of plant roots (Raven and Hossner, 1993),

adsorbing P from the soil solution and releasing counter ions (Qian et al. 1992;

Sibbesen, 1983; Tran et al. 1992).The adsorption rate of the AEM is governed by P

desorption rate from the soil solid phase (Schoenau and Huang, 1991; Cooperband and

Logan, 1994). Therefore, the amount of P adsorbed depends on the relationship

between exchange capacities of the soil solid phase and AEM and the duration of

contact between these two. Qian et al. (1992) found that, an extraction time as short as

15 min can be used without reducing the accuracy of predicted P availability for a wide

range of soils. The AEM can act as sinks or as exchangers, depending on their contact

environment (Cooperband and Logan, 1994).

45

The major advantage of the resin extraction method is the capability to extract P from a

variety of soil types irrespective of the properties of the soil (Sharpley et al. 1994).

Durability of AEMs is another economically desirable feature of this method. They can

be recycled a large number of times without losing their physical and chemical

properties (Cooperband and Logan, 1994). Reports indicate that AEM have been re-

used as many as 50 to 500 times without losing their extraction efficiency or showing

detrimental structural effects (Saggar et al. 1990; Schoenau and Huang, 1991). This

feature makes it relatively a cheaper test compare with the use of FeO-coated papers

that can be used only once. Therefore, multiple uses would be a distinct advantage of

AEM over FeO-strip methods.

However, research with resin-based P tests (Fernandes and Coutinho, 1997; Nuernberg

et al. 1998; Fernandes et al. 1999) has shown that relationships between soil P

extracted and indices of P availability are affected by soil properties (soil pH, particle

size composition, and mineralogy) which are known to influence soil P sorption.

Therefore few generalizations are possible across contrasting soils. Since then, there is

a need to develop a suitable soil P testing method to predict soil P losses that can be

used for soils with a wide range of soil properties, so that an environmental plan can be

programmed and BMPs developed.

2.8 Research Issue

Public demand for improved management of P in agricultural systems is linked to the

threat of eutrophication from P pollution as a result of increased applications of

phosphate fertilizer and livestock manure to agricultural lands. Consequently, there

46

have been determined efforts by soil scientists and other researchers to develop

strategies to reduce P losses from agricultural non-point sources. To a large extent,

these strategies depend on the accurate measurement of forms of P in soil, water and

residual material which are often seen as a source of surface water P.

A key component of remedial strategies to minimize P losses from agricultural non-point

sources to aquatic environments is the determination of soil P levels that exceed the

optimum levels required for crop growth. The possibility of soil P levels that exceed the

optimal level for crop growth is dependent on the soil P dynamics, which in turn is

determined by various soil management practices. A better understanding of the long

and short term changes in soil test P values as a result of widespread soil management

practices such as applying dairy or hog slurry, incorporation of composted livestock

manure and inorganic P fertilizer application is imperative to develop remedial strategies

to control the potential P pollution and managing P more efficiently for the sustainable

management of temperate agricultural soils. One of the related research issues

explored in the extant literature is the suitability of agronomic soil P test methods to

evaluate the P polluting potential of a soil and whether environmental soil P tests are

better suited to evaluate that threat.

This study contributes to our ability to evaluate the impact of various soil management

practices on the long and short term soil test P values. The study evaluates impacts of

long-term cropping and mineral fertilizer and livestock manure applications on changes

of soil inorganic and organic P fractions. It relates these changes to the amounts of P

extracted by agronomic and environmental soil testing procedures to establish a

47

relationship between fertilizer and livestock manure P loading to the long-term potential

and risk of losing P from the soil to aquatic environments.

48

Chapter 3: Immediate and residual effects of different forms of swine manure on

soil phosphorus fractions in a clay loam soil under corn-soybean

rotation

3.1 Abstract

Understanding P dynamics in soils applied with different forms of manure is useful for

minimizing negative impacts on environmental water quality, in addition to improving

crop use efficiency of P. The objective of this study was to evaluate both the immediate

(year of application) and the residual (following year) effects of various forms of swine

manure on soil P fractions in comparison with triple super phosphate (TSP) under a

corn-soybean rotation. A field experiment was conducted on a Brookston clay loam soil

in south-western Ontario, Canada. Treatments were three forms of swine manure [liquid

(LM), solid (SM), compost (MC)] and TSP, which were applied at the rate of 100 kg P

ha-1 only to the corn phase, and the no-P control (CK). Soils were sampled at post-

harvest stage in both corn and soybean phases and P was analyzed using a modified

Hedley’s sequential fractionation procedure.

In the corn phase, all P treatments significantly increased the soil labile (17.24 and

34.36 mg P kg-1 of H2O-P and Bicarb-Pi, respectively) and moderately labile (29.05 mg

P kg-1 P) inorganic P (Pi). The effects from all three forms of manure on a given P

fraction were similar to the effects of TSP. Total-Pi and Total-P (sum of all P forms) were

significantly increased by all P treatments, while total-Po (sum of all organic P forms)

was not. In the following year, only some of the treatments had residual effects on some

of the P fractions (MC and SM on H2O-P, all four P treatments on Bicarb-Pi, SM on

Bicarb-Po and MC, SM and TSP on NaOH-Pi). In both phases of cropping, none of the

49

P treatments affected HCl-Pi or stable-P forms. In this study, increased soil labile P

levels associated with all P sources indicate the potential for soil P loss through leaching

and runoff to the aquatic environment with all three forms of swine manure applications

and TSP.

3.2 Introduction

Swine manure contains nutrients which are essential for plant growth, and when it is

applied to the soil at proper rates, can serve as an excellent source of essential plant

nutrients. However, there are uncertainties regarding its availability of nutrients and

cost-effectiveness for crop production while minimizing the adverse environmental

effects to water resources. Reasons for this uncertainty include large variability of P

concentrations in manure and limited research data about its effects on soil P dynamics.

In manure, P is found in both inorganic (Pi) and organic (Po) forms (Mikkelson, 1997),

yet in general these are not fully available for plant uptake. The availability of Pi in

manure is determined by the form of Pi and the capacity of the soil to precipitate or

adsorb that Pi. The Po fraction in manure is not directly available to plants. It must be

mineralized or converted into Pi forms via soil microbial activities over time before plants

can use them. Thus, the availability of Po in manure depends on the rate of

mineralization.

The proportions of organic and inorganic P in manure vary greatly. In the literature,

Brookes et al. (1997) estimated that inorganic P comprises approximately 80% of the P

in liquid hog manure with the rest present in organic form. This figure agrees with the

result of 70-90% as Pi recorded by Schoumans and Groenendijk (2000). However,

50

Mikkelson (1997) estimated that generally up to 50% of the total P in swine manure is in

the organic form. This figure is compatible with the 49% organic P in swine manure

measured by He and Honeycutt (2001). These disparities are probably due to the

variation of manure P contents with the different manure types.

The composition of manure varies considerably due to animal physiology (type and age

of animal), feed rations and additives, type and amount of bedding, handling and

duration of storage (Lindley et al. 1988). Liquid manure (slurry) and solid manure differ

in their dry matter content, N, P and C contents along with their influence on microbial

activity and chemical changes in the soil (N’Dayegamiye and Cote, 1989). Slurries

consist of excreta produced by livestock while in a yard or building, which are mixed

with rainwater and wash water and some cases, waste bedding and feed. Therefore the

solid content of the slurry is usually low, ~ 1- 5 %. Solid manures include farmyard

manure and materials from covered straw yards, excreta with a lot of straw in it, or

solids from mechanical slurry separators. Due to its large water content, up to 99% (wet

basis) for liquid manure and 70% for solid manure, long-range haulage of fresh manure

from a livestock operation is uneconomical. However, concentrated livestock operations

often have an abundance of liquid and solid manures relative to the land on which the

manure can be spread. Accordingly, composting manure has become of great interest in

areas where surpluses of manure are found. Composting reduces volume, mass

(Larney and Hao, 2007) and moisture content of manure (usually less than 35% by

mass), and may increase the uniformity of manure (Rynk et al. 1992). These properties

reduce the transport cost and make it easier to spread the material uniformly.

Composting manure can also provide several potential advantages over the use of fresh

51

forms of manure, such as a considerable reduction in loading of pathogens, parasites,

weed seeds and the level of odour associated with the land application (Eghball and

Lesoing, 2000; Menalled et al. 2002).

Composting of manure generally results in greater P concentrations, because P is

conserved while CO2 and NH3 losses result in 30-50% reduction in mass of C and N

(DeLuca and DeLuca, 1997). As consequence, composting generally increases the P

content but reduces the N: P ratio. Gagnon and Simard (1999) reported that composting

manure decreases the extractability and plant availability of P in the manure due to

immobilization. However, the extent of these effects varies greatly with the source of

compost material and with the method of compost management.

Composting has gained increased attention as a means of reducing the environmental

impact of livestock manure (Kashmanian and Rynk 1996, 1998), because during

composting, manure nutrients are converted to more stable forms such as and are less

likely to reach groundwater or move in surface runoff. However, there are some

disadvantages of composting manure, such as nutrient (especially N) losses,

greenhouse gas emissions and the extra cost of labour, equipment and space

necessary for composting. Furthermore, since composting increases P content of

manure on a dry matter basis (Ott and Vogtmann, 1982), the greater concentration of P

in compost may create environmental problems if applied to soils that have large

background levels of P.

Currently, agricultural producers and researchers explore the suitability of different

forms of livestock manure, including fresh slurry, solid manure and composted solid

52

forms to satisfy the crops’ needs while minimizing adverse environmental impacts to

water resources. Some research suggests that fresh manure application may modify the

magnitude of different soil P fractions. For example, the greatest net increase in resin-P

and labile-P (in terms of percentage of total P added) occurred after mixing soil with

fresh dairy manure (Gagnon and Simard 2003). These same authors also reported that

the moderately labile P fraction was the most abundant form of P found in acidic soils

following the addition of manure compost. Similarly, a very rapid increase in anion-

exchange membrane extractable P was found by Simard et al. (2001) after the

application of liquid hog manure in a calcareous gleysolic soil in Quebec. A significant

increase in labile Pi and Po fractions was measured after 4 years of liquid dairy manure

addition to a silty clay soil (Zheng et al. 2001). However, a single application of liquid

hog manure was found to have little impact on labile P levels (Qian and Schoenau,

2000). According to these studies, the forms and availability of P in soil following

manure additions are dependent to a large extent on the form of manure applied.

Since manures vary in the degree of P availability and thus their immediate and residual

impacts on soil P forms and levels, detailed measurements of soil P forms, their

distribution and changes with time can provide important information to address the

present concerns with P contamination of surface waters. The objective of this research

was to improve the understanding of immediate effects in the year of application and the

residual effects in the following year of applying different forms of swine manures in

comparison with inorganic P fertilizer on soil P forms.

53

3.3 Methodology

3.3.1 Site Description

The field experiment was conducted at the Eugene Whelan Research Farm of

Agriculture and Agri-Food Canada at Woodslee (42013’N, 82044’W), Ontario, Canada.

The site is characterized as humid and cool-temperate, with a mean annual air

temperature of 8.70C and mean annual precipitation of 827 mm. The soil type is

Brookston clay loam soil classified as fine loamy, mixed, mesic, Typic Argiaquoll or

Orthic Humic Gleysol.

3.3.2 Treatments, soil sampling and analysis

There were five treatments selected for this study, including four P sources: solid swine

manure, liquid swine manure, swine manure compost and inorganic fertilizer P as triple

super phosphate (0-46-0), and the control with zero nutrients (control) applied. Liquid

and solid swine manures were obtained from two local pig producers near Harrow,

Ontario and composted swine manure was obtained from Ridgetown Campus,

University of Guelph in Kent County, approximately 136 km east of Woodslee, Ontario.

Samples of each manure type were collected, chilled to 4°C and analyzed prior to

application.

First, the collected manure samples were dried at 105°C to determine moisture content.

Fresh manure samples were used for the determination of nutrient contents. Carbon

and total N were measured using a Leco Analyzer. The samples were digested with

H2SO4-H2O2 for total N, P, and K. A 10 g aliquot of hog manure was added with 100 ml

of 2M KCl, and shaken for one hour. The mixture was filtered with vacuum suction and

54

NH4-N and NO3-N were determined colorimetrically. The physical and chemical

compositions of the various types of swine manure are given in Table 3.1.

Each study plot was 9 m wide by 25 m long. The cropping system was a corn (Zea

mays L.) soybean (Glycine max L.) rotation. Corn (2006) and soybean (2007) were

seeded at local recommended rates of 76852 and 373132 seeds ha-1, respectively

(OMAFRA, 2002). Pesticides were applied for both corn and soybean production.

Treatments were replicated three times in a randomized complete block design.

Treatments were applied at a rate of 100 kg P ha-1. Additional N and K were applied to

all treatments as inorganic fertilizers to satisfy crop needs according to the OMAFRA

(Ontario Ministry of Agriculture, Food and Rural Affairs) recommendation and taking

account of available N and K contents in the manure. Manures and inorganic fertilizer

were applied using the broadcast-incorporation method in the spring prior to planting the

corn. Manures were spread on the soil surface and incorporated immediately, or once

soil conditions allowed, by disk, triple-K tiller and packer.

Soil sampling was carried out at post-harvest stage. Sixty randomly selected soil cores

were taken from the surface soil (0-7.5 cm depth) in each plot using a standard hand

soil probe with 2.54 cm internal diameter; soil samples from each plot were pooled to

produce a composite surface soil sample for laboratory analysis. The samples were air-

dried. The pipette method was used for analyzing soil texture (Gee and Bauder, 1986).

The organic C and total N were analyzed by dry-combustion (Leco CNS-1000 Analyzer,

Leco Corp., St-Joseph, MI). Soil test P was estimated as Olsen P (Olsen and Sommers,

1982) and ammonium acetate (NH4OAC) extractable K, Ca and Mg (Knudsen et al.,

55

1982) were determined. Soil pH was determined in a soil/water ratio of 1:1. The physical

and chemical properties of the soils used in this study are given in Table 3.2.

3.3.3 Hedley sequential Phosphorus fractionation:

Soil samples were sequentially extracted by the modified Hedley’s P fractionation

procedure to quantity the different soil Pi and Po fractions (Hedley et al. 1982),

(Figure.3.1). Phosphate was determined colorimetrically with the molybdate-ascorbic

acid procedure (Murphy and Riley, 1962) using a QuickChem Automated Analyzer

(Lachat Instrument, Milwaukee, WI).

To calculate the recovery of P in the sequential extraction, total soil P was determined

using a separate soil sample digested with H2SO4-H2O2 (Thomas et al. 1967), followed

by colorimetric measurement using a QuickChem Automated analyzer (Lachat

Instruments, Milwaukee, WI). The results indicated that the modified Hedley’s

sequential fractionation procedure had a good recovery range of 97-102 % of the total

soil phosphorus of this study.

56

0.5 g soil

Shake 16 h in 30 ml distilled and de-ionized water,

centrifuge, pass through 0.45 µm filter filtrate for H2O - Pi & Po

Shake 16 h in 30 ml 0.5 M NaHCO3, pH 8.5,

centrifuge, pass through 0.45 µm filter filtrate for Bicarb - Pi & Po

Shake 16 h in 30 ml 0.1M NaOH filtrate for NaOH-1 Pi &Po

centrifuge, pass through 0. 45 µm filter

Shake 16 h in 30 ml 1.0M HCl filtrate for HCl - Pi

centrifuge, pass through 0.45 µm filter

Shake 16 h in 30 ml 0.1M NaOH filtrate for NaOH-2 Pi & Po

centrifuge, pass through 0.45 µm filter

Digest residue in 5 ml of

concentrated H2SO4 + H2O2 Residual P

at 360°C for 3h, for Pt analysis

Figure.3.1 Modified Hedley sequential fractionation procedure for soil phosphorus (Pi,

and Po refer to inorganic and organic P, respectively)

57

3.3.4 Statistical Analysis

To evaluate the effects of different sources of swine manure on soil P fractions, Proc

GLM was carried out using the SAS 9.3.1. To identify the effect of treatment, amounts of

P removed by the various extracting agents were subjected to ANOVA. Subsequently,

treatment means were subjected to multiple mean comparisons to identify statistically

significant differences (SAS Institute Inc. 2001).

3.4 Results and Discussion

The distributions of Pi and Po fractions in the surface soil (0-7.5 cm) for both years (corn

and soybean) are illustrated in Figures 3.2 to 3.8. The contributions from each P fraction

to the total soil P (Total-Pt) fraction in 2006 and 2007 are given in Table 3.4. Analysis of

Variance (ANOVA) of the effect of treatments, time and the interaction (treatment * time)

on different soil P fractions are given in Table 3.5. The mean differences and their

significant levels of the P fractions from 2006 to 2007 are given in Table 3.6.

Results of this study show that addition of different forms of swine manure and inorganic

fertilizer P to the clay loam soil has influenced soil P fractions differently. Some of soil P

fractions were significantly affected by some of the treatments during the first cropping

season (2006). In addition, there were significant residual effects on some of the soil P

fractions in the following cropping season. Overall, changes in soil P concentrations

were found in most of the soil labile and moderately labile P fractions and such changes

varied by the treatments. These changes in the P fractions will be discussed in detail in

the following sections.

58

3.4.1 Labile and moderately labile inorganic P (H2O-P, Bicarb-Pi and NaOH-1-Pi)

The P extracted by water (H2O-P) represents most labile P fraction of the soil. This P

fraction represents the freely exchangeable soil P, and may also be the most

appropriate environmental estimator of P concentrations in runoff, because water is the

solvent and transport medium for P loss from soils with runoff (Pote et al. 1996).

Under corn in 2006, addition of P from all forms of swine manure and inorganic fertilizer

treatments significantly increased the H2O-P fraction (soil solution P) compared with the

CK (Figure 3.1-a). The average concentration of H2O extractable -P in all P treated soils

was 24.51mg P kg-1, showing 17.24 mg P kg-1 increase (i.e. 237%) with the addition of

organic and inorganic P sources compared to the CK (7.27 mg P kg-1). In addition, the

similar H2O-P levels in soils treated with different forms of swine manure and fertilizer P

indicates, the effects on H2O-P were similar from organic and inorganic treatments

(Table 3.3). These results are comparable with the results reported by Du et al. (2011)

that indicated increased levels of soil solution P concentrations in soils treated with both

compost and inorganic fertilizer P. However, they observed that the soil solution P level

in the soils amended with compost was significantly greater than that in the soils treated

with inorganic fertilizer P. They suggested that the larger H2O-P concentration in

compost-treated soils could be related to the greater organic matter content of compost

(Turner et al. 2007). Organic matter may promote movement of P from soil solid phase

to soil solution phase, because organic matter releases large amounts of organic acids

when decomposed by microorganisms. These organic acids can complex Fe, Al and Ca

and thus mobilize sorbed P. In addition, organic compounds in manure directly compete

with P for the sorption sites (Guppy et al. 2005), and could prevent P from adsorbing to

59

soil mineral surfaces, Al and Fe oxides, or carbonates (Cross and Schlesinger, 1995;

Celi et al. 1999; Wandruszka, 2006) and thereby increase P concentration in soil

solution.

Over the two year study period, H2O-P levels were significantly less under soybean

(2007) than under corn (2006) in all the treatments including the CK (Table 3.6).

Possible explanations for such decreases over the two-year period include plant P

uptake, adsorption of P onto soil components that is not water extractable and loss of

this most soluble P fraction from the surface soil through runoff and leaching.

The H2O- P levels in the following year (2007) were taken as the residual effects of the

treatments. In 2007, the average concentration of H2O-P in the soils that received P in

2006 was 10.56 mg P kg-1. There were no significant differences in the H2O-P levels

between fertilizer and manure treated soils (Table 3.3). However, significantly higher

levels of H2O-P were found in soils which were treated with MC (13.19 mg P kg-1) and

SM (10.82 mg P kg-1) in the previous year, compared to the CK (6.06 mg P kg-1). In

addition, significantly higher H2O-P levels were found in soils treated with MC in the

previous year compared to the LM (8.70 mg P kg-1) treated soils (Table 3.3). The higher

H2O-P level in the MC and SM treated soils might be related to the influence of the large

organic matter content associated with MC and SM treatments. The organic acids that

were released from MC and SM decomposition could effectively reduce P sorption to

the soil and increase soil solution P concentration (Turner et al. 2007). Furthermore, the

increased organic matter in the soils treated with MC and SM perhaps covered clay

mineral surfaces and/or chelated metal ions (Fe, Al etc.). Thus, higher level of organic

matter content may prevent Pi from adsorption to clay minerals or precipitation with

60

metal ions (Tang et al. 2006), leading to higher available P in the soil solution. However,

in soybean phase in 2007, H2O-P levels of LM (8.70 mg P kg-1) and TSP treated soils

(9.54 mg P kg-1) were not significantly different from CK (6.06 mg P kg-1) (Table 3.3).

Given that the P in LM and TSP treated soils was more soluble and readily plant

available than P in other treatments (Mallarino et al. 2005), a greater uptake of P by

both corn and soybean crops within these two cropping periods would tend to decrease

the H2O-P levels in LM and TSP treated soils, leaving less residual P in the surface soil.

Overall, in this study, H2O-P fraction made a minimal contribution to soil total-Pt fraction;

in a range of 1- 4% in the corn phase (2006) and 1- 2% in the soybean phase (2007)

where the lowest contribution was from the CK and the highest contribution from MC

treated soils. These values are in line with Hooda et al. (2001) who reported that soil

water-extractable P fraction represented only 1.3% of soil total-Pt.

Sodium bicarbonate extractable P (Bicarb-Pi) fraction represents a major labile-P

fraction, which is plant available and may be subjected to losses through surface runoff

and leaching (Hedley et al. 1982). Bicarb-Pi fraction consists of Pi adsorbed on to

surfaces of some crystalline P compounds, sesquioxides or carbonates (Tiessen and

Moir 1993). In 2006 as well as in the following year, Bicarb-Pi fractions followed a

similar pattern as that of H2O-P (Figure 3.2-b).

In both years, significant increases in Bicarb-Pi levels were detected in all the P treated

soils compared to CK (Table 3.3). In corn phase, the average concentration of 42.75 mg

P kg-1 of Bicarb-Pi in P treated soils showed a 34.36 mg P kg-1 increase (409%)

compared to the CK (8.39 mg P kg-1). Among these P treatments, the highest Bicarb-Pi

61

level was found in the MC treated soils (47.92 mg P kg-1 and 25.32 mg P kg-1 in 2006

and 2007 respectively) and the lowest Bicarb-Pi level was found in the LM treated soils

(37.95 mg P kg-1 and 20.48 mg P kg-1 in 2006 and 2007 respectively) compared to the

CK (Table 3.3). In addition, Bicarb-Pi fractions increased in fertilizer treated soils in a

similar manner to that of all three forms of swine-manure treated soils for both years.

Thus, there were no significant differences of Bicarb-Pi levels found among the four P

treatments (Table 3.3).

However, over the two-year study period, Bicarb-Pi fractions significantly declined for all

the P treated soils including the CK (Table 3.6). Accordingly, in soybean phase in 2007,

the average concentration of Bicarb-Pi in P treated soils was reduced to 22 mg P kg-1.

However, this residual P level was still 15.25 mg P kg-1 higher (i.e. 226%) than that of

the CK (6.75 mg P kg-1). Similar to the decreases in the H2O-P fraction, the decreases

in Bicarb-Pi fractions also could be owing to plant P uptake and the removal of P from

surface soil layer through runoff and leaching. In addition, some of the Bicarb-Pi may

have converted into other P fractions. In the corn phase in 2006, Bicarb-Pi fraction

represented about 6% of the soil total-Pt fraction from the P treated soils and about 1%

of the soil total Pt fraction from the CK. However, in the soybean phase in 2007, Bicarb-

Pi fraction contributed only about 3.5% of the total-Pt from the P treated soils and 1%

from the CK plots (Table 3.4).

On average, labile P (H2O-P and Bicarb-Pi) fraction represented 9% of total-Pt from P

treated soils and 2% of total-Pt from the CK. These results suggest that greater labile P

levels in the surface soil compared to the CK may have directly resulted from the

accumulation of P in the surface soil when P is added as swine manure (SM, MC, and

62

LM) or as inorganic fertilizer P. Similar findings that manure applications have led to

increase in labile Pi levels were reported by Abbott and Tucker (1973), Campbell et al.

(1986) and O’Halloran (1993). Hao et al. (2008) and Zhang et al. (2004) have reported

that additions of manure and/or fertilizer to soil increased the soil test P values if the

additions were more than crop removal.

Since P has relatively low mobility in soil, most P accumulation takes place in the

surface soil. A significant accumulation of resin-Pi (P fraction which is similar to the H2O-

P fraction in sequential fractionation) in surface soil layer after applying liquid swine

manure in a Le bras silt loam soil (Gleysol) was reported by Hountin et al. (2000).

Similarly, significant increases in labile P (H2O-Pi, Bicarb-Pi and Bicarb-Po) fractions

after feedlot manure applications in a surface loamy soil (Dark Brown Chernozemic)

were observed by Dormaar and Chang (1995). Sutton et al. (1986) also reported that

extractable P concentrations in soils increased with manure and fertilizer applications

and a majority of the P accumulated in the 0-15 cm depth of the soil profile. They further

explained that, of all the nutrients applied to the soil with manure, the accumulation of P

was the greatest and, P was the only nutrient in abundant supply after the residual year

of cropping. This was owing to the inherent characteristics of P being sorbed to the soil

with little downward movement by leaching.

After sodium bi-carbonate extractable-P, the next extractable P fraction is the

moderately labile P fraction (NaOH-1-Pi). This is the source of soil labile P when soil

available P is depleted by crop uptake (Zhang et al. 2004). The moderately labile P

fraction represents the soil Pi that is associated with amorphous and crystalline Al and

Fe phosphates (Bowman and Cole 1978; Parfitt 1978) via chemisorption (Williams et al

63

1980; Hedley et al. 1982; Tiessen and Moir 1993). This P fraction is not readily available

for plant P uptake compared to labile P (H2O-P or Bicarb-P) fraction.

Similar to the labile P fraction, in corn phase 2006, the NaOH-1-Pi fractions in soils

treated with all P sources were significantly increased compared to the CK (Table 3.3

and Figure 3.3-d), indicating considerable amounts of P accumulation in moderately

labile Pi fraction from all P sources. The average concentration of moderately labile Pi

was 56.11 mg P kg-1 for P treated soils compared to the CK (27.06 mg P kg-1), showing

29.05 mg P kg-1 average accumulation (i.e.107% higher than CK) with P addition in corn

phase.

Over the two-year period, NaOH-1-Pi levels significantly decreased in TSP, LM and SM

treated soils (Table 3.6 and Figure 3.3-d), indicating that considerable amounts of P

were moved out from the moderately labile-Pi fraction in TSP, LM and SM treated soils.

These decreases could be due to the replenishment of labile-Pi fraction, which tends to

be decreased by greater crop P uptake, or may be due to the transformation of

moderately labile Pi into more stable P fractions.

As a residual effect in the following year, 2007, significantly higher NaOH-1-Pi levels

were observed in all the P treated soils except LM compared to the CK (Table 4.3 and

Figure 4.3-d). The average concentration of NaOH-1-Pi in 2007 was 42.06 mg P kg-1 for

P treated soils compared to the CK (24.75 mg P kg-1). Therefore, as a residual effect of

P treatments, on average 17.3 mg P kg-1 P increase was observed in moderately labile

Pi pool, (i.e. 70% higher) compared to the CK. However, no significant difference was

found for NaOH-1-Pi levels between TSP and manure treatments (Table3.3 and Figure

64

3.3-d). The contribution to the soil total-Pt fraction from NaOH-Pi fraction is in a range of

4- 8% in both crop phases, where the lowest contribution was found to be in soils of the

CK and the highest contribution from MC and SM treated soils (Table 3.4).

These results indicate that labile (H2O-P + Bicarb-Pi) and moderately labile (NaOH-1-Pi)

Pi fractions had significant treatment effects from both inorganic and organic P sources

in both corn and soybean phases compared to the CK. These results are comparable

with the findings reported by other authors from studies involving a wide range of soils.

Many have reported increases in the Bicarb-Pi and moderately labile-Pi fractions with P

applications (Ciampitti 2011; Picone et al. 2007; Wang et al. 2007; Verma et al. 2005;

Guo et al. 2000; O’Halloran 1993; Selles et al. 1995).

Although the amounts of moderately labile Pi were greater than those of Bicarb-Pi,

these two fractions followed a similar pattern of distribution among the treatments where

the highest increase was with the MC treatment followed by SM, TSP and LM

treatments (Table 3.3). The relationship between labile and moderately labile P with a

similar pattern of distribution among treatments is in agreement with the known fact that

available soil P pools are constantly replenished through reactions of dissolution or

desorption of more stable Pi (Tiessen and Moir 1993). Similar relationships were also

reported by McKenzie et al. (1992a, b) in a long-term crop rotation and fertilizer effects

study on P transformations in Chernozemic and Luvisolic soils.

3.4.2 Labile and moderately labile organic P (Bicarb-Po + NaOH-1-Po)

The Po extracted with NaHCO3 consists of loosely held low molecular weight organic

substances, such as ribonucleic acid, nucleotides and glycerophosphates (Bowman and

65

Cole, 1978). This fraction is considered as an active fraction of soil Po because this

easily mineralizable Po form (Oberson and Joner 2005; Chauhan et al. 1981) could

contribute to plant available P after mineralizing to Bicarb-Pi (Bowman and Cole 1978).

Therefore, Bicarb-Po fraction represents a labile pool of soil Po (Frossard et al. 2000),

and the dynamic nature of soil Bicarb-Po may play a significant role in crop production,

especially when the soil is low in plant available P.

In the corn phase 2006, the average concentration of Bicarb-Po fraction was16.39 mg P

kg-1, representing 6% of soil total-Po. However, the Bicarb-Po fractions in all the P

treated soils showed values comparable to the CK, indicating that there were no

significant impacts from all three forms of swine manure and inorganic fertilizer P

treatments in the year of application (Table 3.3 and Figure 3.2-c). These results are in

agreement with those of Campbell et al. (1986) that manure additions did not change

Bicarb-Po fraction in a Black Chernozem soil. Similarly, O’Halloran (1993) also reported

no significant impacts by manure or inorganic fertilizer additions on the Bicarb-Po levels.

Accordingly, Bicarb-Po fraction appears to be relatively insensitive to the effects of

inorganic fertilizer and manure application in this study (Sharpley, 1985; O’Halloran et

al. 1987b). The results showed that the Bicarb-Po fraction accounted for about 2% of

soil total-Pt.

Over the two-year study period, the Bicarb-Po fractions of TSP and MC treated soils as

well as CK significantly declined (Table 3.6). Hence, in the soybean phase of 2007, the

average concentration of Bicarb-Po fraction was 8.2 mg P kg-1 and represented only

about 2-3% of soil total-Pt. The decreases of Bicarb-Po could be explained as

stimulation of mineralization of soil organic matter (Po pools) from newly added crop

66

residues by greater microbial activity (Zhang et al. 2004). Similar results were observed

by Zhang and Mackenzie (1997b) and Zhang et al. (2004) who found a decrease in soil

Bicarb-Po with high levels of corn straw or swine manure application. In addition,

Tiessen et al. (1983) reported that Bicarb-Po was depleted rapidly in a Black

Chernozemic silt loam and a Dark Brown Chernozemic sandy loam soil during

cultivation. The Bicarb-Po in the SM and LM treated soils did not significantly change

from corn phase to soybean phase. As a result, Bicarb-Po fraction of the SM treated

soils (14.41 mg P kg-1) was significantly greater than that in the CK (2.81mg P kg-1). In

addition, SM treated soils showed a significantly higher Bicarb-Po value compared to

the LM (2.59 mg P kg-1). These different behaviours of Bicarb-Po with added P sources

were probably due to the variation of organic components in these different sources of

phosphorus.

Moderately labile Po (NaOH-1-Po) represents the Po held in more resistant or humified

forms, such as humic acids (Bowman and Cole 1978). This fraction was the most

abundant soil Po fraction (53% of total-Po), with the average concentration of 157.40 mg

P kg-1, representing 21-22 % of soil total-Pt in corn phase. These results are consistent

with Zhang and Mackenzie (1997) who reported that NaOH-Po was a sink for both

added Po and newly formed Po in a manure-inorganic fertilizer system in a Chicot

sandy clay loam soil. However, in both cropping seasons, moderately labile-Po fraction

in all the P-treated soils showed values comparable to the CK indicating that, NaOH-1-

Po fraction was not affected by manure or fertilizer treatments (Table 3.3). From corn

phase to soybean phase, NaOH-1-Po levels did not change significantly with manure P

treatments including CK, but changed with TSP treatment (Table 3.6 and Figure 3.3-e).

67

A similar observation was reported by Tran and N’Dayegamiye (1995) that manure

application maintained the NaOH-Po fraction; however, this fraction was decreased by

inorganic P fertilization.

The above findings agree with the some of the findings in the literature reported by

others from a wide range of soils with a wide range of organic and inorganic P sources.

Yet, there are others with contradictory findings; for example, Zamuner et al. (2012)

reported that Po fractions were not affected by fertilization. Gagnon and Simard (2003)

reported that Bicarb-Po and moderately labile-Po fractions were less affected by

manure treatments compared to Pi fractions. Yet, O’ Halloran (1993) reported that

manure applications had decreased the moderately labile Po fraction in the surface

layer of a sandy loam soil under no-till conditions. Similarly, Ciampitti et al. (2011)

reported that fertilization exerted a large influence on the moderately labile Po fraction.

However, Hountin et al. (1999) reported that liquid pig manure application significantly

increased biological Po (NaHCO3-P and NaOH-1-Po) content in the 0-20 cm soil layer.

These diverse findings about the relationship between fertilization and labile and

moderately labile Po may be due to the differences in soil types, nature and the rates of

the P sources applied in each study.

Overall, compared to NaOH-1-Pi levels, NaOH-1-Po levels were three times greater for

all the treatments. However, NaOH-1-Po level in CK was more than five times greater

than NaOH-1-Pi level, indicating that the soil itself contained substantial amounts of

moderately labile-Po. Conflicting results were observed by Hao et al. (2008) that the

concentration of Po (extracted by NaHCO3 and NaOH) was much smaller than Pi. They

further reported that Po accounted for less than 5% of total-Pt and this percentage was

68

not affected by the manure treatment. Other studies have also shown greater increases

in Pi than Po in soils that have received long-term applications of different types of

manure (Gale et al. 2000; Motavalli and Miles, 2002; Sharpley et al. 2004). These

results might be attributed to the fact that most of the P in manure and compost is

present as Pi (Sharpley and Moyer 2000).

3.4.3 Stable Phosphorus fraction (HCl-Pi + NaOH-2-P + Residual P)

The moderately stable P (HCl-Pi) fraction represents primary mineral P, such as apatite-

type minerals (Williams et al. 1980; Tiessen et al. 1984; Frossard et al. 1995) and other

forms of Ca- and Mg-bound P (Cross and Schlesinger 1995; Reddy et al. 1998; Simard

et al. 1995). This P fraction is generally assumed to be of low availability to plants

(Aulakh and Pasricha, 1991; Syers et al. 1972).

In both cropping seasons of 2006 and 2007, the results indicated that the HCl-Pi

fraction was not affected by fertilizer or manure treatments (Figure 3.4-f), given that the

P levels were very similar to those in the CK (Table 3.3). This indicates a minimal

contribution from swine manure and fertilizer treatments to this primary P mineral

fraction. In addition, over this two-year period, there were no measureable changes in

HCl-Pi fraction for all the treatments except LM (Table 3.6). Although decreases were

found with all the P treatments including CK, they were negligible and did not reach

significant levels.

The findings about the relationships between HCl-Pi and fertilization in previous

research are not unambiguous. For example, in one study the content of HCl-Pi was

shown to be not affected by repeated applications of barnyard manure (Campbell et al.

69

1986). Similarly, Wagar et al. (1986) reported that Chernozemic soils which received

160 kg P ha-1 showed no significant change in HCl-Pi level in the soil. In addition,

Lyamuremye et al. (1996) also reported that no significant change occurred in HCl-Pi

content in soils amended with manure. Similar results were reported by Sharpley et al.

(1991) in studying the impact of long-term swine manure application on soil and water

resources in Eastern Oklahoma, where only small amounts of HCl-Pi accumulated in

the soil profile.

Conversely, some past studies noted possible transformation of P into the HCl-Pi pool

with animal manure applications. Shafqat et al. (2009) reported that continuous

application of manure significantly increased HCl-Pi fraction. Similarly, increased HCl-Pi

levels especially in the upper soil layers with a history of pig slurry application were

noted by Gatiboni et al. (2008). These increases could be related with the fact that, in

general, more than 60% of P contained in swine slurry was found in the inorganic

fraction bonded to Ca (Barnett 1994; Sui et al. 1999). This is consistent with the findings

of the study conducted by Sharpley and Smith (1995), to determine the effect of animal

manure on soil P fractions. The most dramatic increase was observed in HCl

extractable P fraction, which might be the result of the addition of large amounts of Ca

with the manure. McKenzie et al. (1992b) and Song et al. (2007) also reported that HCl-

Pi increased due to fertilizer P during long-term crop production. However, it has been

found that, Ca-P originating from fertilizer P could be among the least stable P

compounds in specified soil types and can be readily re-mobilized back into the labile P

fractions (O’Halloran et al. 1987).

70

In both cropping phases, the amounts of HCl-Pi were greater than all other Pi fractions

for all treatments including the CK, indicating that soil itself contained considerable

amounts of HCl-Pi (Table 3.3). The average concentration of HCl-Pi was 192 and 171

mg P kg-1 in corn and soybean phase respectively. These high levels of HCl-Pi content

imply that a considerable amount of P in the soil is associated with Ca.

Considering total inorganic P in soil, the moderately stable P fraction contributes about

60% of soil total-Pi. Thus, moderately stable P fraction was the most abundant form of

Pi found in the soil and represented 26 - 29% of the total-Pt in both phases (Table 3.2).

Similarly, Daroub et al. (2000) reported that Ca-bound P accounted for 15 to 42% of

total-Pt in three long-term research sites. However, a much higher contribution (56%) of

HCl-Pi to total-Pt has been reported in a calcareous soil from Manitoba, Canada where

different rates of mineral P fertilizer were used (Yang et al. 2002).

After the HCl-Pi extraction, the remaining P fraction contains chemically stable organic

P forms and relatively insoluble inorganic P forms. This stable P fraction does not

contribute substantially to meet the plant P needs or to load P into surface water. Adding

another NaOH extraction to the end of the acid extraction to solubilize the occluded P

(Condron et al. 1990), the stable P fraction is further divided into NaOH-2- extractable P

(Pi and -Po) fraction (which is held within the internal surfaces of stable soil aggregates)

and the final residual-P fraction (Res-P), which contains highly recalcitrant Pi (Tiessen

and Moir 1993) and some stable forms of Po (Cross and Schlesinger 1995). These

stable forms of soil P are probably not directly available to plants, but may be involved

in long-term P transformations in soil (Tiessen et al. 1983).

71

In both cropping seasons, NaOH-2-Pi and NaOH-2-Po fractions did not show any

significant differences between the treatments (Figure 3.5-g and -h). All of the

treatments including the CK showed similar levels of NaOH-2-Pi and Po, indicating that

this occluded P fraction had no impacts from manure and inorganic fertilizer treatments.

However, this fraction represented considerable amounts, about 20-24 % of soil total-Pt

in both phases (Table 3.4). These results further indicated that NaOH-2-Po contents

were much greater (more than five times) than NaOH-2-Pi contents in both phases

(Table 3.3 and Figure 3.5-g and -h). Accordingly, the NaOH-2-Po fraction represented

about 42 % of soil total-Po fraction, while NaOH-2-Pi fraction represented about 7.2% of

soil total-Pi. The reason for the higher NaOH-2-Po level is probably attributable to the

fact that the soil itself may contain a significant amount of chemically stable Po

compounds. Over the two-year period, NaOH-2-Pi and Po levels in all the treated soils

including CK, did not change significantly except in LM (Table 3.6).

After the NaOH-2 extraction, the most recalcitrant fraction of P in the soil is Residual P

(Res-P) extracted by concentrated H2SO4 + H2O2 at 360˚C digestion (Tiessen and Moir

1993). This Res-P fraction does not contribute considerably to meet plant P needs or

loading to surface water in soluble forms, because it consists of more chemically stable

organic P forms such as humus and humic acids (Stewart et al. 1980) and relatively

insoluble inorganic P forms (Thomas et al. 1967). In this study, no significant differences

were found for Res-P fraction between the treatments and the CK indicating that

addition of manure and inorganic fertilizer had no immediate impact on Res-P (Figure

3.6-i). The average concentration of Res-P fraction was 101.96 mg P kg-1 representing

14 -15 % of total-Pt in the year of application.

72

Results also indicated that, within this two-year period, the Res-P fraction in all the P

treated soils including the CK did not change significantly except in the SM treated soils

(Table 3.6 and Figure 3.6-i). This reduction of Res-P with SM treatment may be due to

the transformation of Res-P into other P fractions, such as NaOH-Pi. Similarly, the Res-

P transformation to NaOH-Pi was reported by Zhang and Mackenzie (1997) in a study

investigating long-term soil P changes in a monoculture corn system using path

analysis. Since Res-P comprises a very stable P pool, significant changes in Res-P

concentration cannot be expected within this two year short-term period, because

conversion of manure P or fertilizer P into stable forms is a very slow process. However,

it was reported that Res-P increased with increasing rate of fertilizer application (Zhang

et al. 2004). In contrast, Zheng et al. (2003) found that the resistant pools in a Humic

Gleysolic clay soil were not influenced by fertilizer P application. However, Zhang et al.

(2006) reported that especially in exhaustive cropping systems, Res-P appeared to

continue to replenish the available P in soil. These contradictory results could be

attributed to differences in soil properties, experimental conditions, and different

management practices from location to location as well as continuously changing soil P

pools.

3.4.4 Total inorganic P (Total-Pi), Total organic P (Total-Po) and Total P (Total-Pt)

The total-Pi fraction was defined as the sum of all Pi fractions (labile-Pi + moderately

labile-Pi + moderately stable-Pi + stable-Pi) obtained from the sequential fractionation

procedure. In corn phase (2006), total-Pi levels were significantly increased in all the

manure and fertilizer treated soils compared to the CK. The average concentration of

total-Pi with P treatments was 341.65 mg P kg-1, i.e. 96.92 mg P kg-1 increase compared

73

to the CK (244.73 mg P kg-1). This shows application of P from different forms of swine

manure and inorganic P fertilizer resulted in accumulation of considerable amounts

(40% higher than CK) of P in total-Pi fraction. The highest increase was found with MC

(105.17 mg P kg-1) and the lowest increase was found with SM (86.13 mg P kg-1)

compared to the CK. In addition, all the swine manure treatments showed total-Pi

values similar to the TSP, indicating no significant differences between these manure

and inorganic P fertilizer treatments in the year of application (Table 3.3 and Figure 3.7-

k). The reason for this outcome may be that, in the corn phase, P was added at the

same rate from all of the P sources (i.e. 100 kg P ha-1), and also due to the quite similar

amount of crop P removal from these treatments.

Over the two-year period, total-Pi levels have significantly decreased from the corn

phase to the soybean phase for all the P treatments (Table 3.6). These decreases may

be due to the removal of P by soybean crops, transformation of Pi into unavailable P

forms, immobilization of P by microorganisms in order to maintain energy to mineralize

the organic residues added into the soil and also due to the losses from the soil through

surface runoff and leaching.

In the following year 2007, none of these treatments had a significant residual impact on

total-Pi fraction (Table 3.3). However, the average concentration of total-Pi in soils which

received P was 270.11 mg P kg-1. Accordingly, within the two-year period, the average

reduction of total-Pi in soils which received P treatments was 71.54 mg P kg-1. This

total-Pi concentration was 50.69 mg P kg-1 higher than that of in CK (219.42 mg P kg-1).

This indicates, although the residual effects were insignificant, total-Pi levels in the soils

which received P in the previous year was 23% higher than CK. These results reveal

74

that, regardless of the P source, added P from all forms of swine manure and inorganic

fertilizer made a considerable contribution to total-Pi fraction in the year of application

and the positive residual effects on soybean phase in the following year. In addition, this

increment may be partly accounted for by the transformations of soil Po into Pi forms

within this two-year period.

Overall, the total-Pi fraction represented 39- 47% and 39- 44% of the soil total-Pt in corn

phase and in soybean phase respectively (Table 3.4). These findings are somewhat

less than those from Sharpley and Moyer (2000) and Eghball (2003), who reported that

most of the P in swine manure was present as Pi. Dou et al. (2000) also reported that

most (up to 84%) of the P in manure was in available Pi forms; however, this Pi could be

susceptible to runoff loss after land application.

The total-Po fraction can be defined as the sum of all the Po fractions (Bicarb-Po +

NaOH-1-Po + NaOH-2-Po) obtained from the P fractionation. In this study, Bicarb-Po

accounted for 6% of total-Po, while moderately labile Po accounted for 52% of total-Po

and was the major Po fraction of this soil. The stable-Po (NaOH-2-Po) fraction

accounted for 42% soil total-Po. In the year of P application, the average concentration

of total-Po in soils treated with P was 295.77 mg P kg-1, showing 10.79 mg P kg-1 higher

level compared to the CK (284.98 mg P kg-1). However, none of these treatments had

significant impacts on total-Po fraction, indicating 3.8% increase was not enough to

show a significant impact (Table 3.3 and Figure 3.7-l). This indicates, within this short-

term period, total-Po fraction was not affected considerably due to swine manure or

inorganic fertilizer applications. Perhaps rates of manure applied may not have been

high enough for significant impact on soil-Po fraction. Furthermore, generally swine

75

manure contains more Pi than Po and inorganic fertilizer P does not contain Po at all;

therefore, less contribution to soil Po fraction would be expected.

Over the two-year period, the total-Po levels have significantly decreased in TSP and in

SM treated soils while the total-Po fraction in the LM and MC treated soils remained

unchanged as in the CK (Figure 3.7-l and Table 3.6). This significant reduction of total-

Po with TSP treatment over the two-year period was mainly accounted for by reduction

of Bicarb-Po as well as NaOH-Po fractions, which could be a source of labile P when

available P is drawn down due to plant P uptake. Similarly, Oniani et al. (1973) also

reported that super phosphate addition led to a decrease in Po fraction of the soil.

In terms of absolute amounts, total-Pi levels were generally greater than total-Po levels

in soils treated with manure (LM, MC and SM) and fertilizer in the year of application.

Several other studies have also shown more Pi forms than Po forms in soils that have

received different types of manure (Sharlpey et al. 1998, 2004; Gale et al. 2000;

Motavalli and Miles 2002). This is because manure contains higher amounts of P in

inorganic form and inorganic fertilizer P is entirely in Pi form. However, soils in the CK

plots showed relatively higher total-Po levels compared to the total-Pi levels indicating

that the soil itself contained considerable amounts of Po.

In the following year 2007, total-Po levels were greater than total-Pi levels for soils

treated with LM and SM in the previous year, including CK. This relatively lower total-Pi

concentration may be due to plant P uptake, immobilization of Pi into Po by microbes,

and also may be due to Pi runoff or leaching losses from surface soil. Generally,

inorganic fertilizers are in the soluble Pi form just after land application; therefore, higher

76

plant uptake can be expected. Eghball et al. (2002) reported that P in swine manure

was 91% plant-available within the year of application.

In this study, over this two-year period, both total-Pi and total-Po fractions decreased in

terms of absolute values for all P treated soils including CK (Table 3.3). This may

happen due to the internal transformations of P between P pools when labile P pools

were depleted due to plant uptake. When the Pi content of a soil decreases, the Po pool

can be mineralized into Pi pool, decreasing Po levels. Overall, in corn phase (2006), the

contribution from total-Po fraction to soil total-Pt was about 39 - 45% and in the

following year of soybean phase (2007), total-Po fraction accounted about 40 - 45% of

the soil total-Pt. This indicates, within this short-term period, total-Po fraction has not

changed considerably due to swine manure or inorganic fertilizer applications. The

reasons may be that the manure applied did not add enough Po to the soil Po pool to be

significantly detected and\or the depletion of Po was not adequate to result in a

measureable change.

The total-Pt fraction can be considered as the sum of total-Pi and total-Po fractions

together with most stable Res-P fraction obtained from the H2SO4 + H2O2 digestion of P.

In corn phase (2006), significantly higher (i.e.17.8% higher than CK) amounts of total-

Pt, were observed with all swine manure and fertilizer treatments compared to the CK

(Figure 3.7-j and Table 3.3). The average concentration of total-Pt in soils, which

received P treatments was 739.38 mg P kg-1, which was 111.75 mg P kg-1 higher

compared to the CK (627.63 mg P kg-1), indicating a considerable amount of P

accumulation in surface (0-15 cm) soil. This accumulation was mainly due to the P

addition with P sources; however, it may partially be contributed by the P addition from

77

crop residues. There were no significant differences between different forms of swine

manure treatments and TSP for total-Pt levels.

Over the two-year period, total-Pt levels in all P treated soils significantly decreased;

however, the decrease of total-Pt level in CK was insignificant (Table 3.6). These

significant decreases in plots with manure and fertilizer treatments could be due to

greater crop P uptake than CK and also due to soil P losses through surface runoff and

leaching. In the following year in 2007, the results indicate that the total-Pt fraction was

not affected by residual fertilizer or manure treatments, given that residual treatment

effects were not significantly different between all P treatments and the CK. However,

the average total-Pt level in all P treated soils was 632.06 mg P kg-1, showing 65.76 mg

P kg-1 higher total-Pt concentrations compared to the CK (566.31 mg P kg-1). Although

residual treatment effects were insignificant, this 11.6 % increase in total-Pt level in

soils, which received P in previous year, indicate that significant total-Pt changes require

long periods under cultivation. Therefore, a two-year period was not sufficient to detect

significant residual impact of previous year manure and fertilizer treatments on total-Pt

levels (Table 3.3).

3.5 Conclusions

This study revealed that additions of all three forms of swine manure (LM, SM and MC)

and inorganic fertilizer P (TSP) to clay loam soils influenced the labile P (H2O-P and

NaHCO3-Pi) and the moderately labile Pi (NaOH-Pi) fractions in the surface soils in the

year of application. The effects from all three forms of manure on a given P fraction

were comparable to the effects of TSP on the same P fraction, indicating no significant

78

differences between swine manure and fertilizer P treatments in the year of application.

The greatest forms of P found in this soil following swine manure and inorganic fertilizer

P addition were associated with the moderately labile P fraction. Total-Pi and Total-Pt

were significantly increased by all P treatments, while total-Po was not.

In the following year, only some of the treatments had residual effects on some of the P

fractions. However, none of the P treatments showed significant increases of Total-Pi,

Total-Po and Total-Pt as the residual effects of these treatments. In both phases of

cropping, none of the P treatments affected HCl-Pi or stable-P forms. Overall,

contributions of P from different forms of swine manure and inorganic fertilizer P are not

readily incorporated into Po and stable P fractions. In this study, increased soil test P

levels associated with all P sources indicate the potential for soil P loss through leaching

and runoff to the aquatic environment with all three forms of swine manure applications

and TSP.

79

Table 3.1 Physical and chemical compositions of the different forms of swine manure

materials applied to corn phase in 2006 on a Brookston clay loam at

Woodslee, Ontario, Canada.

Parameter

Liquid swine manure Solid swine manure Composted swine manure

g kg-1

Dry matter

60.4 ± 0.27 486 ± 8.50 279 ± 3.70

Organic C

28.6 ± 0.60 188 ± 6.99 88.0 ± 5.08 Total N

5.89 ± 0.08 23.7 ± 5.88 18.3 ± 1.12

NH4-N

2.50 ± 0.11 4.56 ± 0.15 3.63 ± 0.36

Total P

1.73 ± 0.01 3.79 ± 0.42 2.91 ± 0.42

Total K

1.77 ± 0.01 7.84 ± 0.75 4.18 ± 0.22

pH

7.7 8.2 6.6

Dry matter was measured on a wet weight basis and other parameters were measured

on a dry weight basis (Mean ± SE; n = 2 for dry matter and n = 3 for other parameters;

the concentration of NO3-N was negligible.

80

Table 3.2 Physical and chemical characteristics of the Brookston clay loam soil used in

field plots at Eugene Whalen Research Farm, Woodslee, Ontario, Canada.

parameter Units

Soil pH (soil: water 1:1 w/w) 6.1

Bulk density 1.51 ± 0.04 g cm-3

Sand 26%

Silt 34%

Clay 40%

CEC 18.8 ± 0.4 cmol kg-1

Organic C 22.7 ± 0.0 g kg-1

Total N 1.95 ± 0.0 g kg-1

Inorganic N (NH4-N + NO3-N) 14.1 ± 1.55 mg kg-1

Olsen test P 12.0 ± 1.0 mg kg-1

NH4OAc extractable K 131 ± 1.9 mg kg-1

NH4OAc extractable Ca 2.43 ± 0.1 g kg-1

NH4OAc extractable Mg 0.39 ± 0.0 g kg-1

81

Table 3.3 Distribution of Phosphorus fractions (mg P Kg-1) for different phosphorus sources in 2006 and 2007

2006 H2O-Pt

NaHCO3

-Pi NaHCO3

-Po

NaOH-1-Pi

NaOH-1-Po

HCl-Pi NaOH-

2-Pi NaOH-2_Po

Res-P Total-Pi Total-

Po Total-Pt

TSP 19.46 a 41.72

a 20.23

a 56.99

a 161.91

a 197.34

a 23.14

a 119.68

a 105.81

a 338.65

a 301.81

a 746.27

a

LM 26.29 a 37.95

a 10.88

a 51.68

a 160.92

a 209.76

a 21.51

a 127.57

a 98.19

a 347.18

a 299.38

a 744.75

a

MC 30.93 a 47.92

a 17.52

a 58.78

a 152.30

a 187.64

a 24.63

a 122.04

a 101.01

a 349.90

a 291.87

a 742.77

a

SM 21.37 a 43.42

a 16.93

a 56.97

a 154.47

a 185.01

a 24.09

a 118.63

a 102.83

a 330.86

a 290.03

a 723.72

a

CK 7.27 b 8.39

b 18.79

a 27.06

b 139.85

a 179.96

a 22.04

a 126.34

a 97.93

a 244.73

b 284.98

a 627.63

b

2007

TSP 9.54 abc

20.93 a 8.77

ab 41.95

a 147.15

a 184.46

a 21.46

a 115.64

a 96.61

a 278.34

a 271.56

a 646.51

a

LM 8.70 bc

20.48 a 2.59

b 37.06

ab 148.93

a 179.07

a 21.19

a 115.85

a 96.06

a 266.50

a 267.38

a 629.94

a

MC 13.19 a 25.32

a 6.93

ab 47.30

a 136.01

a 164.83

a 22.92

a 111.94

a 97.52

a 273.55

a 254.88

a 625.95

a

SM 10.82 ab

21.27 a

14.41 a

41.93 a 144.29

a 166.36

a 21.65

a 111.41

a 93.71

a 262.03

a 270.11

a 625.85

a

CK 6.06 c 6.75

b 2.81

b 24.75

b 131.27

a 161.43

a 20.44

a 117.18

a 95.62

a 219.42

a 251.26

a 566.31

a

Mean values within the same column followed by the same superscript are not statistically significantly different at P≤ 0.05

TSP: Inorganic Phosphate

LM: Liquid Swine Manure

MC: Manure Compost

SM: Straw Manure

CK : Control

82

Table 3.4 Phosphorus fractions for different phosphorus sources in both cropping phases (% of Total-P)

P source

H2O-Pt NaHCO3

-Pi NaHCO3

-Po NaOH-

1-Pi NaOH-1-Po

HCl-Pi NaOH-

2-Pi NaOH-2-Po

Res-P Total-

Pi Total-

Po

2006

TSP 2.61 5.59 2.71 7.64 21.69 26.44 3.10 16.04 14.18 45.38 40.44

LM 3.53 5.10 1.46 6.94 21.61 28.17 2.89 17.13 13.18 46.62 40.20

MC 4.16 6.45 2.36 7.91 20.50 25.26 3.32 16.43 13.60 47.11 39.29

SM 2.95 6.00 2.34 7.87 21.54 25.56 3.33 16.39 14.21 45.72 40.08

CK 1.16 1.34 2.99 4.31 22.28 28.67 3.51 20.13 15.60 38.99 45.41

2007

TSP 1.48 3.24 1.36 6.49 22.76 28.53 3.32 17.89 14.94 43.05 42.00

LM 1.38 3.25 0.41 5.88 23.64 28.43 3.36 18.39 15.25 42.31 42.44

MC 2.11 4.05 1.11 7.56 21.73 26.33 3.66 17.88 15.58 43.70 40.72

SM 1.73 3.40 2.30 6.70 23.06 26.58 3.46 17.80 14.97 41.87 43.16

CK 1.07 1.19 0.50 4.37 23.18 28.51 3.61 20.69 16.89 38.75 44.37

TSP: Inorganic Phosphate

LM: Liquid Swine Manure

MC: Manure Compost

SM: Straw Manure

CK: Control

83

Table 3.5 The significance levels of Analysis of Variance (ANOVA) for main effects

Source H2O-Pt NaHCO3-Pi

NaHCO3-Po

NaOH-1-Pi

NaOH-1-Po

HCl-Pi

NaOH-2-Pi

NaOH-2-Po

Res-P Total-Pi

Total-Po

Total-Pt

Model * * * * NS * NS NS NS * NS *

Treatment * * NS * NS NS NS NS NS * NS *

Time * * * * * * NS * * * * *

Treat*time * * * NS NS NS NS NS NS NS NS NS

R2 0.84 0.92 0.73 0.86 0.44 0.51 0.34 0.36 0.47 0.80 0.39 0.80

* Significant at P ≤ 0.05 level

Table 3.6 Treatment mean differences between 2006 and 2007, and their significant levels of the Hedley P

fractions

P source

H2O-Pt NaHCO3

-Pi NaHCO3-

Po NaOH-1-Pi

NaOH-1-Po

HCl-Pi NaOH-2-Pi

NaOH-2-Po

Res-P Total-

Pi Total-

Po Total-Pt

TSP 9.91* 20.79* 11.46* 15.05* 14.75* 12.88 1.68 4.05 9.20 60.31* 30.25* 99.76*

LM 17.59* 17.47* 8.29 14.62* 11.99 30.69* 0.31 11.72* 2.13 80.68* 32.00 114.81*

MC 17.75* 22.60* 10.59* 11.48 16.29 22.82 1.71 10.11 3.49 76.35* 37.00 116.84*

SM 10.55* 22.15* 2.52 15.04* 10.18 18.66 2.43 7.22 9.12* 68.83* 19.93* 97.88*

CK 1.21* 1.64* 15.98* 2.31 8.58 18.54 1.60 9.16 2.31 25.31 33.71 61.32

* Significant at P ≤ 0.05 level

84

Figure 3.2 The amounts of water extractable P fraction for soils treated with inorganic fertilizer P (TSP), Liquid

Swine Manure (LM), Swine Manure Compost (MC), Solid Swine Manure (SM) and the control (CK) for corn phase

in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil.

19.4

6

26.2

9

30.9

3

21.3

7

7.2

7 9.5

4

8.7

0

13.1

9

10.8

2

6.0

6

0

5

10

15

20

25

30

35

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Water extractable- Pt vs. P source 2006

2007

(a)

85

Figure 3.3 The amounts of NaHCO3 extractable -Pi and -Po fractions for soils treated with inorganic fertilizer P (TSP),

Liquid Swine Manure (LM), Swine Manure Compost (MC), Solid Swine Manure (SM) and the control (CK) for corn

phase in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil

41.7

2

37.9

5

47.9

1

43.4

2

8.3

9

20.9

3

20.4

8

25.3

2

21.2

7

6.7

5

0

10

20

30

40

50

60

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Bicarb-Pi vs. P source

2006

2007

(b)

20.2

3

10.8

8 1

7.5

2

16.9

3

18.7

9

8.7

7

2.5

9 6.9

3

14.4

1

2.8

1

0

10

20

30

40

50

60

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Bicarb-Po vs. P source

2006

2007

(c)

86

Figure 3.4 The amounts of NaOH-1 extractable -Pi and -Po fractions for soils treated with inorganic fertilizer P (TSP),


phase in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil.

56.9

9

51.6

8

58.7

8

56.9

7

27.0

6

41.9

5

37.0

6

47.3

0

41.9

3

24.7

5

0

20

40

60

80

100

120

140

160

180

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Moderately labile Pi vs. P source

2006

2007

(d)

161.9

0

160.9

2

152.3

0

154.4

7

139.8

5

147.1

5

148.9

3

136.0

1

144.2

9

131.2

7

0

20

40

60

80

100

120

140

160

180

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Moderately labile Po vs. P source

2006

2007

(e)

87

Figure 3.5 The amounts of HCl extractable -Pi fraction for soils treated with inorganic fertilizer P (TSP), Liquid

Swine Manure (LM), Swine Manure Compost (MC), Solid Swine Manure (SM) and the control (CK) for corn phase

in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil.

197.3

4

209.7

6

187.6

4

185.0

1

179.9

6

184.4

6

179.0

7

164.8

3

166.3

6

161.4

3

0

50

100

150

200

250

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Moderately stable P vs. P source

2006

2007

(f)

88

Figure 3.6 The amounts of NaOH-2-extractable -Pi and -Po fractions for soils treated with inorganic fertilizer P (TSP),


phase in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil.

23.1

4

21.5

1

24.6

3

24.0

9

22.0

4

21.4

6

21.1

9

22.9

2

21.6

5

20.4

4

0

20

40

60

80

100

120

140

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

NaOH-2-Pi vs. P source

2006

2007

(g)

119.6

8

127.5

7

122.0

4

118.6

3

126.3

4

115.6

3

115.8

5

111.9

4

111.4

1

117.1

8

0

20

40

60

80

100

120

140

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

NaOH-2-Po vs. P source

2006

2007

(h)

89

Figure 3.7 The amounts of Residual P fraction and Total-Pt fraction for soils treated with inorganic fertilizer P

(TSP), Liquid Swine Manure (LM), Swine Manure Compost (MC), Solid Swine Manure (SM) and the control (CK)

for corn phase in 2006 (blue) and soybean phase in 2007 (red) in Brookston clay loam soil.

105.8

1

98.1

9

101.0

1

102.8

3

97.9

3

96.6

1

96.0

6

97.5

2

93.7

1

95.6

2

80

85

90

95

100

105

110

115

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Soil Residual P vs. P source

2006

2007

(i)

746.2

7

744.7

5

742.7

7

723.7

2

627.6

3

646.5

1

629.9

4

625.9

5

625.8

5

566.3

1

0

100

200

300

400

500

600

700

800

900

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Total soil Pt vs. P source

2006

2007

(j)

90

Figure 3.8 The amounts of Total Pi and Total Po for soils treated with inorganic fertilizer P (TSP), Liquid Swine Manure

(LM), Swine Manure Compost (MC), Solid Swine Manure (SM) and the control (CK) for corn phase in 2006 (blue) and

soybean phase in 2007 (red) in Brookston clay loam soil.

338.6

5

347.1

8

349.9

0

330.8

6

244.7

3

278.3

4

266.5

0

273.5

5

262.0

3

219.4

2

0

50

100

150

200

250

300

350

400

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Total inorganic P vs. P source

2006

2007

(k)

301.8

1

299.3

8

291.8

7

290.0

3

284.9

8

271.5

6

267.3

8

254.8

8

270.1

1

251.2

6

0

50

100

150

200

250

300

350

400

TSP LM MC SM CK

So

il P

(m

g P

kg

-1)

P source

Total organic P vs. P source

2006

2007

(l)

91

Chapter 4: Phosphorous fractions in a grassland soil following long-term

application of dairy manure slurry and inorganic fertilizer

4.1 Abstract

Impairment of freshwater quality by accelerated eutrophication has focused attention on

manure management and the potential for P loss through runoff and leaching. The

objective of this study was to assess the long-term effects of dairy manure slurry (DMS)

and inorganic fertilizer (Ammonium nitrate; AN) on changes and distribution of soil P

forms in the soil profile. Plots of tall fescue (Festuca arundinacea) sward in the south

coastal region of BC, Canada, were treated with DMS or AN at 50 or 100 kg NH4-N ha-1

up to four times per year. Control plots received no manure or fertilizer. Soil samples

were collected at four depths (0-7.5, 7.5-15, 15-30 and 30-60 cm) and analyzed for P

fractions using a modified Hedley’s sequential technique. Application of DMS had

significantly different effects on soil inorganic P (Pi) fractions compared to the AN

fertilizer and such impacts were higher in the near-surface soil than in deep layers.

Phosphorus accumulation with either rate of DMS application was mainly in labile P

(H2O- Pi + Bicarb-Pi) and moderately stable P (HCl-Pi) fractions in top (0-15 cm) soil

layer, while no significant impacts on soil organic P (Po) and resistant fractions. The

main conclusion of this study was that P supplied with DMS can have both short-term

(labile P) and long-term (moderately stable P) impacts on soil P availability and risk of P

loss to the aquatic environment. On a short-term scale, the contribution from surface

soil layers was around 22% (2 and 20% from H2O-Pt and Bicarb-Pt from high rate of

manure treated soils, respectively) of total soil P from soil labile fraction representing a

considerable potential for surface water pollution.

92

4.2 Introduction

Dairy manure slurry (DMS) is a potentially important source of essential plant macro

nutrients, and can be used as a supplement for inorganic fertilizers in crop production.

There is interest in applying DMS on perennial forages, because these crops often take

up more nutrients than annual crops, and also allow several applications per year. This

is especially beneficial in the regions of increased livestock density. Since forages

provide year-round ground cover and typically have deep roots, manure may pose less

risk of run-off (Bittman et al. 2006). However, there are uncertainties over availability of

nutrients in DMS in the immediate and long-term, and the impacts of DMS on soil

inorganic P (Pi) and organic (Po) fractions compared to that of inorganic fertilizer. This

uncertainty also contributes to concerns regarding potential risk of soil P loading to

aquatic environments.

A wide range of effects of inorganic fertilizer and livestock manure applications on

different fractions of soil P have been reported in past studies. Such effects on soil P

fractions depend mainly on the rates of fertilizer or manure applications, P removal by

crops, inherent soil properties and climatic conditions. In general, continuous and

prolonged application of manure could lead to accumulation of both Pi and Po forms in

soil (Oniani et al. 1973; McKenzie et al. 1992a, b). Some studies have reported that

manure application has increased the concentrations of both total and soluble P, as well

as stable Po (Erich et al. 2002; Ylivainio et al. 2008; Waldrip-Dial et al. 2009). However,

Zhang and Mackenzie (1997) found that addition of manure has significantly increased

the labile (Bicarb-Pi and -Po) and moderately labile P (NaOH-Pi and -Po) fractions,

whereas stable P forms (HCl-P and residual-P) remained unaffected. Similarly, Meek et

93

al. (1982) and O’Halloran (1993) reported that Bicarb-Pi and NaOH-Pi fractions had

significantly increased after manure applications. Abbott and Tucker (1973) and Mnkeni

and Mackenzie (1985) also reported that repeated applications of manure over a long

period of time have increased the labile Pi fraction. However, Greb and Olsen (1967)

reported that 38 years of manure application has increased Po levels in three

calcareous soils. In contrast, Campbell et al. (1986) reported that, when manure was

applied once in three years, no significant changes were observed in the Po fraction of

the soil during the 36 years of the study period, although labile Pi levels increased.

According to these research findings, manure P availability and its impacts on soil P

fractions in the soil profile are not consistent, nor are they well understood. Thus, to

manage manure P for profitable and economically sustainable crop production and to

minimize environmental impacts, there is a need to further investigate its chemical

behaviour in soils by assessing soil P forms and their vertical distributions in the soil.

Thus, the objective of this study was to determine the long-term effects of DMS and AN

fertilizer on soil P forms and their changes and distributions in the soil profile.

4.3 Materials and Methods

4.3.1 Site description:

The study was conducted at the Pacific Agri-Food Research Centre at Agassiz in south-

coastal British Columbia, Canada (49° 10’ North, 125° 15’ West). The 30 year average

annual precipitation and mean temperature (from 1 February to 31 October) in Agassiz

are 1024.6 mm and 12.7 °C, respectively. The soil at the experimental site was derived

from medium-textured (silty loam to sandy loam) stone-free Fraser River deposits with

94

moderately good drainage. The soils (Eluviated Eutric Brunisols) were mapped as

Monroe series, which are equivalent to Typic Dyatroxerepts. Soil physical and chemical

properties are shown in Table 4.1. Soil pH was determined in a soil/water ratio of 1:1.

The hydrometer method was used for analyzing soil texture. Dry-combustion method

was used for analyzing organic matter (Leco CNS-1000 Analyzer, Leco Corp., St-

Joseph, MI). Phosphorus, K, Ca and Mg were extracted by Kelowna soil test extract

(Van Lierop 1988).

Table 4.1 Physical and chemical characteristics of the silty loam soil used in field plots

at Agassiz, British Columbia, Canada.

Parameter Unit (average)

Soil pH (soil: water 1:1 w/w) 5.0

Organic matter % 6

Sand (%) 28.9

Silt (%) 57.2

Clay (%) 13.9

Kelowna soil test extract:

P (mg kg-1) 66 (High)

K (mg kg-1) 82 (Medium)

Ca (mg kg-1) 723

Mg (mg kg-1) 31 (Low)

4.3.2 Treatments and soil sampling:

The experiment was initiated on a stand of tall fescue (Festuca arundinacea Schreb.

var. Festorina) established in 1993 and continued till 2002. The tall fescue stands were

restored through cultivation (conventional ploughing and disking) and reseeding in

95

2003. The fertilizer treatments were interrupted in 2003 and were reinstated in 2004 in

the same plots.

There were six treatments, including unfertilized control, two DMS treatments applied at

target rates of 50 kg N ha-1 (47,000 l ha-1) (M-Low) and 100 kg N ha-1 (90,000 l ha-1) (M-

High) of total ammonia nitrogen (TAN) per application; two ammonium nitrate (AN)

fertilizer treatments at 49 kg N ha-1 (F-Low) and 95 kg N ha-1 (F-High) per application;

and a treatment with alternating manure and fertilizer (Alt) each at 100 kg mineral-N ha-1

yr-1. The application rate of P for manure was averaged 15 and 29 kg ha-1 per

application. For slurry treatments total-N application was approximately 2xTAN. The

details of the treatments and fertilizer applications are given in Table 4.2. These

treatments were completely randomized with four replicates. The treatments were

applied to the plots (3 x 65 m) each year in the early spring and after each harvest,

except the final harvest.

Dairy manure slurry was obtained from manure storages on local high-input dairy farms

in which wood shavings were used for bedding. The slurry averaged 92% water. The

chemical composition of the dairy slurry is given in Table 4.3 (Bittman et al. 2004).

Manure was applied with a 3-m wide sleigh-foot or drag-shoe slurry applicator mounted

behind a 4000-L tank (Bittman et al. 1999).The sleigh-foot was designed to float freely

over the soil surface with little downward force so that there was no soil penetration.

Other nutrients (P, K and S) were applied only to fertilizer plots in spring at rates

according to local recommendations as indicated by the soil tests. In total, applications

in 2004 on both fertilized plots (F-low and F-high) were 0, 135, 11 and 22 kg ha-1 of P,

K2O, S and Mg, respectively.

96

Table 4.2 Fertilizer and dairy manure slurry application rates (Annual- 4 applications per

year)

Treatment

Manure Vol. (L/ha)

Manure TAM kg/ha

Fertilizer N kg/ha

Manure P kg/ha

Fertilizer P kg/ha

Fertilizer K2O

Fertilizer S kg/ha

Fertilizer Mg kg/ha

Control 0 0 0 0 0 0 0 0

F-Low 0 0 196 0 0 135 11 22

F-High 0 0 380 0 0 135 11 22

M-Low 180000 200 0 60 0 0 0 0

M-High 360000 400 0 116 0 0 0 0

Alt 180000 200 190 58 0 0 0 0

Table 4.3 Chemical and physical composition of the dairy manure slurry applied to a tall

fescue sward in a multi-year study in south coastal British Columbia

Parameter

Dairy manure slurry (average)

Dry matter %

8.0

Organic C %

4.1

Total N %

0.31

NH4-N %

0.16 Total P (g kg-1)

0.90

pH

6.7

In 2004, post-harvest soil sampling was done at four depths of 0-7.5 cm, 7.5-15 cm, 15-

30 cm and 30-60 cm from random locations on the plots. Thirty soil cores were taken

from each plot using a standard hand soil probe 2.5 cm in diameter and pooled to

produce composite samples from each depth. Samples were air-dried at room

temperature. After removal of visible crop residues, soils were crushed to pass through

97

a 2-mm sieve. Sub samples of soil were further ground to pass through No.140 mesh

for sequential P fractionation.

4.3.3. Soil P fractionation

Soil samples were sequentially extracted by the modified Hedley’s P fractionation

procedure to quantity the different soil Pi and Po fractions (Hedley et al. 1982). Detailed

methodology is given in Chapter 3 (Figure.3.1). Phosphate was determined

colorimetrically with the molybdate-ascorbic acid procedure (Murphy and Riley, 1962)

using a QuikChem Automated Analyzer (Lachat Instruments, Milwaukee, WI).

To calculate the recovery of P in the sequential extraction, total soil P was determined

using a separate soil sample digested with H2SO4-H2O2 (Thomas et al. 1967), followed

by colorimetric measurement using a QuickChem Automated analyzer (Lachat

Instruments, Milwaukee, WI). The results indicated that the modified Hedley’s

sequential fractionation procedure recovered a range of 95.3-102.2 % of the total soil

phosphorus of this study.

4.3.4. Statistical Analysis

To evaluate the effects of AN and DMS treatments on P fractions, results were analyzed

using Proc GLM in SAS 9.3.1. The contents of soil P fractions were first subjected to

ANOVA to identify the respective contribution of the treatment, soil depth and the

interaction term between soil depth and treatment to the total variance. Subsequently,

the treatment and the depth means were subjected to multiple mean comparisons to

identify statistically significant mean differences (SAS 9.3.1).

98

4.4 Results and Discussion

The significance levels of the ANOVA model for the main effects and the coefficient of

determination (R2) values of the model are given in Table 4.4. The depth effect was

found to be statistically significant (at p<0.05), in all P fractions, because generally, P

decreases in bioavailability with depth. However, the treatment effects of the ANOVA

model were found to be statistically significant (p< 0.05) only in some of the P fractions:

H2O- Pi, Bicarb- Pi, HCl-Pi, NaOH-2-Pi and residual- P. A significant treatment effect

was also found in Total-Pi fraction (Table 4.4). There was a significant interaction

(treatment * depth) observed only for H2O-Pi fraction. This may be due to the greater

decrease of P levels with depth in treatments where excess P was applied.

The distributions of different soil P fractions in the soil profile (0-60 cm depth) are

illustrated in Figures 4.1a to 4.1g. Along with the soil depths, all the P fractions, as

determined by the sequential fractionation, decreased from top to bottom layers for all

the treatments. In order to explain the distribution of P levels across soil depths,

average P values across different treatments were calculated at a given depth and are

given in Table 4.5. These results indicate that the average P levels found in the 0- 7.5

cm and 7.5 -15 cm soil depths were significantly greater than those found in the 15- 30

cm and 30- 60 cm soil depths, though the P levels in top two layers were not

significantly different from one another. The reason for this lack of difference of P levels

in top two layers was probably due to the in-field mixing of the soils, because the depth

of cultivation was greater than 7.5 cm. The P levels of all P fractions in the 15- 30 cm

soil depth were significantly greater than that in the 30-60 cm soil depth except HCl-Pi

and NaOH-2-Pi levels (Table 4.5). Relatively greater P level in the shallow (0-15 cm)

99

soil layer might have mostly been attributable to the accumulation of P in the surface

soil when P was applied with organic manure, and also due to greater biological activity

in this layer than deeper layers.

The results shown in Figures 4.1a- 4.1g generally indicate that the application of DMS

affected almost all of the P fractions mostly in the surface soil (0-15 cm) and slightly in

the sub-surface (15 - 60 cm) soils. Since the treatment impacts were mostly limited to

the top soil layer (0- 15 cm), and also due to the very few significant (treatment * depth)

interaction effects, the treatment mean comparisons were performed only with the mean

P values of the surface soil layer (0-15 cm) and results are discussed as follows.

100

Table 4.4: The significance levels of Analysis of Variance (ANOVA) for the main effects (treatments, depth

and the interaction and R2 values of the model (* significance at P ≤ 0.05 level)

H2O-Pi H2O-Po Bicarb-Pi Bicarb-Po NaOH-1- Pi NaOH-1-Po HCl-P NaOH-2-Pi NaOH-2-Po Res-P Total-Pi Total-Po Total-Pt

T reat * NS * NS NS NS * * NS * * NS NS

Depth * * * * * * * * * * * * *

Trt x Dep * NS NS NS NS NS NS NS NS NS NS NS NS

R2 0.78 0.44 0.93 0.88 0.93 0.84 0.67 0.55 0.9 0.9 0.93 0.91 0.94

Table 4.5: Statistical significance of Depth comparisons from ANOVA (P fractions (mg P kg-1 soil) averaged over

treatments)

Depth(cm) H2O-Pi H2O-Po Bicarb-Pi Bicarb-Po NaOH-1- Pi NaOH-1-Po HCl-P NaOH-2-Pi NaOH-2-Po Res-P Total-Pi Total-Po Total-Pt

0 - 7.5 12.8 a 7 . 1 a 200.3 a 133.2 a 576.2 a 399.5 a 173.2 a 27.6 a 135.6 a 97.5 a 990.0 a 675.4 a 1762.9 a

7.5 - 15 11.2 a 5 .4 a b 203.2 a 125.6 ab 577.0 a 397.6 a 177.2 a 27.5 a 143.1 a 98.0 a 996.2 a 671.6 a 1765.8 a

15 - 30 5 . 0 b 2 . 9 b 146.6 b 96.1 b 431.2 b 311.3 b 151.0 b 24.8 b 95.7 b 81.3 b 758.5 b 506.1 b 1345.9 b

30 - 60 0 . 8 c 0 . 9 c 33.7 c 28.6 c 180.2 c 111.4 c 143.7 b 23.6 b 23.4 c 56.2 c 382.0 c 164.2 c 602.4 c

Means in the same column followed by the same letter are not significantly different

101

Table 4.6: Mean values of Hedley P fractions (mg P kg-1 soil) of top 0-15 cm soil depth for all treatments

Treat H2O-Pi H2O-Po Bicarb-Pi Bicarb-Po NaOH-1- Pi N a O H - 1 - P o HCl-P NaOH-2-Pi N a O H - 2 -P o Res-P Total-Pi Total-Po Total-Pt

F-low 7 . 5 c 5 . 2 a 193.8 b 123.7 a 545.3 a 385.7 a 160.6 c 27.8 a 129.4 a 90.5 b 934.8 b 644.1 a 1669.4 a

F-High 7 . 7 c 5 . 0 a 186.1 b 129.1 a 542.7 a 401.6 a 161.3 c 27.5 a 131.7 a 98.9 ab 925.1 b 667.5 a 1691.5 a

M-Low 16.3 ab 7 . 4 a 215.8 ab 130.5 a 624.6 a 410.0 a 193.2 a 29.3 a 155.1 a 102.2 a 1079.2 a 702.8 a 1884.2 a

M-High 21.9 a 10.8a 230.6 a 143.3 a 617.0 a 388.1 a 190.4 a 28.2 a 148.7 a 102.2 a 1088.0 a 690.9 a 1881.1 a

A l t 11.9 bc 4 . 8 a 200.6 ab 134.0 a 579.7 a 425.7 a 177.4 ab 25.2 b 139.8 a 94.9 b 994.6 ab 704.3 a 1793.8 a

Control 7 . 0 c 4 . 3 a 183.6 b 115.6 a 550.7 a 380.0 a 168.2 bc 27.4 a 131.4 a 97.9 ab 936.9 b 631.2 a 1666.0 a

Means in the same column followed by the same letter are not significantly different

102

4.4.1 Labile Phosphorus fraction (H2O-P and Bicarb-P)

Labile P fraction refers to the water extractable P and Bicarb extractable P fractions and

corresponds to the P sorbed on the soil surface and to the easily mineralizable organic

P (Chauhan et al. 1981). In this study, the H2O-Pi fraction in surface soil (0-15 cm

depth) showed significant increases with either rate of DMS application compared to all

the other treatments (Table 4.5). It is obvious that average H2O-Pi fraction in 0-15 cm

soil depth was increased by 14.9 mg P kg-1 (21.9 – 7.0 mg P kg-1) in M-high treated

soils and by 9.3 mg P kg-1 (16.3 – 7.0 mg P kg-1) in M-low treated soils compared to the

H2O-Pi fractions of the control plots (7.0 mg P kg-1). In addition, these H2O-Pi levels

were two (in M-low treated soils) to three (in M-high treated soils) times greater than

H2O-Pi levels in AN fertilizer treated soils (Table 4.6). A similar observation was

reported by Simard et al. (2001) that the application of liquid hog manure to a

calcareous gleysolic soil in Quebec produced a very rapid increase in very labile P

(anion-exchange membrane extractable-P) fraction. In addition, H2O-Pi level in M-high

treated soil was significantly greater (by 10.0 mg P kg-1) than that of in Alt treatment.

However, either rate of AN fertilizer treatment or Alt treatment did not show significant

impact on H2O-Pi fraction compared to the control (Table 4.6).

The distribution pattern of H2O-Po fraction followed a similar distribution pattern as H2O-

Pi fraction, showing relatively high P levels in near-surface layers with either rate of

DMS application (Figure 4.1a-B). However, there were no significant impacts found from

either rate of DMS or AN or Alt treatments on H2O-Po fraction. In this study, H2O

extractable P fraction (Pi + Po) made a relatively small contribution (2%) to the total soil

P content.

103

Compared to the H2O-Pi fraction, Bicarb extractable-Pi fractions showed considerably

greater levels for all these treatments. As expected, Bicarb extractable-Pi fraction was

significantly increased in M-high treated soils (230.6 mg P kg-1) compared to either rate

of AN fertilizer treatments (193.8 and 186.1 mg P kg-1 for F-low and F-high treatments

respectively) and the control (183.6 mg P kg-1) plots (Table 4.5 and Figure 4.2b-C).

However, Bicarb-Pi fraction in F-low (193.8 mg P kg-1) and F-high (186.1 mg P kg-1)

treated soils were similar to the control plots (183.6 mg P kg-1) indicating no significant

impact from either rate of AN fertilizer treatments on Bicarb-Pi fraction.

It has been shown that increased labile Pi levels are commonly associated with manure

applications (Abbott and Tucker 1973; Campbell et al. 1986). These results are in

agreement with those findings that significantly higher labile-Pi levels in soils receiving

DMS applications may reflect the fact that greater amounts of Pi were applied in these

treatments. Similarly, Zheng et al. (2001) reported that liquid dairy manure applied to

gleysolic silty clay soil produced three times as much labile Pi increase per unit of P

added in surplus to plant exports than did the mineral fertilizer. They also reported that

this increase was closely related to the changes in soil C. Similar findings were reported

by O’Halloran (1993), in a study with liquid dairy manure applied on a continuous corn

system under reduced tillage, where he found that dairy liquid manure application had

significantly increased the levels of labile Pi fraction. Similarly Dormaar and Chang

(1995) also reported, after long-term cattle feedlot manure application, the amounts of

labile Pi in the Ap horizon of a Lethbridge loam soil (Dark Brown Chernozemic)

increased compared to the control. They also reported that the long-term applications of

feedlot manure had a very large impact on the labile P fractions, and the proportion of

104

the total soil P in Pi forms was greatly increased while labile Po was decreased.

However, these results were somewhat contradictory to a previous study done by Qian

and Schoenau (2000) where a single application of liquid hog manure had little impact

on the labile P fraction. The reason for these contradictory results is possibly due to the

differences in number of manure applications and also due to the total amount of P

added with the manure. Because, a single application of hog manure may not be

adequate to clearly identify where that P ends up, given the errors involved and the

amount of P in the soil. Also, the small impact by a single application may be due to the

applied manure P precipitating as calcium phosphate or converted to organic P in the

soil.

These positive effects of DMS may be related to the accumulation of soil P through

increased C addition and due to enhanced microbial and enzyme activities (Lalande et

al. 2000; Bissonnette et al. 2001), thereby, the rates of biologically-mediated turnover of

organic P could be increased (Sommers and Sutton, 1980; Mozaffari and Sims, 1994;

Tiessen et al. 1994). It can be assumed that the organic matter and organic acids that

are released from the DMS could effectively reduce P sorption/fixation to the soil. This is

possibly due to the competition between phosphate ions and the organic materials for

retention sites in the soil and such competition may lead to enhance the P availability

(Mnkeni and MacKenzie, 1985; Xie et al. 1991).

Similar to the H2O-Po fraction, there were no treatment effects found on Bicarb-Po

fraction from any of these treatments. Similar results were reported by Tran and

N’dayegamiye (1995), who found that long-term application of dairy cattle manure to the

same Le Bras silt loam soil (Humic Gleysol) maintained Po forms in 0-20 cm soil layer.

105

Campbell et al. (1986) also reported no changes in labile Po fractions in a Black

Chernozem with manure applications. Overall, Bicarb-P fraction made a considerable

contribution (9 – 21 % from bottom to top soil layers respectively) to the total soil P.

4.4.2 Moderately labile Phosphorus fractions (NaOH-1-P)

The moderately labile-Pi fraction represents the soil Pi that is associated with

amorphous and crystalline Al and Fe phosphates. In this study, the observed levels of

moderately labile Pi fraction were about three times greater than Bicarb-Pi fraction for

all the treatments (Table 4.5). However, Figure 4.2c-E shows that NaOH-1-Pi fraction

followed a similar distribution pattern as Bicarb -Pi fraction (Figure 4.2b-C). As was

observed for Bicarb-Pi levels, relatively higher NaOH-1-Pi levels at both rates of DMS

applications were observed compared to the AN treatments and the control. However,

these NaOH-1-Pi levels were not significant (Table 4.5).

Moderately labile Po fraction is assumed to be derived primarily from humic compounds

arising either due to addition of manures or decomposition of roots. This organically

bound P fraction could be a source of labile, and thus plant available P (Cross and

Schlesinger, 1995). Similar to the NaOH-1-Pi fraction, NaOH-1-Po fraction also did not

have significant impact due to application of either rate of DMS or AN fertilizers.

However, NaOH-1-Po levels were relatively lower compared to the NaOH-1-Pi levels for

all treatments (Table 4.5). O’Halloran (1993) reported that the amount of moderately

labile-Pi fraction increased with manure applications; however, moderately labile Po

fractions were not affected.

106

The distributions of labile and moderately labile P fractions (Figures 4.2b-C, 4.2b-D,

4.2c-E and 4.2c-F) in the soil profile have shown a very similar pattern. This is not

simply a result of the fact that more P in the surface layer means more P in each

fraction. This similar pattern indicates that these fractions were correlated with each

other. These relationships are in agreement with the previous studies that indicate,

available soil P fractions are constantly replenished through reactions of dissolution or

desorption of more stable inorganic P, and also through the mineralization of organic P

(Tiessen and Moir, 1993). Tiessen et al. (1984) reported that moderately labile P

fraction is sparingly available to plants by desorption. However, it acts as a labile P

fraction in absence of P inputs and may also act as the primary sink for external P

additions (Simard et al. 1995; Tiessen et al. 1984; Beck and Sanchez, 1994).

Although moderately labile Pi or Po fractions did not show any treatment impacts from

application of DMS or AN treatments, this P fraction was the largest and about three

times greater than the Bicarb-P fraction for all treatments. The results indicate that the

largest contribution (47- 58% from bottom to top layers respectively) to the total P was

from the moderately labile P fractions of this soil. However, these higher levels of

NaOH-1-P may not be due to the treatment effects, given that the soils in treated plots

had similar levels of NaOH-1-P in the control plots. Therefore, these higher levels of

NaOH-1-P are probably due to the soil itself containing higher levels of moderately

labile P.

Zheng et al. (2001) reported that NaOH-1-Pi fraction was the largest sink for excess Pi

in manure-treated soils. Similarly, Gagnon and Simard, (2003) reported that the largest

amount of P found in acidic soil following manure and compost additions was

107

associated with the moderately labile P fraction. These findings indicate that long-term

manure applications mostly affect the moderately labile P fraction in soils. In addition,

this high proportion of NaOH-1-P in the soil profile may also be due to P retention

associated with amorphous Al and Fe in soil. Furthermore, the increases in labile and

moderately labile Pi fractions could be attributed to re-sorption of P added in excess of

crop removal. Hedley et al. (1982) stated that the Pi not utilized by plants could be re-

adsorbed to soil components either as weakly or strongly adsorbed fractions (Bicarb-Pi

and NaOH-Pi).

4.4.3 Moderately stable Inorganic Phosphorus fraction (HCl-Pi):

Moderately stable Pi fraction is associated with Ca and Mg primary minerals in the soil.

This fraction is mainly apatite-type minerals and occluded P, and assumed to be of low

availability to plants (McKenzie et al. 1992a). In this study, significantly higher HCl-Pi

levels were found in soils treated with M-high (190.4 mg P kg-1) and M-low (193.2 mg P

kg-1) treatments compared to the F-low (160.6 mg P kg-1), F-high (161.3 mg P kg-1) and

the control (168.2 mg P kg-1) plots (Table 4.3). This confirmed that moderately stable P

(HCl-Pi) fraction showed a significant improvement with DMS application compared to

the AN treated plots and the control. These results were in contrast to the results

reported by Campbell et al. (1986) that the amount of HCl-Pi was shown to be not

affected by applications of barnyard manure. Similarly, in a study conducted to evaluate

the impact of long-term swine manure application on soil and water resources in

Eastern Oklahoma, Sharpley et al. (1991) found that only small amounts of HCl-Pi

accumulated in the soil profile.

108

Soils treated with Alt (177.4 mg P kg-1) have significantly high HCl-Pi fractions

compared to the soils that were treated with AN. This positive effect may be due to the

fact that when manure is applied alternatively with inorganic fertilizer, a considerable

amount of P could be added to the soil with manure application and may transfer to

moderately stable-P pool, resulting in greater concentrations of HCl-Pi. However, soils

treated with either rate of AN did not show such impact on HCl-Pi fraction, indicating

that HCl-Pi fraction was not sensitive to the AN applications alone. Similarly, Zhang et

al. (2004) observed that soil HCl-Pi remained constant after 5-10 years of inorganic

fertilizer application in a St-Rosalie heavy clay soil. However, McKenzie et al. (1992)

and Wagar et al. (1986) reported that HCl-Pi increased with the addition of inorganic

fertilizer. This disparity of results may be due to the differences in climatic conditions,

soil types and the nature of the P source applied.

In this study, the distribution of HCl-Pi fraction in the soil profile shows a somewhat

different pattern compared to all other P fractions (Figure 4.2d-G). In addition, HCl-Pi

contents were relatively low in all treatments compared to the Bicarb-Pi and NaOH-1-Pi

levels. The results also showed that regardless of the treatments, the percentage of

contribution to the Total-Pt from HCl-Pi fraction in the soil profile has increased from top

to bottom layers (17 - 40% from top to bottom layers, respectively). The reasons for a

greater contribution from deeper soil layers were probably due to either the presence of

more calcium associated phosphates or due to less weathered minerals.

109

4.4.4 Stable Phosphorus fraction

After HCl-Pi extraction, the remaining P fraction is considered as the stable P fraction.

The stable P fraction is the least plant available P fraction, because it consists of more

chemically stable organic P forms and relatively insoluble inorganic P forms. The stable

P fraction was further divided into NaOH-2-extractable P (NaOH-2-Pi and -Po) fraction,

that is the P from internal surfaces of soil aggregates, and the final residual- P fraction,

that is recalcitrant P.

Results of this study indicated that both NaOH-2-Pi and NaOH-2-Po fractions were not

significantly influenced by either rate of DMS or AN treatments (Table 4.5). However,

NaOH-2-Pi levels in Alt treatments (25.2 mg P kg-1) showed significantly lower P levels

compared to all other treatments and the control. For the final residual P fraction, P

levels in M-low (102.2 mg P kg-1) and M-high (102.2 mg P kg-1) treated soils were

significantly higher compared to the Alt (94.9 mg P kg-1) and F-low (90.5 mg P kg-1)

treatments. However, residual P fraction was not sensitive to the either rate of AN

treatments or Alt treatment (Table 4.5). According to these results, Stable (NaOH-2-Pi

and Po and Res-P) P forms did not show any significant changes with DMS or AN

application. A similar observation was reported by McKenzie et al. (1992a, 1992b). The

contribution to the total soil P from residual P fraction was from 5-10% from top to

bottom layers, respectively.

110

4.4.5 Total inorganic (Total-Pi), total organic (Total-Po) and total P (Total Pt)

In this study, the Total-Pi in 0-15 cm soil depth has significantly increased in M-high

(1088.0 mg P kg-1) and in M-low (1079.1 mg P kg-1) treated soils compared to those in

F-low (934.8 mg P kg-1), F-high (925.1 mg P kg-1) treatments and in the control (936.9

mg P kg-1) (Table 4.5). However, the Total-Pi in either rate of AN fertilizer treated soils

did not show any difference compared to the control plots, indicating that Total-Pi

fraction was not influenced by AN fertilization (Table 4.5). The Alt treatment did not

show a significant effect on Total-Pi fraction (994.6 mg P kg-1) either. Accordingly, only

DMS applications (both M-low and M-high) have influenced soil Total-Pi fraction; this is

because manure contained higher amounts of P in inorganic form. Overall, the Total-Pi

fraction contributes 54 – 66% of the Total-Pt in the soil profile and the contribution

increased with soil depth regardless of the treatment.

Overall, Total-Po fraction did not show any impact from application of either rate of DMS

or AN fertilizer treatments or Alt treatment. Compared to the Total-Pi fraction, a

comparatively small Total-Po fraction was observed for all the treatments. This is

because the accumulated contents of Pi in all three (H2O- Pi, Bicarb- Pi and NaOH- Pi)

P fractions were greater than the Po (H2O- Po, Bicarb- Po and NaOH- Po) fraction. The

same pattern was observed in all depths for all treatments and in the control. However,

for NaOH-2-extracted P fraction, Po fraction was 5- 6 times greater than Pi fraction for

all treatments. As a cumulative effect, Total- Pi fraction was greater than the Total-Po

fraction in the soil profile for all treatments and the control. In this study, Total-Po

fraction accounted for 24- 41% of the Total-Pt and the largest contribution was from the

111

upper layers and smallest contribution was from deeper layers regardless of the

treatment.

In terms of absolute amounts of soil Total-Pt, soils treated with M-high (1881.1 mg P kg-

1) and M-low (1884.2 mg P kg-1) have the greatest Total-Pt levels; however, these P

levels were not significantly greater compared to the control (1666.0 mg P kg-1) and

both rates of AN treated soils. In addition, the Total-Pt in Alt (1793.8 mg P kg-1), F-low

(1669.4 mg P kg-1) and F-high (1691.5 mg P kg-1) treated soils also were not

significantly different from Total-Pt levels of the control plots (Table 4.5). Overall,

application of P with DMS at either rates (M-low and M-high) or P application with Alt

treatment were not enough to give significant effects on Total-Pt fraction of the soil.

However, in soils treated with AN fertilizer, there has been an increased demand for soil

P due to increased root mass and faster crop growth due to N fertilization, that may lead

to lower Total-Pi and then Total-Pt levels compared to all other treatments.

The application of DMS has increased Total-Pt levels mostly in surface soils (0-15 cm)

and slightly in sub-surface (15-60 cm) soils. This may result from the accumulation of P

mostly in the surface soil layer where P was applied through slurry manure, and also

due to low vertical movement of P in soils, which may also lead to accumulation of P

mostly in the surface horizon (Sharpley et al. 1993; Mozaffari and Sims, 1994; Zheng et

al. 2001). The literature indicates that, in most soils, the P content of surface horizons is

greater than that of the subsoil. This is due to the sorption of added P, greater biological

activity, cycling of P from roots to above-ground plant biomass and more organic

material in surface layers. Some studies have found that addition of P did not influence

the concentration of any P fraction below the plow layer (Saleque et al. 2004 and Han et

112

al. 2005). However, in this study, the mean values of different P fractions in various soil

depths showed that although most of the Pi and Po were located in the 0-15 cm soil

depth, there were still certain amounts of P in the lower soil depths, especially in 15- 30

cm soil layer. This indicates, over time, P could be slowly leached into the deeper soil

horizons. Similarly, several studies have demonstrated appreciable downward

movement of P following the field application of manure, resulting in elevated P levels at

60-120 cm soil depth (Martin, 1970; Halstead and Mckercher, 1975).

4.5 Conclusions

Application of DMS has significantly different effects on soil Pi fractions compared to the

AN fertilizer and such impacts are greater in the surface soils than in deep layers. In this

study, P accumulation with either rate of DMS application was mainly in Total-Pi

fractions including labile P (H2O- Pi + Bicarb-Pi) fraction and moderately stable P (HCl-

Pi) fraction in top (0-15cm) soil layer. However, soil organic P fraction and resistant P

fraction were not influenced by either rate of DMS or AN application.

Application of AN fertilizer in either rate did not influence on soil P fractions. Thus, P

supplied with DMS can have both short-term (labile P) and long-term (moderately stable

P) impacts on soil P bio-availability and P loss. On a short-term scale, the contribution

from surface soil layers is around 22% of Total- Pt from soil labile P fraction (2 and 20 %

from H2O-Pt and Bicarb-Pt fractions from M-high treated soils, respectively)

representing a considerable potential for surface water pollution.

113

Figure 4.1a Water extractable Pi (A) and Po (B) fractions in the soil profile of 0-60 cm soil depth.

Treatments: F-low = 50 N kg ha-1, F-high =100 N kg ha-1, M-low = 50 kg ha-1TAN, M-high = 100 kg ha-1 TAN,

and Alt = alternating manure and fertilizer each at 100 kg mineral-N ha-1 yr-1 and the control.

0

10

20

30

40

50

-5 0 5 10 15 20

Dep

th (

cm

)

Soil P (mg kg-1)

Water Extractable Po

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

-5 0 5 10 15 20 25 30 35

Dep

th (

cm

) Soil P (mg kg-1)

Water Extractable Pi

F-Low

F-High

M-Low

M-High

Control

Alt

A B

114

Figure 4.1b Bicarb-Pi (C) and Bicarb-Po (D) fractions in the soil profile of 0-60 cm depth

Treatments; F-low = 50 N kg ha-1, F-high =100 N kg ha-1, M-low = 50 kg ha-1TAN, M-high = 100 kg ha-1 TAN,

and Alt = alternating manure and fertilizer each at 100 kg mineral-N ha-1 yr-1 and the control

0

10

20

30

40

50

0 50 100 150 200 250

Dep

th (

cm

)

Soil P (mg kg-1)

Bicarb-Pi

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

0 50 100 150 200

Dep

th (

cm

)

Soil P (mg kg-1)

Bicarb-Po

F-Low

F-High

M-Low

M-High

Control

Alt

C D

115

Figure 4.1c Moderately labile-Pi (E) and moderately labile-Po (F) fractions in the soil profile of 0-60 cm

depth. Treatments; F-low = 50 N kg ha-1, F-high =100 N kg ha-1, M-low = 50 kg ha-1TAN, M-high = 100 kg

ha-1 TAN, and Alt = alternating manure and fertilizer each at 100 kg mineral-N ha-1 yr-1 and the control

0

10

20

30

40

50

100 200 300 400 500 600 700

Dep

th (

cm

)

Soil P (mg kg-1)

NaOH-1-Pi

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

50 150 250 350 450

Dep

th (

cm

)

Soil P (mg kg-1)

NaOH-1-Po

F-Low

F-High

M-Low

M-High

Control

Alt

E F

116

Figure 4.1d HCl-Pi (G) and residual- P (H) fractions in the soil profile of 0-60 cm depth



0

10

20

30

40

50

100 125 150 175 200 225

De

pth

(c

m)

Soil P (mg kg-1) HCl-Pi

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

25 50 75 100 125

Dep

th (

cm

)

Soil P (mg kg-1)

Residual P

F-Low

F-High

M-Low

M-High

Control

Alt

G H

117

Figure 4.1e NaOH-2-Pi (I) and NaOH-2-Po (J) fractions in the soil profile of 0-60 cm depth



0

10

20

30

40

50

15 20 25 30 35D

ep

th (

cm

)

Soil P(mg kg-1)

NaOH-2-Pi

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

0 50 100 150 200

Dep

th (

cm

)

Soil P (mg kg-1)

NaOH-2-Po

F-Low

F-High

M-Low

M-High

Control

Alt

I J

118

Figure 4.1f Total-Pi (K) and total-Po (L) fractions in the soil profile of 0-60 cm depth



0

10

20

30

40

50

250 500 750 1000 1250

Dep

th (

cm

)

Soil P (mg kg-1)

Total Inorganic P

F-Low

F-High

M-Low

M-High

Control

Alt

0

10

20

30

40

50

0 200 400 600 800 1000

Dep

th (

cm

)

Soil P (mg kg-1)

Total Organic P

F-Low

F-High

M-Low

M-High

Control

Alt

K L

119

Figure 4.1g Total-Pt (M) in the soil profile of 0-60 cm depth.

Treatments; F-low = 50 N kg ha-1, F-high =100 N kg ha-1, M-low = 50 kg ha-1 TAN, M-high = 100 kg ha-1

TAN, and Alt = alternating manure and fertilizer each at 100 kg mineral-N ha-1 yr-1 and the control

0

10

20

30

40

50

500 1000 1500 2000 2500

Dep

th (

cm

)

Soil P (mg kg-1)

Total-Pt

F-Low

F-High

M-Low

M-High

Control

Alt

M

120

Chapter 5: Development of a soil phosphorus test for predicting long-term

phosphorus losses from agricultural soils

5.1 Abstract

Phosphorus losses from terrestrial systems to the aquatic environment contribute to

eutrophication of freshwater bodies. To reduce eutrophication and maintain freshwater

ecosystems, it is important to prevent P entering into freshwater bodies through

leaching and runoff from agricultural soils which are often enriched with P as a result of

over-fertilization and manuring. Current environmental soil P tests provide useful

information on the potential for immediate losses of P from agricultural soils. However,

such tests may not be sufficient for predicting long-term P losses in runoff/leaching

water. The objective of this study was to develop a suitable soil P test for predicting

long-term P loss potential from agricultural soils. Soil samples were collected from field

plots located in four different agro-ecological areas across Canada, including Harrow,

ON; Swift Current and Indian Head, SK; and Agassiz, BC. Soil P was analyzed using; 1)

original or modified soil P tests that are used for agronomic prediction of soil P supply

(Mehlich-3-P and Olsen-P) or are under consideration for environmental soil P testing

(resin membrane strips (RMS)-P, and FeO-strips-P) for indication of immediate soil P

losses and, 2) new procedures proposed for this study, including various combinations

of NaOH with different shaking time periods, with and without EDTA. The soil P

extracted with each individual extractant was correlated to the cumulative amount of soil

P that was removed by sequential extraction by RMS that was considered as the

amount of “Total Releasable P (TRP)” of that soil.

The amount of P extracted by different extractants varied widely. The Mehlich-3

extracted a greater amount of P than did Olsen, while RMS extracted more than twice

121

the amount of P than did FeO-strips. The mean extractable P values for Mehlich-3 were

similar to those of the RMS-P for soils from Harrow, Swift Current and Indian Head

sites, suggesting that Mehlich-3 test is as effective as RMS method for measuring the

TRP of these soils. All new tests extracted greater amounts of P than did existing

agronomic and environmental soil P tests. In addition, all the combinations of NaOH

with EDTA extracted 4-5 times greater amounts of P than did such combinations of

NaOH without EDTA.

The Olsen (r = 0.97), Mehlich-3 (r = 0.93) and FeO-strips (r = 0.97) methods were well

correlated with RMS-P. All NaOH with EDTA extractants were better correlated (r = 0.94

to 0.95) with RMS-P than all NaOH without EDTA extractants (r = 0.91 to 0.92). Overall,

highly significant linear (R2 = 0.89 to 0.91) and quadratic (R2 = 0.93 to 0.94)

relationships between RMS-P and NaOH with EDTA extractants suggests that these

extractants are as effective as sink-based RMS method for measuring TRP of that soil.

Among these newly proposed soil P extraction methods, the strongest linear (R2=0.91)

and quadratic (R2=0.94) relationships between RMS-P and 0.025M NaOH with EDTA

extractant indicates that this extractant might be the most suitable extractant for

predicting long-term P loss potential of agricultural soils.

5.2 Introduction

Phosphorus is often the most limiting nutrient of biological productivity both in terrestrial

and freshwater environments (Sharpley et al. 1994). Hence, P inputs to surface water

resources can increase biological productivity of these water bodies, leading to

accelerated eutrophication (Sharpley et al. 1999). Eutrophication has been identified as

122

one of the main causes of impaired surface water quality (USEPA, 1996), which

restricts water use for fisheries, recreation, industry and potable supplies. Thus, control

of P inputs to waters is of prime importance in reducing the accelerated eutrophication

of freshwater bodies.

Point sources, such as wastewater treatment plants, industrial facilities, sewage, and

drainage pipes are the main sources of P pollution. However, loss of P from agricultural

soils is now considered to be a significant source of pollution (USEPA, 2000), because

many point sources are largely under control due to easier identification (Daniel et al.

1994). Hence research attention has been focused on developing remedial strategies to

mitigate non-point (diffuse) source impacts of agricultural P.

Although P management is an integral part of profitable agronomic systems, previous

studies have consistently shown that long-term continued inputs of fertilizer and manure

P in excess of crop requirements have led to a build-up of soil P levels, which are of

environmental concern rather than agronomic concern, particularly in areas with

intensive agriculture associated with concentrated livestock production. Research has

also shown that runoff (surface and subsurface) and erosion from high P-level soils may

be the major factors contributing to surface water eutrophication (Sharpley et al. 1994;

Sims et al. 2002). Thus, the one of the main issues facing the establishment of

economically and environmentally sound P management systems are the identification

of those agricultural soils that contain high levels of P and have the greatest potential to

lose P to surface water bodies. This concern has led to an increased interest in

environmental soil P tests that may better assess a soil’s propensity to contribute to

nonpoint source P pollution than pre-existing agronomic soil P tests.

123

There is evidence that magnitude of soil P loss to water bodies is influenced by many

soil factors, management practices and environmental factors. These include; the soil

type (Miller 1979; Hooda et al. 1997), land use (Nearing et al. 1993), type, rate and

timing of fertilizer and manure application (Edwards and Daniel 1993; Hooda et al.

1999; Kleinman et al. 2001), amount, form and the distribution of soil P (Sharpley, 1995;

Hooda et al. 1999; Heckrath et al. 1995; Pote et al. 1996; Ige et al. 2005), soil physical

and chemical characteristics (Nearing et al. 1993), the amount and intensity of rainfall

(Edwards and Daniel, 1993), and the field slope and proximity to surface waters. For

example, surface water contamination has become of considerable concern due to high

soil P levels in many regions of Canada as a result of intensive crop and livestock

production, where there is also considerable runoff or soil erosion. Runoff, especially

associated with sloped land and rain events, transports an excessively large amount of

particulate P. Phosphorus loss through erosion is more severe in regions with intense

rainfall and where the soil on sloping land is not protected by a permanent cover of

vegetation. In flat fields, phosphate may be lost mainly by leaching from soils in which

the phosphate sorption capacity has been saturated by P fertilizer application. For

example, in the flat lands of Ontario, more than 50% of the total P losses might be lost

through subsurface drainage water (Culley et al. 1983). Moreover, on very flat prairie

landscapes in western Canada, movement of P into surface water occurs mostly as a

result of snowmelt rather than from rain water and the transported P is in the form of

dissolved reactive P rather than particulate P. Generally, snowmelt seems to carry more

highly reactive dissolved P which may be released from freezing and thawing of living

cells and tissues. Furthermore, there is evidence that P leaching can be a significant

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problem in poorly drained soils with high organic matter levels (Sharpley et al. 1994),

soils with a long history of manure application (Breeuwsma et al. 1995; Heckrath et al.

1995) and also in agro-ecosystems characterized by soils excessively rich in P with

small P sorption capacity (Gaynor and Findlay, 1995; Sims et al. 1998; Smith et al.

2001b). According, in any agro-climatic region, P losses are critically dependent on soil

factors, management practices and environmental factors that are highly variable both

in space and time. Among these factors, soil P level is the baseline risk indicator of

potential P loss to the aquatic environment through runoff and leaching.

There has been increasing interest in using agronomic soil P tests as indices of

potential soil P losses to water bodies (Mallarino et al. 2001). For example, Olsen-P,

Bray-1-P, Mehlich-3-P and Kelowna-P tests are often used in risk assessment for

estimating soil P loss potential in many regions of North America (Sims et al. 2000;

Sharpley et al. 1994). Because of the widespread use of these soil tests and the large

data base they provide on soil P, there is a considerable appeal to use such methods

for estimating the potential for P loss to surface waters. Another advantage of using

such soil P tests to assess risk of soil P loss is that, if one soil test can provide

information on both agronomic P recommendation and soil P loss potential, it would

practically save time, money and other resources required for assessments.

An issue with all these traditional soil P tests is that they are designed for soils with

particular characteristics (acidity and alkalinity) and they use dilute solutions of strong

acids, bases, and chelates to dissolve soil P depending on the soil properties. Thus,

their application over a range of soils with different properties result in the buffering of

acid or base extractants, leading to inefficiency with consequences of solubilizing non-

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labile P which are more tightly bound to Al, F, or Ca complexes (Myers et al. 2005).

When this occurs, such methods could overestimate labile soil P and then test results

for P loss potential are inaccurate. Furthermore, these soil test methods do not take into

account the slow release of sorbed P (Steffens, 1994) and soil organic P mineralization

(Tiessen et al. 1994). Therefore, these methods may not necessarily measure actual

amount of available P for loss and may not correlate well with soil P loss potential.

Given these limitations, use of existing soil P tests to manage P for water quality

protection, especially in the long run, although practical may not be worthwhile.

To overcome the problem of location and soil type dependency of the readings of soil

test P, a distilled water extraction method has been used extensively. This method is

designed to extract easily desorbable soil P (Sissingh 1971) and assumes that

extraction with water replicates the reaction between soil and runoff water. However, the

small amounts of soil P extracted by distilled water for most soils and difficulties related

to chemical analysis of these small amounts limit the use of distilled water as an

effective extractant.

As an alternative to these chemical and water extraction methods, ion-sink-based tests

have been proposed for environmental soil P testing. These ion-sink-based methods

rely on P sorption-desorption reactions instead of extracting soil P with strong

chemicals. Thus, the estimates of available P based on these desorption-based tests

could be better correlated with potential of P loss, because the extraction mechanisms

do not involve an arbitrary chemical extraction. The major advantage of these ion-sink

tests is the capability of extracting P from a variety of soil types irrespective of the

properties of the soil (Sharpley et al. 1994). Furthermore, previous studies have shown

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that ion-sink tests could be more suitable to simulate long-term desorption by using a

near-infinite sink for P, i.e. once P ions are adsorbed to the sink, they are not desorbed

(Sibbesen, 1978) and thus measure the potential of a soil to continue to release P

during a runoff or leaching event (Qian et al. 1992; Saggar et al. 1992).

Some of the most promising ion sink methods are Fe-oxide (FeO) impregnated filter

paper strips and anion exchange resin membranes strips (RMS). Research has shown

that the FeO-impregnated filter paper strips method (Chardon et al. 1996) effectively

estimated available P in a wide range of soils and management systems (Menon et al.

1989, 1990; Sharpley, 1991) and accurately predicted the quantity of P susceptible to

runoff, better than most agronomic soil P tests (Pote et al. 1996). Sharpley (1993) also

observed that the P content of runoff that was extracted by FeO-strips was closely

correlated (R2 = 0.92 - 0.95) with growth of several algal species. Accordingly, the P

adsorbed to FeO-strips would be a good indicator of the biological availability of P to

algae in runoff waters. However, there are some limitations to the use of FeO-coated

papers. FeO-coated papers are not available in the market, and this has led to different

methods for their preparation and use (Myers et al. 2005). Another concern is

contamination of the FeO-coated papers with fine soil particles during shaking (Chardon

et al. 1996), which can lead to error in estimating desorbable P. This can be minimized

by the use of CaCl2 solution as the background electrolyte which tends to minimize soil

dispersion (Myers et al. 2005). However, this can lead to a reduction in the amount of P

extracted (Koopmans et al. 2001). All these disadvantages of the FeO-coated papers

make the resin membrane strips (RMS) method is a more appealing sink-based method

for assessing available P. Furthermore, these resin strips can be re-used several times

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without losing their extracting power (Schoenau and Huang, 1991) and therefore they

are particularly cost effective. Research has also shown that the P extracted by RMS is

strongly correlated with bioavailable P (Abrams and Jarrell, 1992), the fraction of soil P

most likely to induce eutrophication (Sharpley et al. 1994).

Accordingly, the RMS method overcomes the disadvantages of chemical extractions,

water extractions, as well as FeO-strip method. It also has the advantages of cost

effectiveness, and applicability in soils of different regions with diverse chemical and

physical properties and also irrespective of management history. Moreover, the

advantage of RMS over chemical extractions is that it simulates long-term desorption by

providing a near-infinite sink for P. Therefore it can be used for predicting long-term P

loss potential. However, this method is tedious, time-consuming and not simple enough

for use by practitioners with varied technical backgrounds. Furthermore, it is difficult to

carry out on a large scale. These limitations make the procedure unsuitable for routine

use for soil P testing for environmental purposes. In addition, the lack of a consistent,

uniform and widely accepted standard procedure for this method limits its use as an

environmental test.

Currently, there are no methods available to assess long-term risk of P loss potential

from agricultural soils. As such, the objective of this study was to develop a suitable soil

P test for prediction of long-term P loss potential from agricultural soils.

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5.3 Materials and Methods

5.3.1 Site and experiment descriptions

For this study, soil samples were collected from four existing field experimental plots

located in different geographical locations across Canada (Figure 5.1) representing soils

of major agro-ecosystems in Canada. The locations selected were at the Eugene

Whelan Research Farm in Harrow, Ontario; the Semiarid Prairie Agricultural Research

Centre in Swift Current, Saskatchewan; the Pacific Agri-Food Research Centre in

Agassiz, British Columbia and the Agriculture and Agri-food Canada Research Farm in

Indian Head, Saskatchewan. These experimental sites provided a unique opportunity to

implement the present study by covering the major soil types with different physical and

chemical properties and various management histories in the typical agro-ecological

systems of Canada. The experimental field plots at each site provided a range of soil P

levels within each soil-agro-ecosystem as a result of the various treatments that have

been applied historically. Treatments used in these experimental sites are given in

Table 5.1. Soil sampling was done at post-harvest stages. Randomly selected soil cores

were taken from the 0-7.5 cm soil depth in each plot using a standard hand soil probe

with 2.5 cm internal diameter and pooled to produce a composite soil sample for

laboratory analysis. Analytical methods used for analysis of basic soil properties are

explained in next section and results are given in Table 5.2.

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5.3.2 Analysis of basic soil properties

Soil samples were air-dried at room temperature and crushed to pass through a 2 mm

sieve prior to chemical analysis. Soil pH was measured using 1:1 soil to water ratio

(Thomas, 1996). Soil organic carbon was determined by dry-combustion method with a

Leco CNS-1000 Analyzer, Leco Corp., St.Joseph, MI). Particle size distribution was

determined using a hydrometer method (Kroetsch and Wang, 2008).

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Figure 5.1 Major soil types in the typical agro-ecological systems of Canada

Harrow, ON

(Gleysolic)

Swift Current, SK

(Brown chernozemic) Indian Head, SK

(Black chernozemic)

Agassiz, BC (Brunisolic)

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Table 5.1 Different treatments used in field experiments located in four different agro-ecological areas across Canada

Agassiz (4 reps) Indian Head (3 reps) Swift Current (3 reps) Harrow (3 reps)

Dairy manure slurry - high Continuous Pea Alfalfa Solid swine manure

(100 kg TAN ha-1) (0 kg N ha-1) (0 kg P2O5 ha-1) (100 kg P ha-1)

Dairy manure slurry - low Continuous Pea Alfalfa Liquid swine manure

(50 kg TAN ha-1) (20 kg N ha-1) (20 kg P2O5 ha-1) (100 kg P ha-1)

Inorganic fertilizer - high Continuous Pea Alfalfa Swine manure compost

(95 kg N ha-1) (40 kg N ha-1) (40 kg P2O5 ha-1) (0

kg P2O5 ha-1)

(100 kg P ha-1)

Inorganic fertilizer - low Wheat-Pea rotation Alf-RWR Triple Super Phosphate

(49 kg N ha-1) (0 kg N ha-1) (0 kg P2O5 ha-1) (100 kg P ha-1)

Alternate manure & fertilizer Wheat-Pea rotation Alf-RWR Control

(100 kg N ha-1) (20 kg N ha-1) (20 kg P2O5 ha-1) (0

kg P2O5 ha-1)

(zero P)

Control Wheat-Pea rotation Alf-RWR

(zero N) (40 kg N ha-1) (40 kg P2O5 ha-1)

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Table 5.2 Physical and chemical characteristics of the soils (0-7.5 cm depth) of field experiments located in four different agro-

ecological areas across Canada

parameter Harrow Agassiz Indian Head Swift Current

Soil classification Humic Gleysol Eutric Brunisol Black Chernozemic Brown Chernozemic

Mean annual air temperature (˚C) 8.7 12.7 2.5 3.7

Mean annual precipitation (mm) 827 1025 434 330

Location (latitude and longitude) 42°13’N, 82°44’W 49°10’N, 125°15’W 50°33’N,103°39’W 50°15’N, 107°43’W

Soil pH (soil: water 1:1 w/w) 6.1 5.0 7.5 6.6

Sand % 26.0 28.9 24.1 35.96

Silt % 34.0 57.2 20.4 46.96

Clay % 40.0 13.9 55.5 17.08

Organic C (g kg-1) 22.7 34.8 23.2 16.70

P (mg kg-1) 12.0 (Olsen) 66.0 (Kelowna) 6.66 (Olsen) 5.73 (Olsen)

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5.3.3 Soil test phosphorus

Soil testing procedures for P were developed to estimate the amount of plant-available

P in the soil for agronomic purposes and fertilizer recommendations (i.e. to determine

how much P amendments are required for optimum crop growth) and not the amount of

bioavailable P that might be transported in runoff waters. However, it is logical to

consider that these both types of P measurements would be reasonably well correlated,

because, the forms of soil P that are required for terrestrial plants are those that are

soluble, readily desorbable, or "labile" and available to plants in the growing season.

Thus, if a soil is very high in soil test P, it is also likely to contain a high amount of

soluble or readily desorbed P which might be easily moved in runoff water.

The correlation between soil P in different extracting solutions and plant growth is

affected by soil and weather conditions hence regions use different extracting solutions.

The typical soil P tests that are used to measure soil P levels and recommend fertilizer

requirements for the various types of soils in different agro-ecological regions of Canada

are given in Table 5.3.

In this study, two agronomic P tests (Olsen and Mehlich-3), two existing environmental

P tests (Anion Resin Membrane strips and Iron oxide impregnated strips) and two newly

proposed procedures (NaOH with and without EDTA) were used to measure soil

extractable P of the surface soils (0-7.5 cm) collected from four experimental sites.

Olsen extractant (Olsen et al.1954) and Mehlich-3 extractant (Mehlich 1984) were

selected for this comparison, because they are the most widely used soil P tests to

extract available P from soils in Canada. Olsen extractant, the current agronomic soil P

test in Ontario, has been widely used for extracting P from wide range of soils including

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calcareous, alkaline and neutral soils and even in acidic soils (Kamprath and Watson

1980). As a soil test, Olsen-P is sensitive to management practices that influence

bioavailable soil P levels, such as fertilizer applications (O’Halloran et al. 1985) or

manure additions (Qian et al. 2004). Furthermore, the Olsen-P test has been used as a

alternate measure of soil P loss potential through runoff (Pote et al. 1996; Turner et al.

2004) and in regions using the Olsen-P as the recommended soil P test it is often a

criterion in soil P indices for assessing risk of P loss and impact on surface water

(Sharpley et al. 1994). Mehlich-3 extractant is commonly used as a multi-element

extractant (using ICP analysis), which is suitable for removing P and other elements in

acid and neutral soils. Thus, Mehlich-3 method is widely used in most provinces in

Canada especially in western Canada for agronomic soil P testing. Further, the Mehlich-

3 soil test extraction solution in association with the use of the measurement of

aluminum to estimate degree of P saturation has been proposed as a method to

determine environmental risk in North America (Pellerin et al. 2006).

Resin membrane strips and iron oxide strips methods were selected for this study,

because both of these ion sink methods are capable of extracting available P from soils

with large variation of physical and chemical properties. Further, these strips act as an

“infinite sink” to measure desorbable soil P, and thus measured the potential of a soil to

continue to release P during a runoff or leaching event.

Two newly proposed extraction methods include different concentrations of NaOH, a

strong chemical extractant that measures stable forms of P that are involved in the long-

term transformations of P in soils (Batsula and Krivonosova, 1973). From the literature,

it was established that the NaOH extractant measures moderately labile P sorbed on

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amorphous Fe and Al minerals and also the occluded P contained within aggregates.

Furthermore, addition of EDTA which is a strong chelating ligand, complexes Fe and Mn

cations in the extract, and thereby increases the extraction efficiency of soil P and also

the diversity of extracted P compounds (Bowman and Moir, 1993).

Generally, the efficiency of the extraction could be influenced by the strength of the

extractant and the length of the extracting period. Accordingly, for the newly proposed

extraction methods, soils were extracted with sixteen different combinations (4*4) made

out of four different concentrations of NaOH solution (0.05M, 0.1M, 0.15M and 0.2M)

with four different shaking durations (0.5h, 1h, 1.5h and 2h). For 0.1M EDTA + NaOH

extractant, soils were extracted with twelve different combinations (3*4) made out of

three different NaOH concentrations (0.025M, 0.05M and 0.1M) with four different

shaking durations (1h, 2h, 5h, and 16h).

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Table 5.3: Typical soil phosphorus tests used in Canada

Analysis method Extractant Comments

Olsen 0.5M NaHCO3 at pH 8.5 - best suited for neutral, alkaline and calcareous soils - current agronomic soil P test used in Ontario - critical level of soil test P is 20 ppm by Olsen-P in Ontario - process of maintaining pH level, driving off CO2, and filtering extractant through activated charcoal makes the procedure awkward (Qian et al. 1994)

Mehlich-3 0.2M CH3COOH

0.25M NH4NO3

0.015M NH4F

0.013M HNO3

0.001M EDTA

- common method for assessing crop- available P (using colourimetric method) - critical level of soil test P is 60 ppm by Mehlich-3-P (ICP) - multi-element extractant (using ICP analysis), which is suitable for removing P and other elements in acid and neutral soils

Bray-1 0.03N NH4F

0.025N HCl at pH 3.5

-designed for neutral - acidic soils (pH ≤ 7.0)

-not suited for alkaline soils (pH > 7.0)

Modified Kelowna 0.015M NH4F

0.25M ammonium acetate

0.25M acetic acid

- best method for a wide range of soil -pH levels - considered accurate for Canadian prairie soils (Havlin et al. 1999). - measures available P and K - performs similar to the Olsen test at high soil pH levels,

(however it does not require charcoal filtration and does not

evolve CO2 (Qian et al. 1994).

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5.3.3.1 Agronomic Phosphorus tests

Olsen extractable P was obtained by shaking 1.0 g of soil (2 mm sieved) with 20 ml of

0.5M NaHCO3 solution (pH of 8.5) in the presence of 0.25 g of charcoal, for 30 minutes

at 180 rpm on a reciprocating shaker and filtering the suspension through Whatman

No.40 filter paper (Olsen and Sommers 1982). Mehlich-3 extractable P was determined

by shaking 2.5 g of air dried soil sample with 25 ml of Mehlich-3 extracting reagent

(0.2M CH3COOH, 0.25M NH4NO3, 0.015M NH4F, 0.013M HNO3, 0.001M EDTA) for 5

minutes at 120 rpm and filtering the suspension through Whatman No.40 filter paper

(Mehlich,1984). Phosphate was determined colourimetrically with the molybdate-

ascorbic acid procedure (Murphy and Riley, 1962) using a QuikChem Auto Analyzer

(Lachat instrument, Milwaukee, WI).

5.3.3.2 Environmental Phosphorus tests

5.3.3.2.1 Anion resin membrane strips: (RMS)

The procedure followed for the RMS test was described by Tiessen and Moir (1993).

Sheets of a commercially available resin impregnated plastic material were cut into 2 x

10 cm strips. The strips were washed in distilled water to remove all propylene glycol

and stored in distilled water. The resin strips were saturated with HCO3

- by soaking

them in 0.5M NaHCO3 twice, for 3 hours each time and allowed to dry at room

conditions. Phosphorus was extracted from the soil by shaking 1 strip with 1g of soil and

30 ml of 0.01M CaCl2 solution in 40 ml bottles for 1hour shaking in an end-over-end

reciprocating shaker. Resin strips were replaced at 1st, 4th, 9th, 16th, 25th and 36th days of

time period, for a total period of 91 days with six sets of resin strips (i.e. 1st resin strip-

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after 24 hour contact period, shaken 1hour and removed from the solution; 2nd strip was

then placed and removed after 96 hours (4 days) contact period etc. (Sharpley et

al.1994). The strips were carefully removed after slow up-and-down movement of the

strip in the supernatant to remove any adhering soil particles back into the bottles,

thereby minimizing extraction of P from adhered soil as well as membrane P and any

loss of soil particles for subsequent extraction. Phosphorus retained on the resin

membrane strips was extracted by shaking the strip end-over-end with 30 ml of 1.0 M

HCl for 1hour. The membrane strip was removed, rinsed with deionized water, and

shaken end-over-end with an additional 30 ml of 1.0 M HCl for 1hour. Phosphorus in the

two HCl extracts was measured separately and summed to give resin P (freely

exchangeable inorganic P) assuming “Total releasable P” (Tiessen and Moir, 1993).

Phosphate was determined colourimetrically with the molybdate-ascorbic acid

procedure (Murphy and Riley, 1962) using a QuickChem Auto Analyzer (Lachat

instrument, Milwaukee, WI).

5.3.3.2.2 Iron oxide impregnated filter paper strips: (FeO-strips)

The procedure followed for the FeO-strips test was described by Chardon et al. (1996).

Iron-oxide impregnated filter papers were prepared by immersing paper discs (15 cm

diameter, Whatman No.50) in acidified FeCl3 using tweezers for 5 minutes (acidified

FeCl3; 0.65M FeCl3.6H2O + 0.6M HCl). The paper discs were removed from the solution

and allowed to drip dry at room temperature for 1hour. After air drying the papers were

immersed in 2.7M NH4OH for 30 seconds to neutralize the FeCl3 and produce

amorphous Fe oxide. They were then allowed to drain for 15 seconds and thoroughly

rinsed in two containers of clean distilled water to remove adhering FeO particles. After

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air drying, the filter papers were cut into strips (2 x 10 cm with reactive surface area of

40 cm2) and stored for subsequent use.

Phosphorus was extracted from soil by weighing 1 g soil into a 100 ml polyethylene

shaking bottle. One FeO strip was placed between Spectra /Mesh polythene screens

(2 x 11 cm), held together by a plastic clip on one end, making a paper-screen

assembly to insert into the shaking bottle that contained the soil and 40 ml of 0.01M

CaCl2 (soil: solution 1:40). The bottle was sealed and shaken horizontally on an end-

over-end reciprocating shaker for 16 hours reaction time (130 oscillations per minute).

The FeO strips were removed from the screens and rinsed under a stream of deionized

water for a few seconds to remove adhering soil particles. The strips were coiled and

placed in the neck of a 125 ml Erlenmeyer flask to air dry, pushed to the bottom and P

retained on the strips extracted by adding 40 ml of 0.1M H2SO4 to the flasks and

shaking for 1 hour (Chardon et al.1996). Phosphate was determined colourimetrically

with the molybdate-ascorbic acid procedure (Murphy and Riley, 1962) using a

QuikChem Auto Analyzer (Lachat instrument, Milwaukee, WI).

5.3.3.3 Newly proposed extraction methods:

For 0.1M EDTA + NaOH extraction method, 0.5 g of finely ground soil (140 mesh

sieved) was extracted with 15 ml of 0.1M EDTA and 15 ml of NaOH solution (three

different concentrations of NaOH: 0.025M, 0.05M and 0.1M) by shaking for four different

shaking times (1h, 2h, 5h and 16h). For NaOH extraction method, 0.5 g soil was

extracted with 30 ml of NaOH solution (four different concentrations of NaOH: 0.05M,

0.1M, 0.15M and 0.2M were used) by shaking for four different shaking times (0.5h, 1h,

1.5h and 2h).

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After shaking, samples were immediately centrifuged for 10 minutes at 25,000 g at 0°C

and vacuum filtered using 0.45 µm micro-filter papers. For both extractions, a portion

(10 ml) of the filtrates was pipetted into a 30 ml centrifuge tube, acidified (3.5 ml of 0.9M

H2SO4) and centrifuged at 25,000 g for 10 minutes at 0°C to remove precipitated

organic P, Po (Tiessen and Moir, 1993). Inorganic P (Pi) was then measured

colourimetrically using QuikChem Auto Analyzer (Lachat Instruments, Milwaukee, WI),

using the ammonium molybdate and ascorbic acid colourimetric procedure (Murphy and

Riley 1962).

A separate aliquot (10 ml) of the filtrates was acidified by ammonium persulfate

oxidation (autoclaved at 103.4 KPa and 121°C for 1h) and total P was determined using

QuikChem Automated Analyzer (Lachat Instruments, Milwaukee, WI), using the

ammonium molybdate and ascorbic acid colourimetric procedure (Murphy and Riley

1962). The difference between total P and Pi was considered as Po.

5.3.4 Statistical Analysis

The amounts of P extracted by the various extractants were subjected to multiple mean

comparisons for mean differences using STATA. The suitability of soil P tests as

indicators of the potential for long-term P losses were evaluated by comparing soil P

extracted by different extractants with the cumulative amount of soil P that was

determined using the sequential extraction of the resin membrane strips, assuming

within the 91 days extraction period, all desorbable P was adsorbed by resin strips

(Sharpley et al. 1994). This cumulative amount of P was assumed to represent the

“Total Releasable P” from that soil. Correlation and linear regression analyses were run

using STATA to study the relationships between P extracted by existing (agronomic,

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environmental) and new soil test P methods and the total releasable P. Stepwise

regressions (using STATA) were run to assess the significances of coefficient of

determination (R2) values of quadratic model compared to the linear model.

5.4. Results and Discussion

5.4.1 Comparison of soil Phosphorus extractability of existing agronomic and

environmental soil Phosphorus tests

Considerable variation was observed in the soil P amounts extracted by both agronomic

and environmental soil P tests (Tables 5.4a,and 5.4b). Across all soils tested, the

amounts of P extracted by agronomic soil P tests ranged from 4.0 to 204.8 mg P kg-1

and from 8.1 to 242.0 mg P kg-1 for Olsen and Mehlich-3 tests, respectively. Based on

the mean extractable P values of all soils tested (n=57), in terms of absolute amounts,

Olsen extracted less amount of P (i.e. mean = 46.3 mg P kg-1), compared to Mehlich-3

(i.e. mean = 63.9 mg P kg-1) (Figure 5.2).

When soils from individual locations were considered, for soils from Harrow, Swift

Current and Indian Head sites, Mehlich-3 extracted significantly greater amounts of P

compared to the Olsen. However, for soils in Agassiz site, approximately the same

amount of P was extracted in both the Olsen (mean = 144.2 mg P kg-1) and the Mehlich-

3 test (126.9 mg P kg-1) (Figure 5.2 and Tables 5.4a and 5.4b). The different behaviour

of Agassiz soil compared to the soils in Harrow and Saskatchewan sites is probably due

to the low soil pH in Agassiz site. These findings agree with the results reported by Sims

(2000a), that Olsen extractant has less ability to remove P from soils compared with the

Mehlich-3 extractant. Kleinman et al. (2001) found that Mehlich-3 values were roughly

1.5 times greater than corresponding Olsen P values for 24 soils from across the United

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States. Buondonno et al. (1992) found that Mehlich-3- P values were almost twice as

much as those of the Olsen-P. These findings indicate that Mehlich-3 solution extracted

some forms of labile P that are not immediately available to Olsen extractant. Generally,

acid based extractions such as used in the Mehlich-3 solution could extract

proportionally more soil P compared with other agronomic P tests. This is probably due

to the fact that as pH of the extracting solution decreases, more of the phosphate

associated with Ca is dissolved. Thus, Mehlich-3 being a more acidic reagent, extracted

a greater amount of P, whereas Olsen, being more alkaline, extracted a lower amount

of P.

When viewed across all sites, the amounts of P extracted by two environmental soil P

methods ranged from 9.3 to 151.8 mg P kg-1 and from 25.9 to 405.8 mg P kg-1 for FeO-

strips and RMS method, respectively (Table 5.4a). The mean extractable P values of

the RMS (91.3 mg P kg-1) were over twice as much as that of the FeO strips (36.3 mg P

kg-1). The same pattern was observed for each individual soil from the four experimental

sites (Figure 5.2). The greater amount of extractable P obtained by RMS method is not

surprising, because this RMS-P value is the cumulative amount of P extracted twice

from six resin strips within the period of 91 days compared to the one time extraction by

FeO-strips.

When results from agronomic P extraction methods were compared with the cumulative

amount of P extracted by RMS, the mean extractable P values for Mehlich-3 were

almost equal to those of the RMS-P for soils from Harrow, Swift Current and Indian

Head experimental sites (Figure 5.2). This suggested that Mehlich-3 method was as

effective as RMS method for measuring the total releasable P in these soils. However,

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for soils in Agassiz site, the amount of P extracted by RMS method was significantly

greater than the P extracted by Mehlich-3 test (Table 5.4a). In addition, when all soils

across the sites were considered, RMS extracted significantly greater amount of P

compared to the Mehlich-3 test. The amount of P extracted by Olsen test was

approximately half of the amount extracted by the RMS for soils from all four

experimental sites as well as for all soils across all sites (Figure 5.2).

The differences among these extracted P amounts by different extraction methods

probably arose from the fact that, extracting agents preferentially extract P from different

fractions depending on their reactions with soil constituents involved in P sorption

(CAST, 2000). Furthermore, each extracting solution has a different ability to extract

varying portions of soil P because they were targeted at different pools of soil P (Zhang

et at. 2004).

When comparing soils from different locations, the extractable soil P values by all these

extracting methods varied significantly. According to the mean extractable soil P values,

the soils from Agassiz site had the greatest amounts of extractable P with each

agronomic and environmental soil P extraction method and the soils from Swift Current

site had the smallest amounts (Figure 5.2 and Tables 5.4a and 5.4b). This greater

concentration of extractable P in Agassiz soil is probably due to its greater organic

matter content, associated with a history of receiving long-term dairy manure

application. It has been reported in the literature that a fairly close relationship exists

between organic C and available soil P (Bunemann et al. 2006), because the

accumulation of organic C may increase the availability of P in soil due to the

competition between organic anions and PO4-P for the same sorption sites (Muukkonen

144

et al. 2007). In Agassiz, due to increased livestock industry, excessive P applications

have occurred over the past two decades (Schreier et al. 2003) and 85% of all fields

were in the high (50-100 mg P kg-1 Kelowna Test-P) to very high (>100 mg P kg-1

Kelowna Test-P) environmental risk class for P in the 0-15 cm soil depth (Kowalenko et

al. 2007).

(Means with different letters are significantly different at P<0.05)

Figure 5.2 Means of soil P (mg P kg-1) extracted using existing agronomic (Olsen

and Mehlich-3) and environmental (resin membrane strips and iron oxide strips)

soil P tests for soils across four sites in Canada. (n= 57, 15, 12, 12 and 18 for all

soils across sites, Harrow, Agassiz, Swift current and Indian head, respectively).

c

b

c

c

b a a

a

b a

ab

a

b

a a

b b

b

c

b

0

50

100

150

200

250

300

So

il te

st P

mg P

kg -

1

RMS FeO Olsen Mehlich-3

145

Table 5.4a Descriptive statistics for P (mg P kg-1

) extracted by different extractants for all four experimental sites. RMS = Resin Membrane Strips, Total P (TP) = (Pi + Po), ENa = EDTA + NaOH, Na = NaOH, (0.025M, 0.05M and 0.1M) = concentrations of NaOH

Soil P extraction All soils (n=57) Harrow (n=15) Agassiz (n=12)

method Range Mean SD Range Mean SD Range Mean SD

RMS-TP 25.9-405.8 91.3 92.5 25.9-109.3 63.6 26.6 131.2-405.8 245.7 93.7

FeO-strips-P 9.3-151.8 36.3 29.1 9.3-63.1 33.5 14.9 51.7-151.8 81.1 31.8

Olsen-P 4.0-204.8 46.3 54.3 4.0-59.8 27.9 15.6 95.5-204.8 144.2 34.2

Mehlich-3-P 8.1-242.0 63.9 43.9 8.1-116.5 57.3 30.2 83.6-242.0 126.9 48.3

ENa_0.025M_1h_TP 166.8-1916.9 561.8 506.2 249.3-516.3 380.4 65.0 1052.2-1916.9 1490.0 281.2

ENa_0.025M_2h_TP 169.9-2012.9 581.5 528.1 238.9-516.7 381.7 68.9 1146.2-2012.9 1552.6 284.6

ENa_0.025M_5h_TP 174.9-2047.6 602.5 524.4 255.3-547.8 402.2 69.4 1182.3-2047.6 1566.5 275.7

ENa_0.025M_16h_TP 190.9-2027.5 610.5 526.6 250.2-567.1 410.6 75.1 1192.1-2027.5 1578.4 277.2

ENa_0.05M_1h_TP 169.2-2084.9 604.2 559.9 270.8-545.8 401.8 68.9 1218.4-2084.9 1638.5 278.2

ENa_0.05M_2h_TP 183.7-2085.4 632.9 574.7 276.2-556.8 423.5 69.9 1328.9-2085.4 1698.4 260.1

ENa_0.05M_5h_TP 199.2-2136.9 650.7 574.5 280.8-511.4 431.9 59.9 1325.4-2136.9 1713.5 273.8

ENa_0.05M_16h_TP 206.3-2153.8 663.8 577.3 289.2-513.5 419.2 71.5 1354.6-2153.8 1729.4 279.1

ENa_0.1M_1h_TP 146.7-2062.7 584.5 573.6 219.1-430.9 356.9 60.1 1189.2-2062.7 1644.5 291.7

ENa_0.1M_2h_TP 151.9-2098.4 602.6 583.5 233.5-455.7 371.1 64.5 1339.7-2098.4 1688.1 255.0

ENa_0.1M_5h_TP 174.4-2178.3 635.0 600.8 250.2-458.5 390.9 64.9 1389.4-2178.3 1751.6 261.1

ENa_0.1M_16h_TP 172.1-2187.3 637.1 586.3 291.6-479.2 405.9 51.7 1378.1-2187.3 1723.3 272.5

Na_0.05M_1h_TP 17.5-1627.6 306.0 515.9 41.3-101.5 72.3 17.7 781.7-1627.6 1266.7 279.3

Na_0.05M_2h_TP 17.5-1779.7 318.6 533.7 45.8-112.4 81.8 21.0 894.2-1779.7 1313.5 282.1

Na_0.05M_3h_TP 17.6-1749.2 337.1 555.7 49.3-125.7 87.3 22.9 1029.9-1749.2 1382.9 238.8

Na_0.05M_4h_TP 23.2-1672.3 339.9 556.9 51.7-125.6 89.7 22.4 1008.2-1672.3 1388.4 236.6

Na_0.1M_1h_TP 1.3-1788.4 343.1 571.0 45.7-121.6 87.3 23.4 1028.6-1788.4 1412.1 275.9

Na_0.1M_2h_TP 2.5-1843.8 355.1 576.9 43.1-129.3 95.9 27.2 1026.4-1843.8 1437.6 262.1

Na_0.1M_3h_TP 17.9-1952.4 383.7 629.5 57.4-149.2 107.1 27.6 1018.7-1952.4 1558.6 320.6

Na_0.1M_4h_TP 2.3-1855.5 369.8 593.2 58.5-137.2 107.6 25.7 1094.5-1855.5 1484.4 260.6

Na_0.15M_1h_TP 70.7-1854.7 364.2 540.9 93.3-147.8 123.7 17.1 809.8-1854.7 1362.1 333.8

Na_0.15M_2h_TP 69.6-1902.6 382.4 567.1 95.7-158.8 127.9 18.9 941.1-1902.6 1439.9 298.2

Na_0.15M_3h_TP 70.9-1949.9 396.8 588.6 96.1-156.2 130.1 17.8 1038.5-1949.9 1498.4 290.3

Na_0.15M_4h_TP 69.8-1706.8 355.7 498.3 99.4-164.7 134.4 19.7 1003.4-1706.8 1437.7 226.4

Na_0.2M_1h_TP 60.6-1870.4 373.9 565.1 81.1-154.2 115.7 19.8 897.9-1870.4 1424.8 312.2

Na_0.2M_2h_TP 57.4-1984.3 390.8 588.5 80.5-157.7 121.8 23.9 1050.3-1984.3 1487.2 314.2

Na_0.2M_3h_TP 54.6-2012.9 408.9 615.6 87.5-169.6 131.5 24.9 1118.9-2012.9 1560.5 304.4

Na_0.2M_4h_TP 59.8-2032.6 395.9 591.9 91.4-163.6 132.9 22.4 1130.3-2032.6 1494.5 334.5

146

Table 5.4b Descriptive statistics for P (mg P kg-1

) extracted by different extractants for all four experimental sites. RMS = Resin Membrane Strips, Total P (TP) = (Pi + Po), ENa = EDTA + NaOH, Na = NaOH, (0.025M, 0.05M and 0.1M) = concentrations of NaOH

Soil P extraction Swift Current (n=12) Indian Head (n=18)

method Range Mean SD Range Mean SD

RMS-TP 26.4-35.6 31.9 3.6 30.6-78.3 50.9 13.5

FeO-strips-P 12.9-20.7 17.1 2.2 13.2-37.4 21.6 6.6

Olsen-P 8.8-23.9 13.6 4.3 6.6-39.4 18.2 8.8

Mehlich-3-P 24.1-38.8 30.7 5.4 21.9-78.0 49.8 15.9

ENa_0.025M_1h_TP 166.8-203.8 185.3 11.1 250.9-457.4 345.3 50.8

ENa_0.025M_2h_TP 169.9-206.6 192.9 11.4 255.7-458.9 359.6 51.5

ENa_0.025M_5h_TP 174.9-220.4 205.1 12.9 290.8-493.1 391.6 52.7

ENa_0.025M_16h_TP 190.9-227.0 211.5 12.0 306.8-508.9 397.9 47.7

ENa_0.05M_1h_TP 169.2-200.2 186.6 9.0 274.6-474.2 361.7 46.6

ENa_0.05M_2h_TP 183.7-210.7 197.7 8.7 292.6-510.4 387.3 50.2

ENa_0.05M_5h_TP 199.2-232.8 213.6 9.9 316.8-527.6 415.9 53.8

ENa_0.05M_16h_TP 206.3-249.4 220.4 12.1 347.5-565.7 452.8 55.9

ENa_0.1M_1h_TP 146.7-173.9 162.0 9.9 264.7-437.3 349.1 45.9

ENa_0.1M_2h_TP 151.9-203.2 170.9 14.9 256.9-471.1 359.7 51.6

ENa_0.1M_5h_TP 174.4-207.3 188.9 10.9 265.3-576.9 391.5 72.9

ENa_0.1M_16h_TP 172.1-206.3 188.3 12.1 295.6-503.7 404.9 53.4

Na_0.05M_1h_TP 43.7-67.2 58.0 7.6 17.5-30.1 25.8 2.8

Na_0.05M_2h_TP 47.8-71.7 62.7 15.1 22.1-30.5 26.3 2.6

Na_0.05M_3h_TP 53.3-77.4 68.2 8.8 17.6-32.4 27.2 3.5

Na_0.05M_4h_TP 54.5-82.0 72.6 8.5 23.2-32.3 27.6 2.7

Na_0.1M_1h_TP 51.0-93.0 76.9 11.1 1.3-40.8 21.1 11.5

Na_0.1M_2h_TP 47.0-143.9 91.4 22.8 2.5-52.5 25.0 10.4

Na_0.1M_3h_TP 80.0-99.7 90.5 8.2 17.9-36.9 25.3 4.7

Na_0.1M_4h_TP 50.7-124.6 94.5 18.9 2.3-62.6 28.9 13.4

Na_0.15M_1h_TP 86.7-108.3 100.7 6.4 70.7-90.2 74.9 4.4

Na_0.15M_2h_TP 90.6-111.1 104.2 6.5 69.6-80.3 74.9 2.9

Na_0.15M_3h_TP 93.8-124.8 109.6 8.0 70.9-81.9 76.1 2.9

Na_0.15M_4h_TP 95.1-152.0 115.0 15.0 69.8-82.7 76.9 3.7

Na_0.2M_1h_TP 85.1-116.1 103.8 9.5 60.6-81.6 68.8 6.1

Na_0.2M_2h_TP 94.1-143.2 112.9 13.4 57.4-77.6 69.4 5.5

Na_0.2M_3h_TP 103.8-124.2 116.9 5.9 54.6-74.6 66.9 5.4

Na_0.2M_4h_TP 97.5-132.8 116.8 10.5 59.8-78.3 69.0 5.1

147

5.4.2 Comparison of Phosphorus extracted by newly proposed methods

For all soils across all sites, all three concentrations of NaOH with EDTA solution

extracted greater amounts of P (mean values ranged from 561.8 mg P kg-1 to 663.8 mg

P kg-1) than did all NaOH without EDTA extractants (mean values ranged from 306.0

mg P kg-1 to 408.9 mg P kg-1) (Table 5.4 and Figures 5.3- a and b). This suggests that

NaOH with EDTA extracted P from different pools or with different mechanisms than

NaOH alone extraction. This is most likely due to EDTA chelating metal cations and

thereby decreasing the amount of P bound via cationic bridges (Bowman and Moir,

1993).

The amounts of P extracted by NaOH with EDTA extractants increased with increased

shaking time regardless of the NaOH concentration (Figure 5.3-a). As expected, the

lowest extracted P values were observed with low concentration of NaOH (i.e.0.025M

NaOH) with EDTA extractant for all four shaking periods. Although higher extractions

were expected with stronger NaOH concentration (0.1M NaOH with EDTA), in this

study, the highest extracted P values were observed with 0.05M NaOH with EDTA

extractant regardless of the shaking period (Figure 5.3-a). Accordingly, the greatest

amount of P was extracted by 0.05M NaOH with EDTA for 16 hours of shaking period.

For NaOH without EDTA, the extractable P values increased with increasing shaking

period up to 1.5 hours (Figure 5.3-b), and with 2 hours shaking, the amount of P

extracted by all four NaOH concentrations decreased. This may be due to the re-

adsorption of P with longer periods of shaking. As expected, the amounts of P extracted

by all four concentrations of NaOH increased with increasing concentration of the NaOH

148

solution. Hence, the maximum amount of P was extracted by 0.2M NaOH with 1.5 hours

of shaking period (Figure 5.3-b).

For soils in Harrow, all three concentrations of NaOH with EDTA extracted significantly

greater amounts of P (mean values ranged from 356.9 mg P kg-1 to 431.9 mg P kg-1)

than did extractants without EDTA (mean values ranged from 72.3 mg P kg-1 to 134.4

mg P kg-1) (Table 5.4). Among all these NaOH with EDTA extractants, the maximum

extracted amounts were observed with the 0.05M NaOH with EDTA extractant while the

lowest was observed with 0.1M NaOH with EDTA extractant. The amounts of P

extracted by all three extractants increased with increased shaking time; however, for

0.05M NaOH with EDTA extractant, the extracted amounts tend to decrease after 5

hours of shaking period (Figure 5.4-a). Thus, the maximum extraction combination was

observed with 0.05M NaOH with EDTA for 5 hours of shaking period.

For NaOH without EDTA extractants, as expected, the extracted P amounts increased

with increased shaking period for all the NaOH concentration. Thus the maximum

amounts extracted by all four extractants were observed with 2 hours shaking period.

The highest amount of P extraction was observed with 0.15M NaOH extractant.

However, results indicate that both 0.2M NaOH and 0.15M NaOH extractants extracted

similar amounts of P for both 1.5 hours and 2 hours shaking periods (Figure 5.4-b).

For soils in Agassiz, the P extracted by NaOH with EDTA ranged from 1490.0 mg P kg-1

to 1751.6 mg P kg-1 and by NaOH without EDTA ranged from 1266.7 mg P kg-1 to

1560.5 mg P kg-1 (Table 5.4). The amounts of P extracted by NaOH with EDTA

increased with increased shaking period up to 5 hours for all three extractants, and then

149

tend to decrease (Figure 5.5-a). Results indicate that the maximum extraction

combination was with 0.1M NaOH + EDTA solution for 5 hours shaking period.

However, both 0.05M NaOH + EDTA and 0.1M NaOH + EDTA extracted similar

amounts of soil P at all four shaking periods. For NaOH without EDTA extractants, the

maximum extraction combinations were observed with 0.1M NaOH and 0.2M NaOH for

1.5 hours shaking period (Figure 5.5-b).

For soils in Swift Current, all three concentrations of NaOH with EDTA extracted

significantly greater amounts of P (mean values ranged from 162.0 mg P kg-1 to 220.4

mg P kg-1) than did NaOH without EDTA extractants (mean values ranged from 58.0 mg

P kg-1 to 116.9 mg P kg-1) (Table 5.4). The extracted amounts increased with increasing

shaking period, giving the maximum extractions with 16 hours for all four extractants.

However, greater amounts of P extraction were observed with 0.05M NaOH with EDTA

extractant, indicating the maximum extraction with 0.05M NaOH with EDTA for 16 hours

shaking period (Figure 5.6-a). For NaOH without EDTA, 0.1M NaOH and 0.2M NaOH

with 2 hours shaking period gave the maximum extractions (Figure 5.6-b).

For Indian Head site, the extracted P by all three concentrations of NaOH with EDTA

were significantly greater (mean values ranged from 345.3 mg P kg-1 to 452.8 mg P kg-

1) than the NaOH without EDTA (mean values ranged from 21.1 mg P kg-1 to 76.9 mg P

kg-1) (Table 5.4). The amounts of soil P extracted by all three NaOH with EDTA

extractions increased with increasing shaking time. The maximum extraction

combination was with 0.05M NaOH with EDTA for 16 hours shaking period (Figure 5.7-

a). However, for NaOH without EDTA, extracted amounts did not show such increase

with increasing shaking time. Both strong NaOH solutions (i.e. 0.15M NaOH and 0.2M

150

NaOH) extracted significantly greater amounts of soil P compared to low concentrations

of NaOH extractants (0.05M NaOH and 0.1M NaOH). The 0.15M NaOH solution

extracted the maximum P amounts regardless of the shaking period (Figure 5.7-b).

Among these new extraction methods, all NaOH with EDTA solutions extracted

significantly greater amounts of soil P compared to NaOH without EDTA extractants for

soils in Harrow, Swift Current and Indian Head Sites. The greatest amount of P

extraction was observed with 0.05M NaOH with EDTA extractant. For soils in Agassiz,

the results were slightly different compared to other three locations. Both strong NaOH

solutions (0.1M NaOH with EDTA and 0.05M NaOH with EDTA) gave similar extracted

amounts (Figure 5.5-a). This may be related to the higher organic matter content in this

grassland soil, because of the rapid breaking down of cationic bridges between organic

matter and PO4-P by EDTA with the presence of strong (0.1M NaOH) alkali solution.

This finding is supported by Bowman and Moir (1993) who concluded that NaOH-EDTA

extraction is more effective in soils high in organic matter where the chelation with metal

cations plays an important role in tying up organic P. Among NaOH without EDTA

extractants, the highest amounts of P were extracted by 0.15M NaOH solution for

Harrow and Indian Head soils and by 0.2M NaOH solution for Agassiz and Swift current

soils. As expected, greater P extractions occurred with1.5 to 2 hours shaking period for

all four sites.

Overall, both these reagents (NaOH with EDTA and NaOH without EDTA) extracted

significantly greater amounts of P than did agronomic (Olsen and Mehlich-3) and

environmental (RMS and FeO-strips) P tests (Table 5.4). These results agree with

findings reported by Mallarino, (1999) that NaOH extracted two to three times more P

151

from soils that received liquid swine manure than did Mehlich-3 or resin method. This

can be explained by the fact that NaOH solution changes the physical structure of

organic molecules in a way that enhances their solubility, given that, at high pH, many

organic functional groups are ionized and the increased charge density leads to

increased solubility. Furthermore, it is reported that the inclusion of EDTA, a strong

chelating ligand, complexes paramagnetic cations, such as Fe and Mn in the extract

and thereby increases soil P extraction efficiency and the diversity of P compounds

extracted (Bowman and Moir, 1993).

152

(a) (b)

Figure 5.3 Means of soil P extracted using various concentrations of NaOH with EDTA (a) and without EDTA (b) for

four different shaking periods (1h, 2h, 5h and 16h for NaOH with EDTA, and 0.5h, 1h, 1.5h and 2h for NaOH without

EDTA) for all soil samples collected from all four sites in Canada.

y = -2.9021x2 + 31.217x + 532.79 R² = 0.99, p<0.05

y = -3.9135x2 + 39.223x + 569.19 R² = 0.99, p<0.05

y = -4.0047x2 + 39.062x + 547.19 R² = 0.95, p<0.05

540

560

580

600

620

640

660

680

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)

Poly. (EDTA+0.025M NaOH)



y = -2.4371x2 + 24.195x + 283.19 R² = 0.97, p<0.05

y = -6.4494x2 + 43.127x + 303.48 R² = 0.81, p<0.05

y = -14.817x2 + 73x + 303.39 R² = 0.87, p<0.05

y = -7.435x2 + 45.583x + 334.22 R² = 0.92, p<0.05

250

270

290

310

330

350

370

390

410

430

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)

Poly. (0.05 M NaOH) Poly. (0.1 M NaOH)


153

(a) (b)



EDTA) for soil samples collected from Harrow field experimental plots.

y = 1.7581x2 + 2.3188x + 374.72 R² = 0.93, p<0.05

y = -8.6163x2 + 49.175x + 360.79 R² = 0.99, p<0.05

y = 0.201x2 + 15.686x + 340.48 R² = 0.99, p<0.05

350

375

400

425

450

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h) Poly. (EDTA+0.025M NaOH)



y = -1.7882x2 + 14.712x + 59.407 R² = 0.99, p<0.05

y = -2.0414x2 + 17.41x + 71.268 R² = 0.97, p<0.05

y = 0.0478x2 + 3.1911x + 120.68 R² = 0.99, p<0.05

y = -1.1726x2 + 11.992x + 104.28 R² = 0.96, p<0.05

60

80

100

120

140

0 1 2 3 4 5E

xtr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)



154

(a) (b)



EDTA) for soil samples collected from Agassiz field experimental plots.

y = -12.683x2 + 91.322x + 1413.7 R² = 0.98, p<0.05

y = -11.009x2 + 83.807x + 1568 R² = 0.98, p<0.05

y = -17.967x2 + 119.81x + 1537.1 R² = 0.90, p<0.05

1450

1500

1550

1600

1650

1700

1750

1800

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)




y = -10.33x2 + 95.129x + 1177.5 R² = 0.96, p<0.05

y = -24.936x2 + 158.48x + 1264 R² = 0.66, p<0.05

y = -34.626x2 + 201.64x + 1190.1 R² = 0.95, p<0.05

y = -32.102x2 + 188.75x + 1260.6 R² = 0.88, p<0.05

1250

1300

1350

1400

1450

1500

1550

1600

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)



155

(a) (b)



EDTA) for soil samples collected from Swift Current field experimental plots.

y = -0.3189x2 + 10.684x + 174.37 R² = 0.9867

y = -1.0546x2 + 17.007x + 169.96 R² = 0.9861

y = -2.3604x2 + 21.498x + 141.48 R² = 0.9277

150

170

190

210

230

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)




y = -0.0773x2 + 5.3035x + 52.693 R² = 0.9986

y = -2.6133x2 + 18.251x + 62.305 R² = 0.8885

y = 0.473x2 + 2.4754x + 97.632 R² = 0.9986

y = -2.3132x2 + 15.87x + 90.31 R² = 0.9994

40

60

80

100

120

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)



156

(a) (b)



EDTA) for soil samples collected from Indian Head field experimental plots.

y = 18.98x + 326.14 R² = 0.94, p<0.05

y = 30.166x + 329 R² = 0.99, p<0.05

y = 19.931x + 326.47 R² = 0.96, p<0.05

300

350

400

450

500

0 1 2 3 4 5

Extr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)

Linear (EDTA+0.025M NaOH)



y = 0.5079x + 25.562 R² = 0.90, p<0.05

y = 2.3516x + 19.186 R² = 0.92, p<0.05

y = 0.7563x + 73.826 R² = 0.90, p<0.05

y = -0.1752x + 68.944 R² = 0.04, p<0.05

0

20

40

60

80

0 1 2 3 4 5E

xtr

acte

d s

oil P

(m

g P

kg

-1)

Shaking period (h)

Linear (0.05 M NaOH) Linear (0.1 M NaOH)

Linear (0.15 M NaOH) Linear (0.2 M NaOH)

157

5.4.3 Correlations between amounts of P extracted by existing (agronomic and

environmental) soil P extraction methods and the cumulative amounts of P

extracted by resin membrane strips; “Total Releasable P”

For all soils sampled, the amount of P extracted by the FeO-strips method was

significantly correlated (r = 0.97, p < 0.05) with the amounts of P extracted by the RMS

method (Table 5.5). The high correlation between these two environmental tests was

expected, because they both have sink-based extracting mechanisms. Although not

expected because of the markedly different extraction mechanisms, good correlations

(p < 0.05) were also found between the amounts of P extracted by agronomic extraction

methods and RMS-P, with correlation coefficients of 0.97 and 0.93 for Olsen and the

Mehlich-3 methods, respectively (Table 5.5). These high correlations indicate that, even

though each of these extracting agents extracted a different proportion of available P,

they all are capable of estimating TRP of that soil. Similarly a significant correlation

between P desorbed by anion exchange membranes and 0.5M NaHCO3 extractable P

(Olsen method) was reported by Qian et al. (1992). Nuernberg et al. (1998) found a

significant correlation between P desorbed by anion exchange membranes and P

extracted by two acidic extractants (Mehlich-1-P and Bray-1-P).

When individual sites were considered, for soils from Harrow, the amounts of P

extracted by Olsen method showed the most significant correlation (r = 0.94) while the

Mehlich-3 method showed the smallest (but still significant) correlation (r = 0.89) with

RMS-P (Table 5.5). For soils from Agassiz, the P identified by all three methods showed

strong significant correlations with RMS-P, having correlation coefficients of 0.95, 0.95

and 0.93 (p < 0.05) for Olsen-P, FeO-strips-P and Mehlich-3-P respectively. Stronger

correlation is likely due to the greater P content of the soil from this site compared to all

158

the other sites, as indicated by the amount of P extracted using existing soil P testing

methods. For soils from Swift Current and Indian Head plots, none of these existing

methods showed significant correlation with RMS-P.

5.4.4 Correlations between P extracted by newly proposed methods and the

cumulative amounts of P extracted by resin membrane strips; Total Releasable P

For all soils across sites, the amounts of P extracted by all new methods were

significantly correlated (p < 0.05) with RMS-P. The correlation coefficients ranged from

0.94 to 0.95 for NaOH with EDTA extractants and from 0.91- 0.93 for NaOH without

EDTA extractants (Table 5.5). These results indicate that all the combinations of NaOH

with EDTA extractants showed stronger correlations with total releasable P, than did all

combinations of NaOH without EDTA extractants.

When soils from individual sites were considered, significant correlations between

amounts of P extracted by NaOH with EDTA and the RMS-P were observed for soils

from Indian Head and Agassiz (Table 5.5). For Agassiz soils, all three concentrations

(0.025M, 0.05M and 0.1M) of NaOH with EDTA for 5 hours and 16 hours shaking period

showed significant correlations with RMS-P. For soils from Indian Head plots significant

correlations between P extracted by all three concentrations of NaOH with EDTA

extractants and RMS-P were observed with all four shaking periods except a few

combinations. However, none of these NaOH with EDTA extractants shows significant

correlations with RMS-P for the soils from Harrow and Swift Current plots. Further, there

were no significant correlations between RMS-P and the P extracted by different

combinations of NaOH without EDTA extractants for soils from all four locations. The

159

reason may be the small sample size when they were considered individually within the

sites and also due to the range of values observed at any one site.

Overall, these results indicate that P extracted by different concentrations of NaOH with

EDTA was well correlated with total releasable P for all soils across the sites, as well as

for soils from Agassiz and Indian Head sites (Table 5.5). However, P extracted by

different concentrations of NaOH without EDTA extractants were well correlated with

total releasable P, only when soils across all sites were considered. This likely reflects

the bigger sample size when the entire set of soils was considered or a greater range of

soil P levels, and thus more variability.

160

Table 5.5 Pearson correlation coefficients (r) between amounts of P extracted by P

extraction methods and Total Releasable P (RMS-P) for the whole soil collection (n=57),

and for individual locations (The correlations with * are significant at p < 0.05)

P extraction method Whole soil collection Harrow Agassiz

Swift current

Indian Head

n=57 n=15 n=12 n=12 n=18

FeO-strips-P 0.97* 0.92* 0.95* 0.83 0.43

Olsen-P 0.97* 0.94* 0.95* 0.70 0.41

Mehlich-3-P 0.93* 0.89* 0.93* 0.89 0.29

ENaOH_0.025M_1h_TP 0.95* 0.54 0.85 0.78 0.84*

ENaOH_0.025M_2h_TP 0.95* 0.65 0.87 0.71 0.84*

ENaOH_0.025M_5h_TP 0.95* 0.59 0.89* 0.74 0.80*

ENaOH_0.025M_16h_TP 0.95* 0.63 0.89* 0.58 0.81*

ENaOH_0.05M_1h_TP 0.95* 0.64 0.85 0.52 0.86*

ENaOH_0.05M_2h_TP 0.94* 0.64 0.87 0.70 0.84*

ENaOH_0.05M_5h_TP 0.95* 0.72 0.90* 0.71 0.86*

ENaOH_0.05M_16h_TP 0.95* 0.68 0.92* 0.61 0.76

ENaOH_0.1M_1h_TP 0.95* 0.72 0.85 0.57 0.84*

ENaOH_0.1M_2h_TP 0.94* 0.77 0.88 0.72 0.81*

ENaOH_0.1M_5h_TP 0.95* 0.76 0.92* 0.76 0.74

ENaOH_0.1M_16h_TP 0.95* 0.59 0.90* 0.68 0.80*

NaOH_0.05M_0.5h_TP 0.91* 0.70 0.58 0.55 0.62

NaOH_0.05M_1h_TP 0.92* 0.73 0.66 0.47 0.79

NaOH_0.05M_1.5h_TP 0.92* 0.69 0.69 0.45 0.58

NaOH_0.05M_2h_TP 0.91* 0.74 0.68 0.48 0.70

NaOH_0.1M_0.5h_TP 0.91* 0.78 0.63 0.46 0.41

NaOH_0.1M_1h_TP 0.92* 0.60 0.72 0.46 0.43

NaOH_0.1M_1.5h_TP 0.92* 0.66 0.66 0.29 0.45

NaOH_0.1M_2h_TP 0.92* 0.82 0.79 0.44 0.58

NaOH_0.15M_0.5h_TP 0.92* 0.75 0.68 0.60 0.50

NaOH_0.15M_1h_TP 0.92* 0.70 0.70 0.62 0.73

NaOH_0.15M_1.5h_TP 0.92* 0.75 0.70 0.61 0.73

NaOH_0.15M_2h_TP 0.92* 0.72 0.78 0.60 0.79

NaOH_0.2M_0.5h_TP 0.92* 0.74 0.65 0.61 0.61

NaOH_0.2M_1h_TP 0.92* 0.72 0.72 0.35 0.63

NaOH_0.2M_1.5h_TP 0.92* 0.78 0.72 0.37 0.78

NaOH_0.2M_2h_TP 0.92* 0.84 0.69 0.62 0.85*

161

5.4.5 Relationships between soil test Phosphorus methods

The linear relationships between the amounts of P extracted by existing soil P extraction

methods and the RMS-P for entire soils across sites were significant at p< 0.05 (Figures

5.8-a, b and c). These strong linear relationships between these STP methods and

RMS-P suggest that they may extract P from the same soil P fractions and hence

should have a similar relationship with the amounts of total releasable P. Further,

significant linear relationships between these soil test methods indicates that the results

from one soil test P method can be converted to the other method by using their

regression equations.

As expected, a highly significant linear relationship was observed between RMS-P and

FeO- strips methods with the coefficient of determination (R2) of 0.93 (Figure 5.8-a),

because both these methods have a similar sink-based extracting mechanism. A highly

significant linear relationship was found between RMS-P and Olsen-P with R2 value of

0.94 (Figure 5.8-b); however, compared with the Olsen method, a smaller R2 value

(R2=0.86) was observed for the linear relationship between RMS-P and the Mehlich-3

(Figure 5.8-c). The R2 for RMS-P with Olsen P was significantly improved by applying a

quadratic model giving R2 value of 0.97 (Figure 5.8-b1); however, the relationships

between RMS-P and FeO-strips-P or between RMS-P and Mehlich-3-P did not improve

by applying a quadratic model (Figures 5.8-a1 and 5.8-c1).

For the newly proposed extraction methods, significant linear relationships were

observed between RMS method and the different concentrations of NaOH with EDTA

extractants with R2 values ranging from 0.89 to 0.91 [(Figures 5.9 (a to d), 5.10 (a to d),

and 5.11 (a to d)]. When the quadratic regression model was applied, the R2 in the

162

relationships between all the combinations of NaOH with EDTA and RMS-P were

significantly improved giving the R2 values ranging from 0.93-0.94 [(Figures 5.9 (a1 to

d1), 5.10 (a1 to d1), and 5.11 (a1 to d1)].

The R2 for the linear relationships between RMS and the different concentrations of

NaOH without EDTA extractants [(Figures 5.12 (a to d), 5.13 (a to d), 5.14 (a to d) and

5.15 (a to d)] were smaller (ranging from 0.83 to 0.86) than those obtained for the linear

relationships between RMS and the different concentrations of NaOH with EDTA

extractants. Among these extractants, both 0.15M NaOH and 0.2M NaOH extractants

showed greater R2 values with RMS method compared to the 0.05M NaOH and 0.1M

NaOH concentrations. When a quadratic regression model was applied for the

relationships between NaOH without EDTA extractants and RMS-P, the R2 values were

improved (ranging from 0.87 to 0.89) than for linear regressions. However, significant

improvements were observed only for some of these relationships (Figures 5.12-c and

d, 5.13-b and d, 5.14-d, 5.15-c).

Overall, the Olsen method had the greatest R2 for both linear (R2 = 0.94) and quadratic

(R2 = 0.97) regressions among existing soil P tests. For new methods, the comparisons

of linear regressions reveal that the most significant linear relationships (R2 = 0.91) with

RMS-P was with the 0.025M NaOH with EDTA extractant for 2, 5 and 16 hours shaking

periods. Instead of 5 and 16 hours, 2 hours can be considered as the most effective and

practical shaking period. For quadratic regressions, no such difference was found for R2

values among NaOH with EDTA extractants. Accordingly, 0.025M NaOH with EDTA

extraction with 2 hours shaking period can be considered as effective as a sink-based

163

RMS method for determining the total releasable P in that soil and it can be used for

predicting long-term soil P loss potential.

5.5. Conclusion

Amounts of P extracted by different tests varied widely. This is because different

extractants have varying ability to extract different portions of soil P as a result of their

different reactions with soil components controlling soil P availability. Mehlich-3

extracted a greater amount of P than Olsen and RMS extracted more than twice as

much P as did FeO-strips. The new methods extracted greater amounts of P than did

existing agronomic and environmental P tests. All combinations of NaOH with EDTA

extracted greater amounts of P compared to all combinations of NaOH without EDTA

indicating the inclusion of EDTA appeared to enhance the amounts of P extracted in

comparison with the NaOH alone. Among these existing and new methods 0.05M

NaOH with EDTA extractant gave the greatest extraction for all soils across the sites as

well as for soils within individual sites. However, Mehlich-3 extractant showed similar

levels of P extraction as RMS method did for Harrow, Swift current and Indian Head

soils, suggesting Mehlich-3- method was as effective as RMS method for measuring the

total releasable P in the soil.

The strong correlation between two environmental tests was found due to their common

sink-based extracting mechanisms. Amounts of P extracted by two agronomic tests

were highly correlated with RMS, indicating even though each of these reagents

extracted a different proportion of P, they both are capable of estimating the total

releasable P in the soil. The P extracted by all combinations of NaOH with EDTA and

without EDTA extractants were significantly correlated (at p ≤ 0.05) with RMS-P when

164

soils across all sites were considered. In addition, all the combinations of NaOH with

EDTA had better correlation with RMS-P than did NaOH alone for all soils across all

sites. However, less satisfactory relationships were found for soils from individual sites.

Very significant linear relationships (p ≤ 0.05) were observed between the RMS method

and all the new methods. However, 0.025M NaOH with EDTA extractant had the

strongest linear relationship (R2 = 0.91) with RMS-P. The coefficient of determinations

(R2) were significantly improved by adopting a quadratic model for all NaOH with EDTA

extractants and some of the NaOH alone extractants. Results confirm that 0.025M

NaOH with EDTA for 2 hours shaking period was as effective as a sink-based RMS

method for measuring total releasable P in the soil. This extraction combination might

be the most suitable test for predicting long-term P loss potential of agricultural soils.

165

y = 0.3041x + 8.5304 R² = 0.93

0

10

20

30

40

0 20 40 60 80 100

FeO

-str

ips-

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a) FeO-strips_P

y = 5E-05x2 + 0.2857x + 9.3738 R² = 0.93

0

10

20

30

40

0 20 40 60 80 100

FeO

-str

ips-

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a') FeO-strips_P

y = 0.569x - 5.616 R² = 0.94

0

20

40

60

0 20 40 60 80 100

Ols

en-P

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b) Olsen_P

y = -0.0009x2 + 0.8893x - 20.305 R² = 0.97

0

20

40

60

0 20 40 60 80 100

Ols

en-P

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b') Olsen_P

166

Figure 5.8 Linear (left) and Non-linear (right) relationships between soil P extracted by three existing soil P tests (FeO-,

strips-a, Olsen-b and Mehlich-3-c) and cumulative amount of P extracted by resin membrane strip method (Total

Releasable P).

y = 0.4407x + 23.755 R² = 0.86

0

20

40

60

80

0 20 40 60 80 100

Meh

lich

-3-P

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c) Mehlich-3_P

y = -5E-05x2 + 0.4596x + 22.887 R² = 0.86

0

20

40

60

80

0 20 40 60 80 100

Meh

lich

-3-P

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c') Mehlich-3_P

167

y = 5.1959x + 87.605 R² = 0.90

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a) ENaOH_0.025M_1h_TP

y = -0.0097x2 + 8.7623x - 75.948 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a') ENaOH_0.025M_1h_TP

y = 5.4326x + 85.645 R² = 0.91

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b) ENaOH_0.025M_2h_TP

y = -0.01x2 + 9.1164x - 83.29 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b') ENaOH_0.025M_2h_TP

168

Figure 5.9 Linear (left) and non-linear (right) relationships between soil P (mg P kg-1) extracted by 0.025M NaOH + EDTA,

and cumulative amount of P extracted by resin membrane strip method (Total Releasable P)

y = 5.3953x + 110.05 R² = 0.91

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c) ENaOH_0.025M_5h_TP y = -0.0098x2 + 8.9977x - 55.15

R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c') ENaOH_0.025M_5h_TP

y = 5.4212x + 115.72 R² = 0.91

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-16

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(d) ENaOH_0.025M_16h_TP

y = -0.0101x2 + 9.1426x - 54.934 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.02

5M

-16

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(d') ENaOH_0.025M_16h_TP

169

y = 5.7296x + 81.26 R² = 0.90

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.05

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a) ENaOH_0.05M_1h_TP y = -0.0117x2 + 10.058x - 117.25

R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.05

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = 5.8635x + 97.747 R² = 0.89

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.05

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = -0.0125x2 + 10.465x - 113.28 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.05

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


170



y = 5.8927x + 112.9 R² = 0.90

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.05

M-5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(c) ENaOH_0.05M_5h_TP

y = -0.0118x2 + 10.237x - 86.327 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.05

M-5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = 5.9268x + 122.84 R² = 0.90

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.05

M-1

6h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = -0.0111x2 + 10.018x - 64.777 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.05

M-1

6h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


171

y = 5.8677x + 48.935 R² = 0.90

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.1M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a) ENaOH_0.1M_1h_TP

y = -0.012x2 + 10.295x - 154.1 R² = 0.94

0

500

1000

1500

2000

2500

0 200 400 600

ENaO

H-0

.1M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = 5.954x + 59.214 R² = 0.89

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.1M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = -0.0125x2 + 10.565x - 152.23 R² = 0.93

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.1M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


172



y = 6.1442x + 74.242 R² = 0.89

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.1M

-5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c) ENaOH_0.1M_5h_TP

y = -0.0127x2 + 10.818x - 140.08 R² = 0.94

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.1M

-5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = 5.9941x + 90.069 R² = 0.89

0

500

1000

1500

2000

2500

3000

0 100 200 300 400 500

ENaO

H-0

.1M

-16

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = -0.0113x2 + 10.177x - 101.75 R² = 0.93

0

500

1000

1500

2000

2500

0 100 200 300 400 500

ENaO

H-0

.1M

-16

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


173

y = 5.0736x - 157.03 R² = 0.83

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.05

M-0

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a) NaOH_0.05M_0.5h_TP

y = -0.0111x2 + 9.1475x - 343.86 R² = 0.87

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.05

M-0

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a') NaOH_0.05M_0.5h_TP

y = 5.2963x - 164.79 R² = 0.84

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.05

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(b) NaOH_0.05M_1h_TP

y = -0.0104x2 + 9.1373x - 340.93 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.05

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(b') NaOH_0.05M_1h_TP

174

Figure 5.12 Linear (left) and non-linear (right) relationships between soil P (mg P kg-1) extracted by 0.05M NaOH and

cumulative amount of P extracted by resin membrane strip method (Total Releasable P)

y = 5.5024x - 165.12 R² = 0.84

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.05

M-1

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c) NaOH_0.05M_1.5h_TP

y = -0.0115x2 + 9.7439x - 359.63 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.05

M-1

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c') NaOH_0.05M_1.5h_TP

y = 5.5025x - 162.3 R² = 0.84

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.05

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(d) NaOH_0.05M_2h_TP

y = -0.0122x2 + 10.001x - 368.6 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.05

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(d') NaOH_0.05M_2h_TP

175

y = 5.6322x - 170.92 R² = 0.83

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.1M

-0.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = -0.0121x2 + 10.103x - 375.95 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.1M

-0.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a') NaOH_0.1M_0.5h_TP

y = 5.728x - 167.73 R² = 0.84

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.1M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b) NaOH_0.1M_1h_TP

y = -0.0111x2 + 9.8349x - 356.07 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.1M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


176



y = 6.2305x - 184.95 R² = 0.84

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.1M

-1.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = -0.0134x2 + 11.174x - 411.65 R² = 0.88

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.1M

-1.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(c') NaOH_0.1M_1.5h_TP

y = 5.9267x - 171.08 R² = 0.85

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.1M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(d) NaOH_0.1M_2h_TP

y = -0.0111x2 + 10.011x - 358.37 R² = 0.89

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.1M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


177

y = 5.3995x - 128.65 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.15

M-0

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = -0.0095x2 + 8.9069x - 289.49 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.15

M-0

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(a') NaOH_0.15M_0.5h_TP

y = 5.6497x - 133.25 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.15

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(b) NaOH_0.15M_1h_TP

y = -0.0109x2 + 9.6492x - 316.67 R² = 0.88

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.15

M-1

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


178


umulative amount of P extracted by resin membrane strip method (Total Releasable P)

y = 5.8528x - 137.39 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.15

M-1

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


y = -0.0114x2 + 10.068x - 330.71 R² = 0.88

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500NaO

H-0

.15

M-1

.5h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(c') NaOH_0.15M_1.5h_TP

y = 4.9826x - 99.017 R² = 0.86

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.15

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(d) NaOH_0.15M_2h_TP

y = -0.0075x2 + 7.7442x - 225.66 R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.15

M-2

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


179

y = 5.6017x - 137.28 R² = 0.84

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-0.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a) NaOH_0.2M_0.5h_TP y = -0.0113x2 + 9.7757x - 328.7

R² = 0.88

-500

0

500

1000

1500

2000

0 100 200 300 400 500

NaO

H-0

.2M

-0.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(a') NaOH_0.2M_0.5h_TP

y = 5.8794x - 145.78 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(b) NaOH_0.2M_1h_TP

y = -0.0107x2 + 9.8413x - 327.47 R² = 0.89

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-1h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


180



y = 6.139x - 151.42 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-1.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)


y = -0.0117x2 + 10.433x - 348.35 R² = 0.88

-500

0

500

1000

1500

2000

2500

0 200 400 600

NaO

H-0

.2M

-1.5

h-T

P (

mg

P k

g-1)

RMS-P (mg P kg-1)

(c') NaOH_0.2M_1.5h_TP

y = 5.9043x - 142.89 R² = 0.85

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)

(d) NaOH_0.2M_2h_TP

y = -0.0093x2 + 9.3292x - 299.95 R² = 0.87

-500

0

500

1000

1500

2000

2500

0 100 200 300 400 500

NaO

H-0

.2M

-2h

-TP

(m

g P

kg-1

)

RMS-P (mg P kg-1)


181

Chapter 6: Summary, Conclusions and Future Studies

6.1 Summary and conclusions

Freshwater eutrophication is accelerated by increased phosphorus (P) inputs, with a

large portion potentially from agricultural non-point sources. To reduce such negative

impacts of P on surface water quality, P inputs from agricultural lands are required to be

diligently managed to minimize them from entering into nearby water bodies. Hence, it

is imperative to identify the agricultural crop lands with high potential for P losses and

develop proper management strategies to minimize soil P loss. To accomplish this,

accurate measurements of soil P forms that related with loss from soil, their changes

and distribution in the soil profile are of prime importance.

The purpose of this thesis was to determine the effects of inorganic fertilizer (Triple

super phosphate and Ammonium nitrate) and different types of livestock manure

(different forms of swine manure and dairy manure slurry) applications on soil P forms,

their changes over time and distribution in the soil profile (objectives one and two). In

addition, a third objective was to develop a new soil P test for assessing the long-term P

loss potential of agricultural soils. To accomplish these objectives, this thesis is

comprised of three studies and their findings are summarized as follows.

The findings of both swine manure and dairy manure slurry studies indicate that

regardless of the P source, accumulation of P from external P sources was mostly in the

uppermost soil layer (0-15 cm). Generally, P has little mobility in soils because of their

strong reactivity with soil components and tendencies to be adsorbed on soil colloid

surfaces and to form insoluble compounds with bi- and trivalent cations (Stevenson and

182

Cole 1999). In addition, this accumulation of soil P in the surface soil may also be

related to greater biological activity in surface soil layers where P was applied with

manure. Phosphorus accumulation in 0-15 cm soil depth is beneficial in terms of crop

growth especially in soils that are deficient in P. However, this accumulation is

accompanied by increases in soil-test P and degree of soil P saturation (Zheng et al.

2001) in the surface soil horizon that might increase the risk of P losses to adjacent

water bodies.

The results also confirmed that both types of manure had a very rapid impact on the

labile-Pi fractions (H2O-Pi and Bicarb-Pi) in soils. The reason may be due to the higher

Pi content in added manure sources. Further, the addition of organic matter to the soil

through manure application would enhance microbial and enzyme activities, which in

turn would increase the rate of biologically-mediated turnover of organic P into inorganic

P. In addition, competition between manure P and organic acids which were produced

during microbial degradation of organic matter, for adsorption sites on soil could lead to

a greater amount of manure P to be more available. Although increases of these labile

forms of P in soil increase plant availability of P, it also increases the risk of P loss and

contamination of adjacent water-bodies, causing environmental problems.

Considering the soil P pools, in the soils treated with dairy manure slurry and all forms

of swine manure, the moderately labile P fraction (NaOH-1-P) was the largest P fraction

compared to all other P fractions. NaOH extractable Pi is known to be the P

chemisorbed on amorphous Al and Fe minerals (Tiessen et al. 1984). Therefore, these

higher levels might be related to larger Al and Fe contents of these soils. Research has

found that the moderately labile P pool can act as a source of plant available P in

183

absence of P inputs and also as a sink for P added in the case of P application in

excess of crop removal (Beck and Sanchex 1994). However, in the study with dairy

manure slurry application (Chapter four), the moderately labile P fraction did not show

significant changes. The potential explanation for the lack of significant changes of

moderately labile P fraction is, the rates of P application with dairy manure may not be

enough to considerably increase this P fraction, or the drawdown of the labile P fraction

due to plant P uptake may not be sufficient enough to result in reduction in this P

fraction. Therefore, moderately labile-P fraction neither acted as a P sink nor a P source

in the dairy manure study. In the study with swine manure and fertilizer P applications

(Chapter three), moderately labile P fraction increased significantly in year of P

application and also in the following year as a residual effect, indicating that moderately

labile P pool acted as the major sink for added P from swine manure and inorganic

fertilizer P. This is probably due to the formation of amorphous or crystalline Fe and Al

phosphates favoured by low pH. This low pH could be due in part to the application of

triple superphosphate that diminishes pH when solubilized, and also due in part to the

higher carbon availability through crop residues returning to the soil that releases acids

when degraded by microorganisms. These results show concomitant increases in labile

Pi and moderately labile Pi fractions along with P addition through different swine

manure forms and inorganic fertilizer P. Increases in both fractions indicate that these

fractions are an important sink of added P; it also suggests that P may be added in

excess of plant removal in this study.

The findings of the study with dairy manure slurry show that, moderately stable P

fraction (HCl-P) significantly increased with dairy manure slurry application, indicating

184

that a considerable amount of P was added from dairy manure slurry to this P pool. The

study with swine manure or fertilizer P additions did not have significant impact on

moderately stable P fraction in the year of application or in the following year. This

indicates there were no measurable immediate or residual effects on moderately stable

P fraction from all these P sources. Generally, moderately stable P fraction is presumed

to be less use to plants and remains unaffected under normal conditions, because this

P fraction represents primary minerals in soils. This disparity of behaviour of moderately

stable P fraction with different manure and inorganic fertilizer P additions may be due to

the nature of the P sources and also due to the soil types and climatic conditions.

The organic P fraction of these soils did not change significantly with manure or fertilizer

application in both swine manure and dairy manure studies. One reason is that manure

contains relatively higher amounts of P in inorganic form and inorganic fertilizer P is

entirely in inorganic form. It has been reported in the literature that over 60% of P in the

manure products is initially present in the inorganic form (Barnett 1994; Sui et al. 1999).

Thus, application of manure products results in accumulation of P mainly in the

inorganic P pool (Gatiboni et al 2008) and it may not readily be incorporated into the

organic P pool to show significant impact. However, soils in the control plots showing

relatively higher organic P levels compared to the inorganic P levels indicate that the

soil itself contained a considerable amount of organic P and external P sources

contribute a significant amount of inorganic P to the treated soils.

In both studies, stable P forms (NaOH-2-Pi and Po and Res-P) did not show any

response to manure or fertilizer application, suggesting these sparingly soluble P

fractions may not likely contribute substantially to meet plant P needs or loading to

185

surface water. Typically, the pools that are less available or stable are weakly

responsive to manure additions (Leinweber et al. 1999). However, some studies have

reported conflicting results which could be attributed to soil type and manure source.

Nevertheless, changes to these fractions may likely occur slowly and may take years to

become significantly quantifiable.

Generally, the forms and availability of P in soil following manure additions are

dependent to a large extent on the source of P applied. Most of the changes in manure-

treated soils are attributable to an increase in soil organic matter, because animal

manure, especially solid manure and manure compost contains a large amount of

organic matter. This increased organic matter in manure-treated soils covered clay

mineral surfaces or chelated metal ions, which could prevent Pi from adsorption or

precipitation by clay minerals or metal ions (Tang et al. 2006). Thus, the inorganic P

content in soil solution becomes higher in soils applied with manure with high organic

matter content.

Overall, the increases in labile and moderately labile P with manure applications are

beneficial for crop production, but excess amounts could negatively impact the

environment. The results of these studies confirm that the P applied with manure can

have both short-term (increases in labile-P fraction) and long-term (increases in

moderately stable-P fractions) impacts on soil P availability and P loss. However, in

both of these manure studies, other soil factors which influence soil P status need to be

considered, such as microbial activities, different agro-climatic factors (such as rainfall

intensity, evaporation and thermal effects) and concentrations of other soil constituents

that may have influence on soil P dynamics; this could be the focus of future research.

186

As per the third objective, relationships between existing agronomic (Mehlich-3-P and

Olsen-P) and environmental (FeO-strips-P and Resin membrane strips, RMS-P) soil P

tests and new soil P tests (different concentrations of NaOH with EDTA and without

EDTA with different shaking periods) were assessed. The results show that the

amounts of P extracted by Mehlich-3 extractant were almost equal to the cumulative

RMS-P amounts for soils from Harrow and both sites in Saskatchewan. This suggests

that Mehlich-3 method was as effective as an RMS method for measuring the total

releasable P in these soils. Further, newly proposed reagents extracted more P than did

existing agronomic and environmental P tests. The reason for higher extraction is NaOH

changes the physical structure of organic molecules enhancing their solubility. Given

that, at high pH, many organic functional groups are ionized and the increased charge

density leads to increased P solubility. In addition, the inclusion of EDTA appeared to

enhance the amounts of P extracted compared to the NaOH without EDTA. Because

EDTA, a strong chelating ligand, complexes paramagnetic cations such as Fe and Mn

in the extract, it thereby increases soil P extraction efficiency and the diversity of P

compounds extracted (Bowman and Moir, 1993).

Considering the relationships between soil P tests, a strong correlation was observed

between two environmental P tests because of their common sink-based extracting

mechanisms. Further, the amounts of P extracted by two agronomic tests were also

highly correlated with RMS, indicating even though each of these reagents have

markedly different extraction mechanisms and extract different proportions of P, they all

are capable of estimating the total releasable P of that soil. These findings agree with

the findings of several past studies which have been conducted in different regions to

187

analyze the relationship between amounts of P extracted by some environmental soil P

tests and P extracted by agronomic soil P tests. Their results demonstrated that the P

values obtained by agronomic P tests were well correlated with those of environmental

P tests (Atia and Mallarino 2002; Kleinman and Sharpley 2002; Maguire and Sims

2002). However, the relationships varied with soil properties such as soil type, soil pH,

particle size distribution and mineralogy and management practices, which are known to

influence soil P sorption. Hence, for regions with contrasting soils, only a few

generalizations can be made about the relationship between agronomic and

environmental P tests. Although P extracted by all the combinations of NaOH with

EDTA and NaOH without EDTA extractants were significantly correlated with RMS-P, all

the combinations of NaOH with EDTA extractants showed better correlation with RMS-P

than did NaOH alone. This suggests that NaOH with EDTA extractants are as effective

as RMS for measuring total releasable P of the soil and may have potential for use as

an environmental soil P test for identifying soils with long-term P loss potential.

6.2 Future Studies

To determine the residual effects of manure and inorganic fertilizer P on changes of soil

P forms, future studies need to be considered for a few more years with manure and

fertilizer applications on every other year and observe the behaviour of P fractions and

P distribution in the soil profile within these multi years. In order to further improve the

accuracy and appropriateness of this new soil P testing method as environmental soil P

test for prediction of long-term P loss potential, further work should be conducted using

a larger number of samples from different soil types from contrasting ecological regions

throughout Canada, and need to consider some other factors which influence on soil P.

188

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