APPLICATIONS OF ACTIVE TRACER TECHNIQUES IN SOIL ...

APPLICATIONS OF ACTIVE TRACER TECHNIQUES IN SOIL EROSION AND CATCHMENT SEDIMENT SOURCE INVESTIGATIONS

Submitted by Philip Greenwood to the University of Exeter as a thesis for the degree of Doctor of Philosophy in Geography, January 2010

This thesis is available for library use on the understanding that it is copyright material and that no quotation from the thesis may be published without proper acknowledgement

I certify that all material in this thesis which is not my own work has been identified and that no material is included for which a degree has previously been conferred upon me

……………………………

ABSTRACT

This thesis outlines the methodology and results from five different soil erosion and catchment sediment source investigations using the artificial radionuclides,

134

Cs

and 60Co as sediment tracers. The overall aims of the thesis were firstly, to evaluate the capabilities of the two radionuclides as tracers of fine-sediment, and secondly, to use the findings from each of the investigations to provide an improved understanding of some of the mobilisation and transfer mechanisms interposed between upslope erosion and downslope sediment yield. For the first investigation, a technique was developed to determine the extent of remobilisation of recently-deposited sediment during a sequence of overbank flood events on river floodplains in two different catchments. The results from both sites confirmed the importance of remobilisation and suggested that variations in the magnitude were influenced by physical controls such as the floodplain geometry, topography, surface roughness and inundation magnitude. The findings have implications for estimating the deposition of sediment over medium-term timescales on river floodplains using established fallout radionuclide techniques. For the second investigation, a technique was developed to quantify the soil translocation distance and vertical mixing within the till-zone during chisel-plough tillage performed upslope at a constant speed and depth. The experiment was conducted on unique experimental tillage-erosion facilities and the results indicated that sediment displacement values were compatible with studies derived using other types of tracer and conducted under similar tillage parameters. Evidence was also obtained of a net downward movement of soil within the till-zone and this could result in the uneven distribution of surface-applied soil-amendments, or the accumulation of soilcontaminants. For the third investigation, a technique was developed to determine the extent of sediment redistribution on small areas of livestock-poached pasture. Although sediment redistribution was documented on all livestock-poached areas investigated, erosion exceeded deposition and this resulted in a net loss of sediment from all sites. This finding has potential implications for environmentally sensitive areas located in close proximity to livestock-poached pasture. For the fourth investigation, a technique was developed to determine the fate of sediment from dispersed earthworm casts. The technique was applied in-situ to a small area of pasture and used natural weather events as the eroding agent. The results 2

indicated that dispersed casts represent a substantial source of eroded sediment and factors controlling the rate of erosion include topography, the availability of fresh casts and the timing of rainfall events. The technique was also applied on tilled soil using an erosion plot and simulated rainfall as the primary eroding agent. The findings indicated that the removal of dispersed casts from unconsolidated soil was considerably less than the removal of dispersed casts after the formation of a surface crust and this difference was attributed to changes in soil surface conditions before and after the formation of the crust. The results also indicated that sediment from casts dispersed both before and after the formation of the crust represented a substantial source of eroded material, and highlights the susceptibility of a crusted soil to mobilisation during erosion events. For the final investigation, a technique was developed for documenting sediment source changes during the transition from inter-rill to rill erosion on cultivated hillslopes. Experiments were conducted using an erosion plot and simulated rainfall and the technique was applied to eight different agricultural soils. Statistically significant differences were recorded between the grain size composition of pairs of inter-rill and rill eroded sediment samples by the majority of soils. These findings have important implications with regard to predicting and estimating the off-site conveyance of soilsorbed nutrients and contaminants by particular mobilisation processes during erosion events, since their transfer is closely associated with the movement of fine-sediment. The results from the five investigations adequately demonstrate the ability of the artificial radionuclides,

134

Cs and

60

Co, to accurately quantify the movement of small

quantities of fine-sediment with a level of sensitivity that far exceeds the majority of other tracers.

3

TABLE OF CONTENTS

Abstract

2

List of Contents

4

List of Figures

11

List of Tables

16

Acknowledgements

19

Dedication

21

LIST OF CONTENTS

CHAPTER 1: Introduction, Literature Review, Aims & Objectives, Research Strategy & Thesis Structure

22

1.1 Introduction

22

1.1.1 The Importance of Sediment Studies

22

1.1.2 Sediment Tracers

23

1.1.3 Passive & Active Tracers: A Working Definition

25

1.2 Literature Review

26

1.2.1 Passive Tracers: A Review

26

1.2.2 Active Tracers: A Review

29

1.3 Aims & Objectives

34

1.4 Research Strategy

35

1.4.1 Added Value

37

1.4.2 Convenience & Availability of Information

38

1.4.3 Scale

39

1.5 Thesis Structure

41

CHAPTER 2: The Radionuclide Selection Process, Origins & Characteristics, Health & Safety & the Principles of Gamma Spectrometry

43

2.1 Radionuclide Selection Process

43

2.2 Cobalt-60 & Caesium-134

45

2.2.1 60Co: Origin & Environmental Characteristics

45

2.2.2 134Cs: Origin & Environmental Characteristics

49

2.3 Health & Safety Issues & Gamma Radiation Measurement 4

52

2.4 Gamma Spectrometry

53

2.4.1 The Basic Principles of Gamma Spectrometry

53

2.4.2 The Major Components of a Gamma Spectrometer

54

2.4.3 Gamma Ray Interactions with Matter & Detector Efficiencies

58

2.4.4 The Laboratory-Based Gamma Spectrometers & Radiometric Equations 2.4.5 The Field Detector

59 62

2.5 The Radionuclides: Purchasing & Provenance

65

CHAPTER 3: Assessing the Extent of Floodplain Sediment Remobilisation During Single Overbank Flood Events

66

3.1 Overview

66

3.2 Introduction

66

3.3 Method

68

3.3.1 Experimental Design: Overview

68

3.3.2 Tracer Selection

68

3.3.3 The Approach

68

3.3.4 The Study Sites and Measurement Programme

72

3.3.5 The River Taw: General Catchment Characteristics

72

3.3.6 The Study Site

72

3.3.7 The River Culm: General Catchment Characteristics

73


74

3.3.9 Measurement Programme

75

3.4 Results

75

3.4.1 River Taw Floodplain: First Inundation Event

76

3.4.2 Second Inundation Event

77

3.4.3 Magnitude of Areal Activity Density Reduction

77

3.4.4 Spatial Variability

78

3.4.5 River Culm Floodplain: First Inundation Event

80

3.4.6 Second Inundation Event

81

3.4.7 Magnitude of Areal Activity Density Reduction

82

3.4.8 Spatial Variability

83

3.5 Interpretation of Results

85

3.5.1 The Extent of Remobilisation 5

85

3.5.2 The Magnitude of Remobilisation

87

3.5.3 Physical Controls

90

3.6 Assessment of the Approach

91

3.7 Conclusion

93

CHAPTER 4: Determining Soil Translocation Distance and Vertical Mixing Within the Till-Zone During Upslope Tillage

95

4.1 Overview

95

4.2 Introduction

96

4.2.1 Measuring Tillage Erosion

96

4.2.2 Chisel Plough Tillage: Soil Movement & Mixing

97

4.3 Method

99


99

4.3.2 The Soil Bin

99

4.3.3 Plot Construction, Dimension & Location

101

4.3.4 The Stratified Configuration of the Radionuclides

103

4.3.5 The Labelling Procedure and Radionuclide Activity

105

4.3.6 Choice & Description of the Tillage Implement

106

4.3.7 Bulk Soil Characteristics

107

4.3.8 Tillage Configuration & Characteristics

109

4.3.9 Sampling & Removal of Tilled Soil

109

4.3.10 Radiometric Protocol

112

4.3.11 Estimating Soil Translocation

112

4.3.12 Estimating Soil Mixing

114

4.4 Results

115

4.4.1 Radionuclide Recovery

116

4.4.2 Backward Soil Translocation

116

4.4.3 Upslope Soil Translocation

116

4.4.4 Soil Movement & Mixing

117


121

4.5.1 Soil Translocation: Mass

121

4.5.2 Constant Speed & Depth of Tillage

123

4.5.3 Multiple Tillage Passes

123

4.5.4 Backward Translocation

123

4.5.5 Patterns of Soil Movement & Mixing

124

6

4.5.6 Implications of Uneven Soil Mixing 4.6 Assessment of the Approach

125 125

4.6.1 The Sampling Strategy

125

4.6.2 Soil Characteristics: Experimental Versus Field

126

4.6.3 Fluctuating Tillage Depths

126

4.6.4 Use of Additional Radionuclides

127

4.6.5 Corroborating the Results

127

4.6.6 Net Soil Movement

127

4.7 Conclusion

128

CHAPTER 5: Assessing the Extent of Sediment Redistribution on Livestock-Poached Pasture

129

5.1 Overview

129

5.2 Introduction

129

5.3 Method

133


133

5.3.2 Experimental Design

133

5.3.3 Tracer Preparation

133

5.3.4 The Labelling Procedure

133


134

5.3.6 Plot Quantity, Dimensions & Locations

136

5.3.7 Installation, Duration & Re-measurement Programme

137

5.3.8 The Radiometric Protocol

137

5.3.9 Shallow Coring of a Control Plot

138

5.3.10 The Calibration Procedure

139

5.4 Results

142

5.4.1 Cumulative Rainfall

143

5.4.2 Active Plots: Soil Redistribution Depths

143

5.4.3 Active Plot 1

143

5.4.4 Active Plot 2

144

5.4.5 Active Plot 3

145

5.4.6 Active Plot 4

146

5.4.7 Active Plot 5

147

5.4.8 Active Plot 6

148

5.4.9 Soil Redistribution Data Summarised & Extrapolated

149

7

5.4.10 Soil Redistribution & Slope Gradient

150

5.4.11 Soil Redistribution & Cumulative Rainfall

150


151

5.5.1 General Interpretations

151

5.5.2 Environmental Influences on Rates of Erosion

151

5.5.3 Implications of the Results

153


153

5.6.1 Soil Erosion or Flux

153

5.6.2 The Role of Vegetation as a Buffer of Mobilised Sediment

154

5.6.3 Site Specific Characteristics

154

5.6.4 Constraints on Plot Size / Area

155

5.5.5 Uncertainty in the Calibration Procedure

155

5.7 Conclusion

156

CHAPTER 6: Determining the Fate of Earthworm Cast-Eroded Sediment on Agricultural Land

158

6.1 Overview

158

6.2 Introduction

159

6.3 Method

162


162


162

6.3.3 Collection of Earthworm Casts

162

6.3.4 Labelling of Earthworm Casts

163

6.3.5 Verification of the Labelling Procedure

164

6.3.6 Radiometric Data Processing

167

6.3.7 In-Situ Experiment on Pasture

167

6.3.8 The Erosion Plot

167

6.3.9 Bulk Soil Parameters

168

6.3.10 Cast Mass, Activity Concentration & Placement

169

6.3.11 The Monitoring Campaign

171

6.3.12 Runoff Eroded Sediment: Recovery & Processing

171

6.3.13 Bulk Soil: Recovery & processing

172

6.3.14 Radiometric Analysis

172

6.3.15 Mass Balance

172

8

6.3.16 Ex-Situ Experiment on Cultivated Soil

173

6.3.17 The Erosion Plot

173

6.3.18 Bulk Soil Parameters

174

6.3.19 134Cs-Labelled Casts: Placement & 1st Simulated Rainfall

175

6.3.20 60Co Labelled Casts: Placement & 2nd Simulated Rainfall

176

6.3.21 Bulk Soil: Recovery & Processing

176

6.3.22 Radiometric Assay

176

6.4 Results

177

6.4.1 In-Situ Pasture

177

6.4.2 Ex-Situ Cultivated

183


187

6.5.1. Pasture

187

6.5.2 Cultivated Soil

189


191

6.6.1 Uncertainty in the Labelling Technique

191

6.6.2 Cast Durability and Resilience to Disaggregation

191

6.6.3 Runoff Eroded Samples: Volume and Quantity

192

6.6.4 Reduced Crusting by Earthworms

192

6.6.5 Scale

192

6.7 Conclusion

193

CHAPTER 7: Documenting Sediment Source Changes During the Transition from Inter-Rill to Rill Erosion

195

7.1 Overview

195

7.2 Introduction

195

7.3 Method

199


199


199

7.3.3 Soil Characteristics & Background Information

204

7.4 Results

208

7.4.1 Identifying the Transition from Inter-Rill to Rill Erosion

209

7.4.2 Sediment Partitioning Using a Mass Balance

210

7.4.3 A Comparison of Particle Size Distributions

213

7.4.4 Enrichment Ratios

215

9


217

7.5.1 General Interpretations of the Results 7.6 Assessment of the Approach

217 220

7.6.1 Mass Balance: Uncertainty

221

7.6.2 Limitations of the Sampling Strategy

221

7.6.3 The Erosion Plot: Dimensions, Scale & Edge Effects

222

7.6.4 Reduced Aggregate Stability

223

7.7 Conclusion

223

CHAPTER 8: Synthesis and Conclusion

225

8.1 Chapter Content

225

8.2 Principal Findings, Limitations & Scope to Extend the Existing Work

225

8.2.1 Chapter 3

225

8.2.2 Chapter 4

226

8.2.3 Chapter 5

228

8.2.4 Chapter 6

229

8.2.5 Chapter 7

230

8.3 General Conclusions

232

8.3.1 Advantages

232

8.3.2 Disadvantages

233

8.4 New Directions & Potential Research Areas

234

8.5 Conclusion

234

References

236

10

LIST OF FIGURES

CHAPTER 1 1.1

A conceptual framework to identify desired research outcomes and potential research themes

37

CHAPTER 2 2.1

A simplified schematic of a Canberra HPGe detector and Vertical Slimline Cryostat cooling system

2.2

55

A simplified schematic diagram showing the main electrical components of a gamma spectrometry system

56

2.3

Two laboratory-based gamma spectrometers

57

2.4

A view inside a lead shield showing the copper liner and

2.5

end-cap at the end of the detector head

58

The field detector mounted in an adjustable support cradle

64

CHAPTER 3 3.1

Deposition of labelled sediment during a simulated overbank flood event

69

3.2

Using the in-situ gamma spectrometer

70

3.3

Diagrammatic representation of the analytical protocol

71

3.4

A map showing the locations of the two floodplains and the position of plots on the River Taw floodplain

3.5

A map showing the position of plots on the River Culm floodplain

3.6

74

Changes in the AAD values for the River Taw floodplain after the first inundation event

3.7

76

Changes in the AAD values for the River Taw floodplain after the second inundation event

3.8

73

77

The relative magnitude of the reduction in AAD values for the River Taw floodplain and the proportion of remaining inventory after both events

78 11

3.9

Proportional symbols based on reduction in AAD values after the first inundation event for the River Taw floodplain

3.10

79

Proportional symbols based on reduction in AAD values after the first and second inundation events for the River Taw floodplain

3.11

80

Changes in the AAD values for the River Culm floodplain after the first inundation event

3.12

81

Changes in the AAD values for the River Culm floodplain after the second inundation event

3.13

82

The relative magnitude of the reduction in AAD values for the River Culm floodplain and the proportion of remaining inventory after both events

3.14

83

Proportional symbols based on reduction in AAD values after the first inundation event for the River Culm floodplain

3.15

84

Proportional symbols based on reduction in AAD values after the first and second inundation events for the River Culm floodplain

3.16

85

The swale on the downstream part of the River Taw floodplain shown in flood

3.17

86

The swale on the downstream part of the River Taw floodplain shown after a large flood event

3.18

Fresh overbank sediment deposited onto foliage and the floodplain surface

3.19

86

88

The inundated River Culm floodplain site showing wind-induced waves

90

4.1

A side view of the soil bin

99

4.2

An upslope view of the soil bin

100

4.3

The lateral and vertical position of the tillage implement

101

4.4

A plan-view of the soil bin showing the position of the

CHAPTER 4

plot and the path of tillage 4.5

102

The use of a base-plate was used to relocate the plot and also provide a means of determining the exact tillage translocation distance

103 12

4.6

Filling of the plot with labelled soil

4.7

A conceptual diagram showing the stratified configuration

104

of the two radionuclides within the plot prior to tillage

105

4.8

An example of a Chisel plough

106

4.9

A 0.325 m wide duck-foot tillage attachment from a chisel plough

107

4.10

Particle size characteristics of the bulk soil

108

4.11

A plan-view showing the sampling-strategy

110

4.12

Removal of the tilled soil as cores

111

4.13

A plan-view showing the cumulative spatial distribution of both radionuclides after tillage

4.14

115

A comparison between the observed and predicted spatial distribution of the radionuclide after tillage

4.15

The relative proportions of unlabelled and labelled soil in cores located at rows X = 0.05 m to X = 0.35 m

4.16

116

119

Tillage parameters depth and speed and the soil translocation values calculated as ratio values

121

CHAPTER 5 5.1

Poaching around focal points within a field, such as gateways, feed-mangers and water troughs

5.2

Slurrified soil caused by excessive poaching on a waterlogged soil

5.3

130

131

An adjustable cradle was used to suspend the raindrop simulator during the direct labelling of livestock-poached plots

134

5.4

Map showing the location of the study site

135

5.5

Map showing the site layout and locations of the plots

136

5.6

A schematic of the analytical grid-system for each plot

138

5.7

A conceptual diagram showing a method for quantifying rates of soil redistribution on undisturbed soils

140

6.1

Earthworm egesting a surface cast

160

6.2

Earthworm casts on field of pasture

163

6.3

Labelling of earthworm casts by immersion in radionuclide

CHAPTER 6

13

Solution 6.4

164

A trial to disintegrate labelled earthworm casts spread across an erosion plot lined with polythene in order to evaluate the labelling procedure

165

6.5

Partially disintegrated earthworm casts on an erosion plot

165

6.6

Scattergraph to demonstrate the uniformity of the labelling procedure

166

6.7

The in-situ erosion plot viewed from upslope

167

6.8

An upslope view of the in-situ erosion plot

168

6.9

Physical characteristics of the surrounding soil on pasture

169

6.10

Schematic plan-view showing the layout of the erosion plot and the relative locations of each batch of casts

170

6.11

The erosion plot used in the experiment on cultivated soil

174

6.12

Physical characteristics of the bulk soil used in the experiment on cultivated land

6.13

175

Labelled casts were evenly spread across the erosion plot prior to their disintegration and dispersal by simulated rainfall

6.14

Temporal variations in sediment concentration measured from the in-situ plot

6.15

178

The relative proportions of labelled and unlabelled sediment were partitioned using uncorrected radiometric data

6.16

181

The (corrected) partitioned data was calculated as sediment concentration (g l-1) in runoff

6.18

179

The relative proportions of labelled and unlabelled sediment were partitioned using corrected data

6.17

176

182

Partitioned data showing the relative proportions of labelled and unlabelled sediment removed over two simulated rainfall events

6.19

185

The partitioned data was converted to percentage values to determine the contribution of labelled and unlabelled soil in each sample

6.20

186

Differences in the rate of inventory removal for both radionuclides during the two rainfall events

14

187

CHAPTER 7 7.1

The erosion plot showing the profiled plate and collection trough

7.2

200

Schematic cross-sectional diagram of the tracer configuration in the erosion plot

7.3

201

The profiled and textured soil surface before a simulated rainfall event

7.4

202

The locations of the sites from where soil samples were collected

205

7.5

Breaky Bottom Vineyard

207

7.6

Evidence of rill erosion from site CC

207

7.7

Evidence of soil erosion in runoff leaving site OL

208

7.8

Close-up of soil erosion in runoff leaving site OL

208

7.9

An example of a rill system in the erosion plot after simulated rainfall

7.10

208

An example of another rill system in the erosion plot after simulated rainfall

7.11

208

Changes in the radionuclide concentrations indicating the transition from inter-rill to rill erosion for soil RT

7.12

The relative proportions of sediment removed in surface runoff by inter-rill and rill erosion for soil RT

7.13

212

An example of the cumulative frequency grain size distributions of eroded sediment samples and parent material for soil OL

7.14

209

214

A comparison of enrichment ratios from pairs of eroded Samples for all soils

216

15

LIST OF TABLES

CHAPTER 1 1.1

A summary of selected investigations showing a comparison between the physical characteristics of point-tracers and the host soil

28

CHAPTER 2 2.1

A checklist used to identify two candidate radionuclides as sediment tracers

2.2

44

A selection of literature showing the wide variety of applications where 60Co has been used as a tracer of environmental processes

2.3

48

A selection of literature showing the wide variety of applications where 134Cs has been used as a tracer of environmental processes

51

CHAPTER 3 3.1

Inundation dates, peak discharges and the timings and duration of re-measurements at each inundation event for the River Taw floodplain

3.2

75

Inundation dates, peak discharges and the timings and duration of re-measurements at each inundation event for the River Culm floodplain

3.3

75

Correlation coefficients for the relationships between the reduction in AAD values of the active plots and two potential controlling variables for the River Taw floodplain

3.4

79

Correlation coefficients for the relationships between the reduction in AAD values of the active plots and two potential controlling variables for the River Culm floodplain

3.5

84

The percentage proportions of active plots on both floodplains recording statistically significant reduction in AAD values 16

86

3.6

A summary comparison of the mean magnitude of remobilisation over both events for both floodplains and also the percentage values of the remaining inventory after the second event

87

CHAPTER 4 4.1

A list of studies to demonstrate the wide range of results that were obtained when attempting to determine the potential erosivity of chisel plough tillage

98

4.2

A summary of detector efficiencies

112

4.3

A summary of the mean and median tillage translocation values and the tillage parameters, speed and depth, which were used to calculate the soil flux values

4.4

117

Soil displacement values for 12 cores recovered from within the original boundary of the plot indicating the upward or downward movement

4.5

120

A comparison of the findings of this investigation with other studies

122

5.1

Background information for each active plot

137

5.2

A summary of control plot inventory losses

142

5.3

A summary of data generated by Active Plot 1 over the

CHAPTER 5

three re-measurements 5.4

144

A summary of data generated by Active Plot 2 over the three re-measurements

5.5

145


5.6

146


5.7

147


5.8

148


5.9

149

A summary of soil redistribution data for all six active 17

plots recorded over the three re-measurements 5.10

Correlation coefficients for the net depth of erosion and the gradient of each plot

5.11

150

150

Correlation coefficients for the net depth of erosion and cumulative rainfall

151

CHAPTER 6 6.1

A summary of data used to evaluate the quality of the technique to uniformly label casts

166

6.2

A summary of data derived from the monitoring campaign

171

6.3

A summary of key data calculated from information derived from the monitoring campaign

6.4

A summary of the initial cast mass, the initial and recovered activities to sediment ratios for both radionuclides.

6.5

183

A summary of the initial cast mass, the initial and recovered activities / sediment ratios

6.7

183

A summary of the partitioned data obtained from samples of runoff during two rainfall events

6.8

178

A summary of the corrected partitioned data for the experiment undertaken on pasture.

6.6

177

184

A summary of the data relating to the removal of labelled earthworm casts by surface runoff over two simulated rainfall events.

184

7.1

Background information for each site

205

7.2

Textural characteristics for all soils

206

7.3

A summary of the results from the partitioning procedure for

CHAPTER 7

each soil 7.4

211

A summary of information relating to pairs of eroded sediment samples selected for statistical analysis

7.5

A summary of the results from the statistical analysis for pairs of samples for each soil

7.6

213

215

A summary of the maximum particle size range of the eroded sediment samples for each soil 18

219

ACKNOWLEDGEMENTS

My time at Exeter University has passed extremely quickly and many people have provided me with invaluable guidance, support and friendship during this period. I would like to express my sincere gratitude to Professors Des Walling and Tim Quine for giving me the opportunity to undertake this research, and to thank them for their patience, for their words of encouragement and for imparting a little of their academic wisdom on to me. I am also indebted to the Natural Environment Research Council for providing the funding (studentship number NER/S/A/2004/12291A) which enabled me to undertake this research. I would like to acknowledge the assistance and support of numerous technical staff within the Geography Department. These include Dr. Yusheng Zhang, Mrs. Sue Frankling, Mr. Neville England, Mrs. Sue Rouillard and Mrs. Helen Grapes. I would especially like to thank Mr. Jim Grapes for his advice regarding gamma-spectrometry, and for his assistance with the often temperamental field-detector. I am grateful to the Environment Agency (EA) for supplying rainfall and river flow data, to the National Trust (NT) for allowing access to land adjacent to the River Culm and especially to Mr. John Snell of Newnham Barton Farm in Kings Nympton, mid-Devon, for granting me unrestricted access onto his land upon which a large proportion of my field-based research was undertaken. I would also like to thank the Department of Soil Science at the University of Manitoba (U of M), Canada, for allowing me to undertake part of my research using their experimental facilities. In particular, I would like to express my sincere gratitude and thanks to Dr. David Lobb who made my research in Canada possible, for his hospitality, and also for facilitating a trip to Churchill which enabled me to fulfil two lifelong ambitions; to photograph polar bears in their natural environment and to view the Aurora borealis in all its stunning glory: the memories of both will stay with me for a long, long time. I would also like to thank Drs. Li Sheng, Kevin Tiessen and Mr. Alex Koiter for their help and assistance in making the experiment operational and to Mrs. Eva Slavicek for overseeing the radiometric analyses. I would also like to thank Drs. Tim Papkiyriakou and Laura Simm for the sailing experience on Lake Winnipeg and Dr. Mario Tenuta and Alex and Harmony Koiter for their company during a wilderness

19

walking / camping trip through the beautiful Whiteshell Provincial Park, which straddles the border of Manitoba and Ontario. I would like to also like to acknowledge the moral support and camaraderie given by my many fellow PhD. candidates, to Penny Briars for asking me all sorts of questions about the nature of my research during the 2 ½ years that I lodged with her in the picturesque village of Kings Nympton, mid-Devon, and to myriad others that I have met during my time at Exeter University and its environs, many of whom have become and will hopefully remain firm friends. I would lastly like to acknowledge the support and encouragement given to me by my partner, Florence (Flo) Bottin, and by my friends and family back home; it really has been very much appreciated.

20

Dedicated to my Father, (John) Maurice Greenwood (1929-2000), who is still sadly missed, to my Mother, whose health recently deteriorated, and to my friend, Paul Wheeler (1954-2008), whose life was tragically cut short.

21

Chapter 1: Introduction, Literature Review, Aims & Objectives, Research Strategy & Thesis Layout

CHAPTER 1: INTRODUCTION, LITERATURE REVIEW, AIMS & OBJECTIVES, RESEARCH STRATEGY & THESIS STRUCTURE

1.1 Introduction 1.1.1 The Importance of Sediment Studies The mobilisation, transfer and deposition of sediment by rainfall and surface runoff are natural processes, but are often accelerated by human activities such as deforestation or inappropriate land use (Barrow, 2000; Zapata, 2003). Recognising soil as a non-renewable resource has led to concerns that accelerated erosion and the concomitant degradation of agricultural land pose significant threats to sustainable food production, to longterm global food security (Walling & Quine, 1992; Zapata, 2003; Walling, 2008) and to the quality of the environment per se (El-Swaify & Flach, 1988; Zapata et al., 2002). In quantitative terms, estimates from the early 1980s suggest that 3 million ha. of land were being ‘lost’ each year by erosion from the global agricultural resource-base (Buringh, 1981), and that the annual rate of soil depletion from global croplands were approximately 23 billion tonnes in excess of that which could be formed under natural conditions (Brown, 1984). More recent estimates have suggested that the productivity of global agricultural croplands are still diminishing by a mean value of around 5% each year, principally by water-driven soil erosion, and exacerbated by inappropriate agricultural practices (Crosson, 1996 in Oldeman, 1998). If this trend continues unabated, it is suggested that between 1.4-2.8% of the global agricultural and forestry resource-base will be irredeemably lost to future generations by the year 2020 (Scherr & Yadav, 1996). Aside from the on-site degradation of both soil structure and fertility associated with soil erosion, it is now widely recognised that eroded soil can give rise to many offsite environmental problems. Fine-sediment acts as a vector for the transfer and conveyance of nutrients and pollutants that contribute to the degradation of aquatic habitats (Walling, 2004; Walling et al., 2003; Walling & Collins, 2008). High concentrations of fine-sediment can physically disrupt the internal functioning of aquatic ecosystems and also act as a diffuse source of contamination in both aquatic and terrestrial ecosystems (Walling, 1983; Bottrill et al., 1999; Walling et al., 2003, 2006; Haygarth, 2005; Wood et al., 2005; Walling & Collins, 2008). In response to this realisation, depositional environments such as river floodplains have recently attracted increasing attention in their role as stores or sources of fine-sediment and sediment22


associated pollutants (Leeners & Schouten, 1989; Walling & He, 1998), or as transitional zones which link the catchment to the river channel (Burt & Haycock, 2000). The implementation of legislation such as the EU Water Framework Directive (2000/60/EC) has sought to promote an holistic approach to river basin management by reducing pollution, preserving and enhancing the biodiversity and amenity status of waterways, and sustaining their future well-being for the benefit of all stakeholders (Europa, 2008). Stimulated by this legislation, the formulation of catchment-scale management plans has, in many instances, highlighted an urgent need for effective sediment control strategies to manage and mitigate the deleterious effects of finesediment in the wider environment. The effectiveness of such measures to reduce sediment inputs and improve water quality ultimately relies on the ability to predict and quantify the spatial and temporal flux of sediment through catchments. Prognoses relating to the catchment sediment flux have typically relied on monitoring techniques which have focused on quantifying sediment yields at catchment outlets (Walling, 2006). Although this approach has provided an initial framework around which overall sediment budgets have been constructed, recognition of the wider implications of finesediment in the environment has directed attention to the internal functioning of catchments (Walling & Collins, 2008) in an attempt to gain an improved understanding of the conveyance pathways which link upstream sediment mobilisation by erosion to the downstream sediment yield, and to determine their relative efficiency in conveying sediment through the catchment (Walling, 2004). Identifying and elucidating those links however, remains one of the most complex and least understood components of a catchment sediment system. Importantly, it has also highlighted limitations in the ability of existing monitoring techniques to provide the detailed information needed to construct more meaningful sediment budgets that fully identify and integrate sediment sources and sinks, and the conveyance pathways interposed between the two components (Walling, 1999, 2004, 2006; Walling & Collins, 2008). These limitations have consequently prompted a shift from monitoring the movement of sediment towards tracing its movement, with the latter approach viewed as a complement to the former (Walling, 2004).

1.1.2 Sediment Tracers A tracer can be defined as any physical object or chemical substance capable of being introduced into a physical, chemical or biological system in order that it may be 23


subsequently tracked over a given timeframe and provide information which is inferred by its redistribution or dispersal (Sauzay, 1973; Evans, 1983). The suitability of any tracer depends on its ability to meet certain requirements, the most fundamental being its ease of identification, in order to replicate the behavior of the media in which it has been placed, and also to remain distinguishable from its surroundings (Sauzay, 1973; Evans, 1983; Guiresse & Revel, 1995; McCubbin & Leonard, 1995; Foster & Lees, 2000). A wide variety of materials and substances meet those requirements and the choice of tracer will ultimately depend on the calibre of the sediment (i.e. fine, medium or coarse) under investigation (Sear et al., 2000). Since most sediment-related environmental problems relate to fine-sediment and the package of erosion investigations described and conducted in subsequent chapters are primarily concerned with the movement and redistribution of fine-sediment, this review of tracer techniques accordingly focuses on types of tracers used to study the movement of material of that size. For clarity, tracers that are included in the review are divided into two groups; passive and active. The interpretation of passive and active tracers typically depends on the characteristics of the tracer, and of the frequency and stage during an investigation at which measurements of the movement or dispersal of the tracer can be undertaken (i.e. during or after an event) (Allan et al., 2006). Consequently, the definition of passive and active tracers frequently differs according to the particular research discipline making the distinction (cf. Sauzay, 1973; Evans, 1983; Sear et al., 2000; Allan et al., 2006; Granger et al., 2007). In light of this uncertainty, the following text provides a working definition of passive and active tracers in the context of this study. It is emphasised that these definitions relate specifically to this study and concern only those tracers used to determine the movement of fine-sediment. In addition, there are three factors which are explicit to this study and which require further clarification. Firstly, it is recognised that some tracer techniques discussed in the review are not necessarily used exclusively to study the movement of fine-sediment. Secondly, finesediment is defined as material of ca. < 2 mm dia., although it is also acknowledged that material of this size-calibre will often be eroded in the form of aggregates, which may be considerably larger than the maximum threshold size stated above. Finally, the review principally focuses on artificial tracers, which, according to Evans (1983), are defined as any material that has been deliberately introduced into a system for the benefit of following its subsequent movement. Some discussion of the fallout radionuclides, 137Cs, 210Pb and 7Be is, however, included in the context of their ability to 24


provide information on rates of soil redistribution over a wide range of timescales (Walling et al., 2009). Given the justification for establishing a firm distinction between passive and active tracers, as well as outlining the predefined constraints associated with this investigation, the following text now provides a working definition of both types of tracer.

1.1.3 Passive & Active Tracers: A Working Definition Passive tracers are defined as those that have no additional source of energy external to the force of movement and whose redetection cannot therefore be undertaken via an internal function of the tracer itself (Evans, 1983). Passive tracers are therefore inert, but can take the form of any individual objects or particulate material that has been placed on or in a host soil. Ideally, their size distribution should be similar to the host soil and their behaviour should remain as compatible as possible with its movement in response to an external force (Granger et al., 2007). Importantly, the inert nature of passive tracers means that the method of recovery and measurement typically involves their relocation by physical-based approaches. For small particulate material, this may involve systematically removing samples of sediment at different locations along the line of erosion and then recovering the tracer material from the sediment. Rates of erosion can then be determined based on changes in the concentration of the tracer relative to the concentration at its original location prior to the erosion event. For larger passive tracers, their relocation may involve more elementary methods and involve field-walking to observe and then record the position of individual objects relative to their original location. By necessity however, all methods of recovery are performed after one or a series of recursive erosion events and since, by necessity, sampling methods are usually invasive, through accessing and trampling the site and by disturbing the host sediment during their recovery, this tends to restrict the use of passive tracers to measuring erosion over single events (e.g. Lobb et al., 1999; Li et al., 2007; Tiessen et al., 2007b), or over events that occur in quick succession (e.g. Lobb, 2006). Active tracers have been defined as any material or substance whose redetection and measurement is facilitated by an internal function of the tracer itself (Evans, 1983), and such functions often have the ability to operate both independently and irrespective of environmental conditions (Lang, 2008). This attribute frequently permits modes of identification and measurement to be conducted using a variety of different techniques, many of which are non-invasive and / or non-destructive and can also be undertaken in25


situ and / or remotely (Evans, 1983). Many of those attributes serve to extend the operational life of an investigation by permitting measurements to be undertaken at different stages during the study-period, which allows the movement or dispersal of tracers to be monitored iteratively over a series of events. In most instances therefore, the characteristics evidenced by active tracers are usually viewed as positive attributes that can afford a major advantage over passive tracers by allowing repeat measurements to be performed over a sequence of erosion events occurring in quick succession or over slightly longer (i.e. seasonal) timescales (Wooldridge, 1965; Fullen, 1982). Given the two extended definitions of the two groups and a brief explanation of key differences between the two, the review will now focus firstly on passive tracers, and then on active tracers.

1.2 Literature Review 1.2.1 Passive Tracers: A Review One of the most common types of passive tracer used in investigations to determine the movement of fine-sediment are physical point-tracers. Physical pointtracers come in many different forms, span a wide range of sizes and densities and have been used in a range of erosion scenarios. Point-tracers should ideally be matched to the texture of the soil under investigation, but choices are often influenced by the availability of materials (Guiresse & Revel, 1995). Among the different types of pointtracers previously used in erosion investigations include dyed gravel, coloured stones, metal cubes and plastic beads and their attractiveness in erosion investigations reflects their low cost, inert nature, wide availability and relative ease of use. Given the exceptionally large number of studies where point-tracers have been used, some of the more common types have been compiled from the published literature. These are listed in Table 1.1 and are ranked in ascending order according to the size of the tracer. A fundamental approach when using point-tracers is the need to match the physical characteristics of the tracer to the approximate aggregate-size and bulk density of the host soil. Although numerous authors have stressed this prerequisite (e.g. Evans, 1983; Parsons et al., 1993; Rahman et al., 2005; De Alba et al., 2006), it is not always addressed, as demonstrated by Table 1.1. Although the list presented in Table 1.1 is neither exhaustive nor fully comprehensive, mean size and bulk density values were noted for both the tracer and the host soil from the investigations, and this indicated that 15 investigations (~ 83%) listed the bulk density of the host soil. In contrast, only 11 (~ 61%) listed the bulk density of the tracer, and only 10 (~ 56%) provided both pieces of 26


information. Surprisingly, just 2 (~ 11%) of the investigations (i.e. Govers et al., 1994 & Rahman et al., 2005(a)) acknowledged the need for, and showed evidence of, matching the bulk density of the tracer to the host soil. It has thus been argued that a failure to comply with this fundamental requirement potentially limits the ability of point-source tracers to provide representative data in many erosion scenarios, and particularly those where the energy-input is low, such as during inter-rill erosion for instance (Parsons et al., 1993). In spite of this limitation however, the varying attributes listed above make physical tracers an attractive and common choice, particularly in studies where estimates of soil translocation distance or redistribution patterns represent the primary objectives of the investigation, as evidenced by similarities in the nature of those studies listed in Table 1.1.

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Table 1.1 A summary of selected investigations from the available literature published between 1994 to 2007, showing a comparison between the physical characteristics of the point-tracers, which have been ranked in descending order of size, and the host soil. Data Source

Year

Nature of Investigation

Tracer Material

Maximum tracer dia. / axis (mm)

Aggregate Class (Wentworth Scale)

Mohler et al .

2006

Movement of weed seeds during tillage

Coloured ceramic beads

1

coarse / very coarse sand

no data

no data

Colbach et al.

2000


Plastic cubes

1


no data

no data

Roger-Estrade et al.

2001


Plastic cubes

1


no data

1360

Guiresse & Revel (1995)

1995

Soil redistribution during tillage

Crushed gravel

6

fine gravel

1330

1630

Spokas et al. (2007)

2007

Soil redistribution during tillage

Plastic beads

6

fine gravel

1090

1400

Rahman et al . (2005(a))

2005

Soil translocation during liquid manure injection

Plastic cubes

10

medium gravel

1300

1200

Rahman et al . (2005(b))

2005


Wood cubes

10

medium gravel

900

1200

Liu (unpublished PhD. Thesis) (2005)

2005

Soil translocation during mechanised tillage erosion

Aluminium cubes

10

medium gravel

no data

1330

Tiessen et al . (2007a)

2007a


Dyed aquarium gravel

13

medium gravel

2700

1185

Tiessen et al . (2007b)

2007b


Dyed aquarium gravel

13

medium gravel

2700

1185

Sharifat (unpublished PhD. Thesis) (1999)

1999


Plastic cubes

15

medium gravel

1750

1100

Govers et al . (1994)

1994


Plastic spheres

15

medium gravel

1750

1500

Marques da Silva et al.

2004


Aluminium cubes

15

medium gravel

2665

no data

Heckrath et al.

2006


Aluminium cubes

15

medium gravel

no data

1495

Rahman et al . (2005(c))

2005


Aluminium cubes

15

medium gravel

2700

1200

Van Muysen et al . (1999)

1999


Aluminium cubes

15

medium gravel

no data

1650

Rahman et al . (2005(d))

2005


Steel cubes

20

medium gravel

7700

1200

Li et al .

2007


Dyed gravel

25

coarse gravel

no data

1170

28

Tracer Bulk Density Soil Bulk Density (kg m3) (kg m3)


1.2.2 Active Tracers: A Review Other tracing approaches have attempted to overcome limitations on the representativeness of larger-sized physical tracers by using smaller particulate material, such as magnetic powder (i.e. magnetite) sorbed onto small polystyrene beads (e.g. Ventura et al., 2001). These were then incorporated into a host soil at known concentrations and subjected to rainfall in order to initiate surface runoff. A comparison of changes in the magnetic susceptibility measured before and after the rainfall event was used to infer the extent of soil redistribution. Although, in this instance, care was taken to match the bulk density of the tracer to the host soil, the physical presence of the beads across the soil surface reportedly provided an amouring effect. This inadvertently afforded the soil aggregates protection against impacting raindrops and consequently resulted in the enrichment of the tracer in inter-rill areas. Although a later study had benefited from certain refinements to the technique (Ventura et al., 2002) and served to confirm its overall viability, problems of preferential movement and enrichment of the tracer were still prevalent. These fluctuations corresponded to the intensity of the rainfall and resulted in a greater accuracy with higher sediment loads. This led the authors to conclude that more work was needed to determine the optimum grain-size and density range of the magnetic material and match those with varying rainfall and runoff conditions (Ventura et al., 2002). In a similar tracing approach, Plante et al. (1999) labelled small ceramic prills with dysprosium oxide (Dy2O3). After subjecting them to rainfall and recovering them in the eroded material they were then subjected to neutron activation analysis (NAA) in order to ‘activate’ the dysprosium oxide and thus make it radioactive, which allowed rates of soil redistribution to be determined (Plante et al., 1999). Riebe (1995) used a more complex labelling process whereby Europium (Eu2O3) was incorporated into molten glass, which was then cooled and ground to different sizes to represent soil aggregates. Lead was also introduced into the molten mixture, which enabled the bulk density of the tracer to be adjusted to the host soil. The mixture was then subjected to NAA, transforming the Europium into the gamma-emitting radionuclide, Eu-125m (125mEu) and enabling the labelled aggregates to be used in a variety of erosion investigations. This approach incurred significant time, effort and expenditure and the size-range of the chosen material prevented it from adequately mimicking the behaviour of the fine-sediment under investigations (Riebe, 1995). Parsons et al. (1993) used crushed magnetite as a tracer and rates of soil redistribution from inter-rill areas were inferred using magnetic susceptibility measurements. Although attempts were taken to 29


match the characteristics of the tracer to the host soil, again, differences in bulk density values resulted in poor agreement between actual erosion rates and those indicated by the tracer. Rare Earth Elements (REEs) have proved particularly useful for studying erosion over relatively short timescales, particularly in areas where high erosion rates reduce the effectiveness of other techniques (e.g. Junliang et al., 1994; Zhang et al., 2001, 2003; Polyakov & Nearing, 2004; Kimoto et al., 2006). Numerous characteristics have contributed to their popularity and these include their availability, their conservative behaviour, their ability to bind to clay particles and also to coalesce and mimic the behaviour of aggregates. Low background concentrations in most environments have also permitted high analytical sensitivity (Zhang et al., 2001). Other positive attributes include their inert nature, benign characteristics and the ability to detect different REEs simultaneously by inductively coupled plasma mass spectrometry (ICP-MS). Disadvantages with this type of material include the need for detailed sampling campaigns which cover the full extent of the area under investigation and thus provide a sufficient level of spatial resolution in order to determine representative estimates over the study area. Rare earth elements also require lengthy and complex laboratory-based reextraction procedures to recover them from samples of eroded soil. Moreover, the need to analyse samples using laboratory-based techniques prevents rates of soil redistribution from being determined by in-situ measuring techniques. In many instances, this latter factor has restricted their use to investigations conducted on erosion plots using different labelling configurations (e.g. Junliang et al., 1994; Zhang et al., 2003; Liu et al., 2004; Polyakov & Nearing, 2004; Li et al., 2006). Although this approach has successfully provided information on the spatial and temporal variability of soil redistribution, differences in the magnitude of redistribution have been recorded between different studies, depending on the proportion of tracer coverage within the plot and / or the proximity of each REE species to the plot-outlet (Zhang et al., 2003; Liu et al., 2004). It has been demonstrated that the level of spatial information derived from this approach is dependent on the number of different REEs used in an investigation (Liu et al., 2004). Evidence of preferential sorting has also been reported from investigations using REEs and this has been attributed to their extremely fine texture and poor ability to bind to soil aggregates under certain conditions (Zhang et al., 2003; Polyakov & Nearing, 2004). Some workers have also used REEs and then subjected the samples to NAA. 30


Although this permits subsequent analyses to be conducted using conventional radiometric techniques such as gamma spectrometry, this potential saving is often negated owing to the additional time and expenditure incurred by using NAA (Showler et al., 1988; Junliang et al., 1994; Ventura et al., 2001; Li et al., 2006). Radioactive tracers have been widely used in sediment studies (e.g. Ritchie & McHenry, 1990; Walling, 2004) and these techniques have provided researchers with a deeper understanding of the dynamics of sediment-systems, owing to the process of nuclear decay, which is independent of environmental conditions (Lang, 2008). This factor has allowed a temporal perspective to be introduced for many sedimentological processes (Campbell et al., 1988) and timeframes can now be attributed to many environmental processes once considered too slow and immeasurable, or simply too innocuous to merit closer attention (e.g. Owens et al., 1999; Walling et al., 1998, 2003; Wallbrink et al., 2002; Walling & Owens, 2002; Walling & Collins, 2008). As previously stated, fallout radionuclides technically lay outside of the predefined remit of this review. It is pertinent at this particular point, however, to briefly mention techniques using the anthropogenic fallout radionuclide,

137

Cs, the naturally-

occurring fallout radionuclide, 210Pb, as well as the short-lived cosmogenic radionuclide beryllium-7 (7Be) (t0.5 = 53 d.), owing to their prevalence and previous success in sediment tracing investigations. The first two radionuclides have permitted information on rates of sediment deposition and redistribution to be obtained retrospectively at wide range of spatial scales and over timescales of ca. 40 to 100 years respectively (e.g. Walling & He, 1994; He & Walling, 1996; He et al., 2002; Zapata et al., 2002). Although this is often viewed as a major advantage, especially in the context of establishing longer-term sediment storage information, or to determine catchment responses to longer-term land-use changes (Walling & He, 1994), such estimates of rates of sediment mobilisation and deposition represent averages for that time period. These techniques are therefore unable to account for soil redistribution over short timescales or over single events, or where redistribution has occurred in association with short-lived land use conditions (He et al., 2002). In situations such as those, recent developments using the short-lived fallout radionuclide 7Be have proved successful (e.g. Blake et al., 1999, 2002; Walling et al., 1999) and are now firmly established as a complementary technique to existing medium-term approaches using longer-lived fallout radionuclides (e.g. Wilson et al., 2003; Schuller et al., 2006; Sepulveda et al., 2008; Walling et al., 2009). However, 7Be can be readily sequestered by surface vegetation (Mabit et al., 2008), which, depending 31


on its type and density, can influence the amount of fallout inventory that reaches the soil surface. Since this can alter the depth distribution of 7Be within the soil profile whilst the investigation is in-progress (Walling et al., 2009), studies to determine estimates of soil redistribution using the 7Be technique have tended to be conducted exclusively on bare soils, to soils that have recently been cultivated, or across areas with sparse vegetation cover or low crop residue (e.g. Blake et al., 1999, 2002; Wilson et al., 2003; Doering et al., 2006; Schuller et al., 2006; Mabit et al., 2008; Sepulveda et al., 2008; Walling et al., 2009). In addition to the above limitation, the technique originally relied on prior knowledge of the pre-event spatial distribution of 7Be over the study area, which increases the scale of the field-collection campaign and the number of radiometric assays by a factor of two, and also introduced additional complexities into the subsequent data analysis (He et al., 2002; Walling et al., 2009). Attempts to overcome these problem have been addressed by making a prior assumption that the spatial fallout distribution of 7Be across the study area is uniform, by virtue of either a substantial preceding dry period in which much the existing inventory will have decayed away, or by recent cultivation, which incorporates the surface soil, and hence the residual 7 Be inventory, into the lower horizons (He et al., 2002). Deviations in inventory values from areas within the study area are then compared against a reference inventory determined from a nearby non-eroding site (e.g. Blake et al., 1999; Walling et al., 1999). A numeric conversion model, based on the depth distribution of the radionuclide in the soil profile as indicated by the reference core, is then used to convert the inventory measurements obtained from individual cores to determine estimates of soil redistribution (Walling & He, 1999). For the assumption of spatial homogeneity in inventory values to be to be correct however, sufficient time must have elapsed since the last erosion event in order to ensure that variations in the inventories inherited from the last event and induced by soil redistribution have been removed through natural radioactive decay (Blake et al., 1999, 2002; He et al., 2002). Since at either the reach or catchment scale, this assumption is often not met (Mabit et al., 2008), measuring and accounting for changing inventory values between erosion events has previously represented an obstacle to obtaining accurate soil redistribution data using 7Be over a sequence of erosion events, or where single events have occurred in relatively quick succession, unless inventories at predetermined locations are known prior to each event (Blake et al., 1999). These problems have been overcome by applying an inventory balance approach (Wilson et al., 2003), which measures the 7Be inventories before and after a storm event, and also monitors the 32


atmospheric flux deposited at the site between erosion events to calculate an inventory budget specific to that particular location. This approach has substantial advantages over the reference inventory method since all data are collected from the study site itself and estimates of erosion losses, calculated by determining the depth distribution of the radionuclide within the soil profile, are based on the actual depth distribution of the soil under investigation (Wilson et al., 2003). Certain disadvantages associated with the inventory balance approach have, however, prevented its wider acceptance within the research community. These include firstly, the need for repeat measurements, which dramatically increases the time needed to obtain samples and to undertake radiometric assays, and secondly, the timing of the sampling campaigns, which need to coincide immediately prior to a storm event, which thus requires the ability to precisely forecast the timing of the next event (Walling et al., 2009). Recent developments to the technique have now managed to resolve many of the latter problems outlined above (Walling et al., 2009), by accounting for variations, firstly, in the temporal distribution of 7Be fallout, secondly, in the temporal distribution of sediment redistribution, and thirdly, the influence of the precipitation history and depositional inputs to the depth distribution of the radionuclide within the soil profile of the study area during the period under investigation. Thus, the technique has now successfully been extended to provide accurate soil redistribution information from multiple storm events over periods of several months (Walling et al., 2009). The use of artificial radionuclides, or to clarify, radionuclides that have been deliberately introduced into or across a predefined study area or used to artificially ‘label’ small quantities of fine-sediment, have the potential to overcome many of the limitations associated with other tracers of fine-sediment described above. This is largely due to their limited distribution in the wider environment, the level of convenience afforded by their use in soil erosion investigations and their ability to faithfully mimic the behaviour of moving sediment during an erosion event. However, in spite of adequate demonstrations of their effectiveness over relatively short timescales or over single events, albeit over relatively small areas (e.g. Toth & Alderfer, 1960a, 1960b; Wooldridge, 1965; Fullen, 1982; Syversen et al., 2001), their use in studying the movement of fine-sediment has attracted surprisingly little attention over the last ca. fifty years. Reasons for this remain unclear, but may be attributed to factors such as the limited areal extent over which artificial radionuclides can be realistically introduced (Syversen et al., 2001). Numerous reasons have been suggested for their under-utilisation, but the most common relate to the logistical constraints and 33


impracticalities associated with attempting to uniformly apply the radionuclide, along with the need to undertake any re-measurement strategy at a sufficient spatial resolution and temporal frequency to provide an adequate level of data coverage commensurate to the size of the area under investigation. These constraints have typically confined the application of artificial radionuclides, as well as numerous other tracers that have been artificially-applied to soils, to small-scale investigations consisting of areas of just a few m2 (cf. Toth & Alderfer, 1960b; Wooldridge, 1965; Fullen, 1982; Junliang et al., 1994; Syversen et al., 2001; Zhang et al., 2003; Liu et al., 2004; Payne et al., 2004; Polyakov & Nearing, 2004; De Alba et al., 2006). Despite this apparent limitation however, some workers have managed to overcome scale-constraints by using point-tracing approaches (cf. Wooldridge, 1965; Fullen, 1982). Such techniques have proven extremely effective and convenient in providing soil redistribution data at representative locations within a given landscape over short timescales or over single or sequences of rainfall events (e.g. Wooldridge, 1960; Fullen, 1982; Zhang et al., 2003; Liu et al., 2004). Since only a few gamma-photons are needed to identify the species of radionuclide, arguably one of the biggest attributes associated with artificial radionuclides is the relative ease with which they can be detected, measured and readily distinguished from other environmental radionuclides, even in depositional environments where burial is likely. This level of sensitivity far exceeds the capability of many other tracers and provides a distinct advantage over other tracertypes (cf. Wooldridge, 1965; Alldredge & Whicker, 1972; Courtois, 1973; Pahlke, 1973; Showler et al., 1988; Tyler et al., 1996a, 1996b; Golosov et al., 2000). Owing to the negligible mass associated with gamma-photons emitted by radionuclides, soil labelled with radionuclide material theoretically behaves in a manner which is identical to unlabelled soil (Toth & Alderfer, 1960a; Wooldridge, 1965). Examining the spatial redistribution of sediment, as inferred by changes in radioactivity values measured after one or a series of erosion events, permits the dispersal, redistribution or burial of very small quantities of fine-sediment pre-labelled with radionuclide material to be accurately and readily determined over a range of spatial and temporal scales with a level of sensitivity that far exceeds the majority of other tracers (Toth & Alderfer, 1960b; Wang et al., 1975; Fullen, 1982; Alam et al., 2001; Zhang et al., 2001).

1.3 Aims & Objectives From the discussion outlined in earlier sections, it has been demonstrated that the use of artificial radionuclides as tracers of fine-sediment have the potential to offer a 34


significant number of advantages over many of the other types of tracer reviewed in previous sections. Many of the attributes associated with artificial radionuclides have the ability to extend the scope of an investigation into novel research areas. In the light of those varying attributes, along with the general success but limited scope of the pioneering efforts of the small number of investigations mentioned earlier that have previously used artificial radionuclides as tracers of fine-sediment, scope exists to explore further and confirm their potential through the development of a series of tracing applications linked to a variety of different erosion and sediment transport scenarios. The first aim of the thesis was to develop five separate tracing applications. These related to a range of different fine-sediment erosion and transport scenarios in which other types of tracer or tracing technique were likely to be unsuitable or ineffective, in order to demonstrate the efficacy of the radionuclide tracers. A second aim of the thesis was to use the information derived from each of the tracing investigations to provide an improved understanding of some of the mechanisms of sediment mobilisation and transfer, particularly those related to its conveyance from source to catchment outlet. A number of objectives were formulated in order to address these aims and these are as follows:

i.

To identify two radionuclides for use as sediment tracers in a series of sediment tracing investigations;

ii.

To design five novel tracing applications in order to evaluate the effectiveness of the radionuclides as sediment tracers over a range of applications over suitable spatial and temporal scales; and

iii.

To interpret the results from each of the investigations and assess whether the aims have been adequately achieved.

1.4 The Research Strategy The research strategy focused on designing five separate investigations, each consisting of a different tracing application and each representing a different erosion scenario to be examined through either field and / or laboratory-based experiments. A conceptual framework was formulated as part of the research strategy and is shown in Figure 1.1. This permitted three different research principles to be identified and a synthesis of those research principles represented a starting-point from which potential research themes were identified and the nature and suitability of each tracing application

35


was ultimately determined. Because of their role as drivers for the entire study, each of the three main research principles is discussed separately in the following text.

36


Figure 1.1 A conceptual framework was formulated based on several research principles and this was used to identify the desired research outcomes and potential research themes.

37

Chapter 1: Introduction, Literature Review, Aims & Objectives & Research Strategy

1.4.1 Added Value The first research principle was categorised as ‘Added Value’ and its aim was to ensure that each of the individual investigations within the overall study demonstrated an adequate degree of experimental ‘uniqueness’, or novelty. The required level of novelty was achieved by ensuring that a number of requirements were met. The first of these was to identify one or more candidate radionuclides with suitable environmental characteristics, but whose use as a tracer of fine-sediment in previous investigations was limited. Before a suitable candidate (or candidates) could be identified however, it was necessary to determine the cost of different radionuclides, as this ultimately dictated both whether their use was feasible and the quantity that could be purchased. Numerous internet-based enquiries were made to (legitimate) suppliers of radioactive products in order to determine the cost of different radionuclide(s) at activity concentrations sufficient to undertake all of the investigations. Based on the quotes obtained, it was deemed that the available research budget was sufficient to purchase two different radionuclides in suitable quantities. Since the use of multiple tracers was considered to be beneficial to the overall study, by potentially increasing the range and complexity of tracing applications, the ability to procure multiple radionuclides represented the second requirement. The use of multiple tracers consequently permitted a range of different ‘simultaneous’ and ‘parallel’ tracing strategies to be undertaken which would not otherwise be possible using one radionuclide. As an example, a simultaneous tracing approach could involve applying both radionuclides using a stratified or layered tracing approach during a single investigation to determine, for instance, the relative magnitude of soil erosion by inter-rill and rill erosion. In contrast, a parallel tracing approach would involve using both radionuclides to perform an identical tracing application. If performed under similar environmental constraints, this approach could allow the two sets of results to be compared, in order to determine the performance and behaviour of each of the tracers. Based on these possibilities, the second requirement listed under the research principle of Added Value, and previously alluded to earlier in this section, dictated that, ideally, one or both of the chosen radionuclides would be either previously untested as a fine-sediment tracer, or their previous use in such investigations would be relatively limited. This strategy would therefore permit the behaviour of the 37


radionuclides to be evaluated as potential tracers of fine-sediment and this would also contribute to achieving the first aim listed previously in Section 1.3. The third and final requirement involved the development of a suite of sufficiently complex research problems, such that the results from each investigation were only obtainable using one or more radionuclides which had been deliberately introduced into a given environment for the specific purpose of investigating the movement and redistribution of fine-sediment over a given time period. Complying with this requirement was thus considered to be of paramount importance in order to ensure that the full tracing capabilities of the radionuclides were comprehensively tested.

1.4.2 Convenience & Availability of Information The second research principle was categorised as ‘Convenience & Availability of Information’ and, as the title of this sub-section implies, was principally concerned with the convenience and availability of information afforded by the use of radionuclides in sediment tracing scenarios to providing reliable estimates of the movement of small quantities of fine-sediment across a range of spatial and temporal scales. A major advantage associated with artificial radionuclides over other types of tracers that have previously been used to estimate the movement of fine-sediment relates to the principle of minimal mass theory (e.g. Sauzay, 1973; Courtois, 1973; McCubbin & Leonard, 1995). The general principle associated with this theory has previously been described in an earlier section, but to briefly reiterate, it states that the behaviour of any tracer must be as compatible as possible with the medium being traced in order to derive accurate and representative information. Because of the negligible mass of gamma-photons, sediment labelled with a radionuclide material is theoretically capable of providing an absolute indication of the movement of the eroded sediment without unduly perturbing or adversely influencing its behaviour during erosion events (Wooldridge, 1965; Fullen, 1982; Showler et al., 1988; McCubbin & Leonard, 1995). It was therefore considered paramount to take full advantage of these attributes and aim to acquire essentially unique information from each of the investigations which would otherwise be unobtainable using other less sensitive types of tracer and tracing techniques. This attribute has proven to be particularly beneficial during attempts to study, for instance, the movement of 38


sediment in low-energy environments where material is being mobilised by inter-rill erosion from the upper surface of the soil (Syversen et al., 2001). Although other techniques using fallout 7Be have proven to be equally capable of providing similar information over a much larger range of spatial scales (e.g. Blake et al., 1999, 2002; Wilson et al., 2003), and latterly, over an extended range of temporal scales (e.g. Walling et al., 2009), radionuclides which have been deliberately placed onto or within a given area have been shown to provide a level of convenience which exceeds that of many other techniques and which permits parallel tracing investigations to be undertaken and replicated at representative locations within the landscape (Wooldridge, 1965; Fullen, 1982). Such strategies have proven extremely useful when attempting to draw comparisons between rates and patterns of sediment redistribution from areas sharing similar topographic features, to provide extremely detailed information on particle transport pathways, and for determining rates, patterns and directions of soil movement between sets of environmental conditions (Wooldridge, 1965; Fullen, 1982, Syversen et al., 2001). In addition, a further convenience which is often associated with the use of artificial radionuclides in sediment tracing investigations arises from the fact that the operator can, within reason, determine the initial radionuclide concentration or inventory (Wooldridge, 1965; Syversen et al., 2001). Since the initial inventory or activity concentration values are not reliant on environmental levels of fallout radioactivity, higher initial applications can readily extend the duration of any experiment to incorporate seasonal or even multiple-seasonal variations, and therefore permit spatial and temporal variations in erosion and deposition to be precisely and accurately documented (Wooldridge, 1965; Syversen et al., 2001).

1.4.3 Scale The last research principle was categorised as ‘Scale’ and concerned three main factors. The first of those relates to the logistical and practical constraints associated with uniformly labelling soil with radionuclide material over large areas. This limitation has been noted by many other researchers (e.g. Wooldridge, 1965; Fullen, 1982; Quine et al., 1999; Syversen et al., 2001) and in the absence of satisfactory methods which permit material containing radionuclides to be uniformly applied directly to soils (Syversen et al., 2001), meant that the majority of erosion investigations would need to be undertaken within relatively small areas and would 39


also require pre-event knowledge of the relative radionuclide inventories measured at predetermined locations. Moreover, in order to derive meaningful information, subsequent re-measurements would need to be repeated precisely at those predetermined locations in order to accurately determine changes in the radionuclide values and to ensure that changes in the inventory values before and after single events or over given time-periods could be assessed (Wooldridge, 1965; Fullen, 1982; Syversen et al., 2001). These factors potentially limited the area of individual labelled plots to a few m2. However, a way in which the areal extent of the study area was extended included the use of small point-scale plots dispersed across particular landforms, or located on representative landform features (Wooldridge, 1965; Fullen 1982). As discussed in the previous sub-section, the convenience offered by applying radionuclide tracers to small point-scale plots at representative locations across a predefined study area for the benefit of determining the movement and redistribution of sediment has overcome many of the perceived spatial constraints associated with their use. It has also permitted the collection of high-resolution data, enabling rates and patterns of sediment redistribution to be determined, and also permitted comparisons to be drawn between data-sets obtained from comparable or similar landforms (Fullen, 1982). The second factor relates to the likelihood that the subtleties associated with monitoring and tracing the mobilisation and transfer of small quantities of labelled sediment during small-scale experiments may provide more detailed and precise information regarding the sediment conveyance mechanisms and also permits the relative efficiency of each mechanism to be determined during the sediment conveyance process. The third and final factor relates to studies by numerous authors (cf. Walling & Quine, 1990, 1992; Dalgleish & Foster, 1996; Walling & He, 1999; Walling et al., 2002), all of whom have identified uncertainties regarding the behaviour of fallout radionuclides in certain erosion scenarios and have thus called for research to improve the current understanding of certain factors. These include a refined determination of their vertical distribution within the soil profile in undisturbed environments, an improved understanding of the relationship between radionuclide concentration and particle-size, and a deeper understanding of the effects of grain-size selectivity during erosion events and its effect on the spatial redistribution of fallout radionuclides. Owing to similarities between the environmental characteristics of the artificial 40


radionuclides used to conduct the series of small-scale experiments in this study and the suite of established tracing techniques using fallout radionuclides to determine soil redistribution over a range of short to medium-term timescales, information obtained from one or all of these investigations may result in a deeper understanding of the behaviour of environmental fallout radionuclides. It may also ultimately lead to an improved understanding of their capabilities as sediment tracers. The remaining section of this chapter now outlines the structure of the thesis and also provides brief details of each of the subsequent chapters.

1.5 Thesis Structure The remaining components of the thesis comprise seven chapters and a broad outline of the structure of each is as follows: Chapter 2 firstly identifies the two radionuclides chosen for the investigations. It then describes how they were selected and then discusses their respective origins and environmental characteristics. It then outlines some of the basic health and safety issues associated with the exposure to radionuclide material and the Principles of Gamma Spectrometry. Chapters 3 to 7 describe five different erosion investigations using a generic format, which consists of a very brief overview, followed by an introduction summarising the need for the investigation and placing each in the context of previous work. A series of aim(s) and objectives specific to each investigation are then listed followed by a detailed description of the method. The results are then presented and interpreted, the tracing application is assessed and conclusions relating to the results of each application are drawn. A summary of each of the five investigations is outlined below: Chapter 3 describes an investigation to determine the extent of sediment remobilisation during individual overbank flood events on a river floodplain within two separate catchments in Devon, UK. Chapter 4 describes an investigation to determine soil translocation distances and patterns of soil movement during upslope chisel plough tillage using unique experimental equipment located at the University of Manitoba, Canada. Chapter 5 describes an investigation to determine the extent of sediment redistribution on livestock-poached pasture on a hillslope located in Devon, UK.

41


Chapter 6 describes an investigation to determine the fate of earthworm casts during erosion on pasture and cultivated soil using both field and laboratory-based approaches. Chapter 7 describes an investigation to determine changes in the grain size composition of eroded sediment during the transition from inter-rill to rill erosion on a sample of eight different agricultural soils collected from locations in the south of England, using a laboratory-based approach. Chapter 8 then summarises the conclusions from each of the investigations and evaluates whether the aims of the overall investigation have been achieved. General conclusions are then drawn to assess the potential for the future use of the two radionuclides in tracing investigations in order to determine the movement and redistribution of fine-sediment.

42

Chapter 2: Radionuclide Selection Process, Origin & Characteristics, Health & Safety Issues & Gamma Spectrometry

CHAPTER 2: THE RADIONUCLIDE SELECTION PROCESS, ORIGINS & CHARACTERISTICS, HEALTH & SAFETY & THE PRINCIPLES OF GAMMA SPECTROMETRY

This chapter describes the radionuclide section process and then outlines the origins and characteristics of the two that were selected to undertake the sediment tracing investigations. Basic health and safety issues and best practice requirements concerning the use of and exposure to gamma-emitting radionuclides are then discussed, followed by a description of the principles of gamma spectrometry and the laboratory and field equipment used to measure radioactivity in soils and sediment.

2.1 Radionuclide Selection Process In order to be effective as sediment tracers, it was vital that the radionuclides selected should possess several characteristics. A simple checklist, shown in Table 2.1, was developed which lists seven fundamental characteristics and this was used to facilitate the selection process. Each characteristic was attributed equal status and for the radionuclide to be considered as a potential candidate, each characteristic needed to be present. Characteristics ranged, for instance, from determining whether they decayed by gamma radiation (for ease of rapid analysis), through their half-life duration, to their availability (i.e. whether they could be readily purchased). Each characteristic is represented as a sub-heading within the checklist and posed as a direct question in order to invoke a Yes (= 1) or No (= 0) response and thus identify whether each radionuclide possessed that particular characteristic.

43

Chapter 2: Radionuclide Selection Process & Environmental Characteristics

Table 2.1 A checklist was developed based on a suite of criteria and this was used to identify two candidate radionuclides as sediment tracers for use in a package of soil erosion investigations. CRITERIA CHECKLIST Gamm aEmitter?

Environmentally Conservative?

Suitable Half-Life Length?

Suitable Radiometric Yield?

Low BioUptake?

Used in Previous Tracing Investigations?

Total

Comm ents

Purchasable?

Am-241

0

1

0

0

1

1

0

3

Alpha-emitter requiring complex laboratory procedure to determine activity concentrations by gamma spectrometry; radiometric yield too low

Ba-133

1

0

1

0

1

1

1

5

Radionuclide

W ater soluble and hence, mobile in environment; radiometric yield too low Abundance in environment due to precipitation-sensitive fallout deposition would complicate any subsequent measurement strategy; half-life too short; not available for purchase Environmental behaviour unverified; tends to become metabilised within bone material; radiometric yield too low

Be-7

1

1

0

1

1

0

1

5

Ca-45

0

0

1

0

0

1

1

3

Cd-109

1

0

1

0

1

1

0

4

Environmental behaviour unverified

Co-60

1

1

1

1

1

1

1

7

Satisfies all check-list requirements

Cr-51

1

0

0

0

0

1

1

3

Environmental behaviour unverified; very easily absorbed through skin and has tendency to bio-accumulate in tissue; radiometric yield too low

Cs-134

1

1

1

1

1

1

1

7

Satisfies all check-list requirements

Cs-137

1

1

0

1

1

1

0

5

Abundance in environment due to fallout deposition would complicate any subsequent measurement strategy; half-life too long

Eu-152

1

0

0

0

1

1

1

4

Environmental behaviour unverified; half-life too long; not available for purchase; radiom etric yield too low

Eu-154

1

0

1

0

1

0

1

4

Environmental behaviour unverified; not available for purchase; radiometric yield too low

Fe-59

1

1

0

0

1

1

0

4

Half Life too short; radiom etric yield too low

Na-22

1

0

1

1

0

1

1

5

Environmental behaviour unverified; potentially high bio-uptake Beta-emitter; does not sorb to sedim ent; potentially bio-accum ulates in tissue; not available for purchase; radiometric yield too low Beta-em itter; gradual desorption from sedim ent; half-life too long, bio-accum ulates in bones

Sb-125

0

0

1

0

0

0

1

2

Sr-90

0

0

0

0

0

1

1

2

Y-88

1

0

0

1

1

1

0

4

Zn-65

1

0

1

1

0

1

1

5

Can become metabilised in bones and in bone-marrow

2

W ater-insoluble and so requires com plex pre-treatment with strong acid; half-life too short; radiom etric yield too low; not available for purchase

Zr-95

1

1

0

0

0

0

0

44

Environmental behaviour unverified; half life too short

Chapter 2: Radionuclide Selection Process, Origin & Environmental Characteristics & the Principles of Gamma Spectrometry

Eighteen radionuclides were selected as potential candidates and each was systematically assessed under the suite of characteristics and accorded an appropriate score using information obtained from some of the available literature (cf. Toth & Alderfer, 1960a, 1960b; Wooldridge, 1965; Fullen, 1982; Showler et al., 1988; Rowan, 1990; Rowan & Walling, 1992; Higgitt et al., 1993; McCubbin & Leonard, 1995; Zehnder et al., 1995; Avery, 1996; Bailly du Bois, 1996; Rühm et al., 1996; Smith et al., 1997; Bailly du Bois & Guéguéniat, 1999; Laissaoui & Abril, 1999; Tyson et al., 1999; Børretzen & Salbu, 2000; Suárez et al., 2000; Lu & Mason, 2001; Syversen et al., 2001; Massas et al., 2002; Real et al., 2002; Ristic et al., 2002; Galabov et al., 2003; Khan, 2003; Caron & Mankarios, 2004; Payne et al., 2004; Twining et al., 2004; Liu et al., 2005; Shinonaga et al., 2005; Shi & Guo, 2006; Wicker & Ibrahim, 2006; Gudelis et al., 2008). Individual values scored by each radionuclide were then summed and those with the highest totals were preliminarily identified as being potentially suitable for use in the package of sediment tracing investigations which were conducted and are described over the following chapters. Using this approach, highest values were scored by the radionuclides cobalt-60 (60Co) (7 points) and caesium-134 (134Cs) (7 points). Prior to purchasing suitable quantities of each species of radionuclide from a reputable supplier (Eckert & Zeigler Isotope Products, Berlin, Germany), closer attention was given to their environmental characteristics.

The following sub-section provides details of the origin and environmental characteristics of

60

Co and

134

Cs and then also briefly reviews previous work where

these two radionuclides have been used as sediment tracers.

2.2 Cobalt-60 & Caesium-134 For clarity and convenience, the characteristics and origin of each radionuclide are discussed separately. 2.2.1 60Co: Origin & Environmental Characteristics Cobalt-60 (60Co) is an anthropogenic gamma-emitting radionuclide with a relatively short half-life (t0.5) of ~ 5.26 years (Khan, 2003; Gudelis et al., 2008). It is classified as an activation product that is inadvertently manufactured in nuclear reactors by neutron activation of the naturally-occurring isotope, Cobalt-59 (59Co). Cobalt-59 is a stable metal element that is relatively ubiquitous in the environment and occurs in 45


various minerals and metal-ores and typically remains as an impurity within most finished metal products. Metals containing 59Co are frequently (and inadvertently) used in the construction of nuclear reactors. Upon exposure to the intense radiation-field within the reactor,

59

Co absorbs a free neutron into its nucleus, transforming into the

unstable radioisotope

60

Co (Khan, 2003). Once transformed,

60

Co begins to emit

ionising radiation in the form of beta particles (β-) and gamma radiation (γ-radiation). After one half-life (~ 5.26 yrs.), radioactive decay results in the transformation of 50% of the original 60Co nuclei into its stable progeny, nickel-60 (60Ni) (EPA, 2002). Cobalt-60 emits γ-radiation at photo-peaks of ~1,173 and ~1,332 Kilo electron Volts (KeV). The energy yield, which is defined as the proportion of (60Co) atoms produced per 100 disintegrations, is ca. 100% at each photo-peak. It is theoretically possible to obtain accurate results from samples with activities as low as 1.0 Bq kg-1 of (dry) sediment, assuming long (~ 1 day (d).) counting times (Grapes, Pers. Comm.). With regard to its presence in the environment, quantities of

60

Co have been

discharged as effluent under license into marine environments from nuclear reprocessing plants such as La Hague in northern France and Sellafield in Cumbria (Bailly du Bois, 1996; Cundy & Croudace, 1996; Matishov et al., 1999). Discharges of 60

Co can be detected in close proximity to its release-point in samples of seawater

(Bailly du Bois, 1996; Bailly du Bois & Guéguéniat, 1999) or in surface-sediments or from sediment cores (Cundy & Croudace, 1996; Matishov et al., 1999). Some workers have exploited these licensed releases and used them as a tracer or geochemical marker to elucidate varying environmental processes (e.g. Cundy & Croudace, 1996; Bailly du Bois, 1996; Bailly du Bois & Guéguéniat, 1999; Thompson et al., 2002). As well as the licensed releases noted above, areas in close proximity to disused underwater nuclear test-sites also represent an additional yet highly localised source of 60Co (Matishov et al., 1999). Owing to its anthropogenic origin, relatively short half-life and limited discharge sources,

60

Co does not occur naturally in the wider environment (Bailly du Bois &

Guéguéniat, 1999; Khan, 2003). This factor represents a potential advantage during sediment tracing investigations for numerous reasons, principally, due to the lack of background contamination, which removes the possibility of obtaining spurious data due to spatially variable background levels (cf. Sutherland, 1991, 1994, 1996), and this also permits tracer inputs and outputs to be precisely quantified (Sauzay, 1973). It also facilitates the likelihood of successfully documenting subtle changes in radionuclide inventories over a range of spatial and temporal scales and, in addition, negligible 46


background contamination also permits the use of lower initial dose-rates in tracing investigations (Quine et al., 1999). Importantly, all of the above factors potentially contribute to increasing the analytical sensitivity, accuracy and longevity of the monitoring campaign for a given initial inventory. A substantial amount of literature has been compiled detailing the characteristics, behaviour and mobility of 60Co under varying environmental scenarios (e.g. Matishov et al., 1999; Galabov et al., 2003; Khan, 2003; Caron & Mankarios, 2004; Shinonaga et al., 2005; Gudelis et al., 2008; Chen & Lu, 2008; Payne et al., 2004, 2009) and this review principally focuses on the same characteristics owing to their relevance to this investigation. These characteristics include the rate and strength of retention (sorption) to sediment, the solubility and vertical migration of

60

Co; and its

rate of biological-uptake (bio-uptake) and accumulation, and thus its potential toxicity to organisms. With respect to the first two parameters, the solid / liquid distribution coefficient (Kd) values have been used to describe the mobility, sorption and desorption of 60Co in response to a range of environmental and physical controls. These range from the particle size characteristics, organic matter (OM) content, pH, Eh, temperature, moisture and time (Cremers et al., 1988; EPA, 1999; Laissaoui & Abril, 1999; Sanchez et al., 2002; Khan, 2003; Payne et al., 2004, 2009). Kd values for

60

Co, obtained from wide

ranging soil-types, indicate that sorption is typically high to very high (Børretzen & Salbu, 2000; Galabov et al., 2003; Khan, 2003; Shinonaga et al., 2005). It is reported that pH is a major controlling factor which often dominates the rate at which sorption occurs, especially in acidic environments (i.e. ≤ pH 6) (Khan, 2003; Chen & Lu, 2008; Payne et al., 2009). However, the influence exerted by pH diminishes very rapidly from ca. pH neutral and upwards, and by ~ pH 8, sorption to sediment is reportedly ~ 100% (Khan, 2003; Sparks, 2003; Chen & Lu, 2008). Under environmental conditions typically encountered in agricultural environments, therefore, sorption of 60Co to soil is both rapid and considered to be largely irreversible (Shor & Dial, 2000; Khan, 2003; Shinonaga et al., 2005; Chen & Lu, 2008). The environmental characteristics of 60Co have resulted in it being used as a tracer in a wide variety of investigations over approximately the last six decades. To demonstrate this, some investigations from the available literature have been compiled and are listed in Table 2.2 by year of publication. Although the list is neither exhaustive nor comprehensive, it provides some indication of the diverse applications in which 60Co has been used and thus highlights its

versatility

as

a

tracer 47

of

environmental

processes.


Table 2.2 A selection of literature that demonstrates the wide variety of applications where 60Co has been used as a tracer for a range of environmental processes. Data Source

Year of Publication


Arnason et al. (in Showler et a l., 1988)

1950

Determine the movement of wireworms (Ctenicera Spp .) in soil

Rings & Layne (in Showler et al ., 1988)

1953

Determine the dispersal and habitat range of beetles (Curculionoidea Spp .) in woodland

Sullivan (in Showler et al ., 1988)

1953

Determine the dispersal and habitat range of beetles (Pissodes Spp .) in woodland

Babers et al. (in Showler et al. , 1988)

1954

Determine the dispersal and habitat range of beetles (Anthonomus Spp. ) in woodland

Green & Spinks (in Showler et al ., 1988)

1955

Determine movement of wireworms (Agroites Spp .) in soil

Green et al . (in Showler et al. , 1988)

1957

Determine movement of moth larvae (Rhyacionia Spp .) in soil

Green et al . (in Showler et al. , 1988)

1957

Determine the dispersal and habitat range of mature moths (Rhyacionia Spp .)

Toth & Alderfer

1960a

Establish a procedure for labelling water stable aggregates (WSAs) by immersion into radionuclide solution

Toth & Alderfer

1960b

Monitor the formation & breakdown of labelled WSAs

Traniello et al . (in Showler et al. , 1988)

1985

Determine territorial feeding patterns of termites (Isoptera Spp.)

Rosengaus et al . (in Showler et al. , 1988)

1986


Thorén et al. ,

1991

Determine location of cancerous brain tumours in humans

Cundy & Croudace

1996

Wastewater effluent used as geochemical marker to determine rates of estuarine sediment accretion and sea-level rise

Alam et al.

2001

Determine rates of nutrient uptake in agro-crops

Capowiez et al .

2001

Determine 3-dimensional trajectories of burrowing earthworms

Sattar et al .

2002


Thompson et al.

2002

Wastewater effluent used as geochemical marker to determine rates of estuarine sediment accretion and sea-level rise

48


Owing to the relatively limited number of

60

Co sources and pathways for its

entry into the environment, very little literature exists regarding its biological uptake, accumulation and potential toxicity to organisms. Studies that have made reference to those particular characteristics do not discuss its toxicity potential in great detail (e.g. Toth & Alderfer, 1960a, 1960b; Showler et al., 1988; Capowiez et al., 2001; Sattar et al., 2002; Shinonaga et al., 2005; Gudelis et al., 2008), but infer that its likely rate of transfer and accumulation from soils into and through the food-chain, and its injurious potential to organisms at low concentrations are minimal. This inference is corroborated, albeit by logic alone, via the wide variety of ecological and medical tracing applications where

60

Co has been introduced onto or into numerous organisms

(including humans) during tracing investigations, some of which have previously been referred to in Table 2.2. With regard to the use of

60

Co in sediment tracing studies, Toth & Alderfer

(1960a) document two of the earliest applications, with the first providing details of the development of a procedure for uniformly tagging (labelling) water stable aggregates (WSAs) by immersion into solutions of

60

Co and water. The success of the labelling

technique culminated in the same authors conducting a 1 yr. investigation into the formation and breakdown of the

60

Co-labelled WSAs (Toth & Alderfer, 1960b). The

reported success of both the labelling technique and the tracing application provided an initial indication of the potential value of

60

Co as a suitable candidate for sediment

tracing work of a similar nature. The apparent dearth of sediment-related tracing investigations since 1960 implies, however, that the versatility and value of

60

Co, as

concluded by Toth & Alderfer (1960a, 1960b), may therefore have been ignored or inadvertently overlooked. 2.2.2 134Cs: Origin & Environmental Characteristics Caesium-134 (134Cs) is an anthropogenic gamma-emitting radionuclide, also with a relatively short half-life (t0.5) of ~ 2.06 years (Avery, 1996; Real et al., 2002; Khan, 2003; Lee et al., 2007). It is a fission product, formed almost exclusively in nuclear reactors by neutron activation of the stable isotope Caesium-133 (133Cs). The activation process involves the capture and absorption of a free neutron into the nucleus of

133

Cs, thereby transforming it into the unstable radioisotope

Upon transformation,

134

134

Cs (Avery, 1996).

Cs also decays by ejecting β-particles and emitting γ-radiation

(Lee et al., 2007). After one half-life (~ 2.06 yrs.), radioactive decay results in the transformation of 50% of its nuclei into barium-134 (134Ba), its stable progeny (Rowan, 49


1990; Lee et al., 2007). Caesium-134 emits γ-radiation at photo-peaks of ~ 563, ~ 569, and ~ 802 KeV, all of which yield at very low levels (i.e. < 15%). It also emits γradiation at two additional photo-peaks of ~ 605 and ~ 796 KeV, both of which are at very high yields of ca. 98% and 85% respectively. Similar to

60

Co, it is also

theoretically possible to obtain reliable radiometric results from activities as low as 1.0 Bq kg-1 of (dry) sediment, again, assuming long (i.e. ~ 1 d.) counting times (Grapes, Pers. Comm.). Due to the very limited opportunity for neutron-capture by 133Cs during nuclear explosions, virtually no

134

Cs was produced and deposited as fallout from the

atmospheric detonation of thermonuclear weapons during their testing from the ca. 1950s to 1970s (Avery, 1996; Rühm et al., 1996). However, substantial trans-boundary contamination occurred across many countries as a direct result of the Chernobyl nuclear reactor explosion in Ukraine in late April, 1986 (Silver, 1987; Walling et al., 1989; Rowan, 1990; Rowan & Walling, 1992; Higgitt et al., 1993; Avery, 1996; Rühm et al., 1996; Quine et al., 1999; Spezzano, 2005). Owing to its short half-life, sufficient time has now elapsed since this accident that any remaining inventory has now decayed to below detection levels (McKenna & Longworth, 1995; Quine et al., 1999). Localised sources of low-level

134

Cs are also released under controlled conditions in wastewater

effluent from a number of nuclear reprocessing plants similar to those previously mentioned for

60

Co (McKenzie et al., 1989; Rowan, 1990; McKenna & Longworth,

1995; Avery, 1996). Like

60

Co,

134

Cs is therefore largely absent from the wider

environment owing to its anthropogenic origin, its relatively short half-life and its restricted release into the environment. For reasons identical to those described previously for

60

Co, this is expected to simplify and facilitate its use during sediment

tracing investigations. With regard to its environmental characteristics, estimates of Kd values in varying soils, minerals and also on urban surfaces indicate that the rate and strength of sorption is rapid and largely irreversible (Cremers et al., 1988; Avery, 1996; Sanchez et al., 2002; Khan, 2003; Spezzano, 2005; Wicker & Ibrahim, 2006). Sorption onto such materials is also not as readily influenced by environmental parameters, such as pH, which typically govern the rate and strength at which many other radionuclides sorb to different media (Real et al., 2002; Payne et al., 2004; Twining et al., 2004). With regard to soil and sediment, it is especially particle size dependent however, and is essentially dominated by the clay and silt fractions within a given soil (Cremers et al., 1988; Avery, 1996; Twining et al., 2004; Spezzano, 2005). 50


Given these potentially beneficial characteristics, and aside from instances where 134Cs derived as fallout from the Chernobyl accident has been used to determine rates of post-Chernobyl sediment redistribution within some UK catchments (cf. Walling et al., 1989; Rowan, 1990; Rowan & Walling, 1992; Higgitt et al., 1993), surprisingly little literature was found where it was deliberately and artificially applied to soil for the specific purpose of using it as a tracer. To demonstrate this point, information regarding its use was accrued from 20 separate investigations dating from 1995 to 2008 (Table 2.3), each of which has been attributed a research theme according to the nature of the work. Table 2.3 A selection of literature that demonstrates the wide variety of applications where 134Cs has been used as a tracer to determine varying environmental processes. Data Source

Year of Publication


McKenna & Longworth

1995

Determine mobility in marine & terrestrial ecosystems

Zehnder et al.

1995

Determine m obility through a biological ecosystem

Avery

1996


Bailly du Bois

1996

Tracing sea water movement in a marine environment

Rühm et al.

1996

Determine mobility in an edaphic ecosystem

Quine et al.

1999

Sedim ent tracing in an agricultural environment

Tyson et al .

1999


Bailly du Bois & Guéguéniat

1999

Tracing sea water movement in a marine environment

Syversen et al.

2001

Sedim ent tracing in an agricultural environment

Brambilla et al.

2002


Massas et al.

2002


Real et al.

2002

Determine mobility in urban environm ents

Sanchez et al.

2002


Khan

2003

Determine mobility in all environments

Twining et al.

2004


Payne et al.

2004

Determine mobility in all environments

Spezzano

2005


Wicker & Ibrahim

2006

As an analogue for another radionuclide

Sandeep & Manjaiah

2008

Determine mobility from an edaphic to a biological ecosystem

Again, although this list is neither exhaustive nor comprehensive, it was found that the majority (i.e. 75%) of the investigations focused on the mobility of

134

Cs in

varying ecosystems and / or environments, or used 134Cs as an analogue for determining the mobility of another radionuclide. In five (25%) investigations, it was used as a tracer of environmental processes per se. Only two investigations (i.e. 10%) used

134

Cs as a

tracer in the context of studying soil erosion or sediment transfer (e.g. Quine et al.,

51


1999; Syversen et al., 2001). Aside from issues regarding the potentially harmful and injurious nature associated with

134

Cs (and all radioactive material in sufficient

quantities), and given its attributes and environmental characteristics outlined above, surprisingly little literature is available which documents the use of 134Cs in soil erosion investigations. As with

60

Co, this again suggests that its potential capability as a

sediment tracer may have largely been ignored or overlooked. Possible reasons for the limited use of either radionuclide are discussed in the following sub-section.

2.3 Health & Safety & Gamma Radiation Measurement During tracing investigations where radioactive materials are used, legal constraints dictate that health and safety issues must be adhered to at all times. Investigators are thus legally and ethically obliged to undertake such investigations in a manner likely to cause minimal interference to themselves, to the general public and to biota within the particular ecosystem under investigation (Toth & Alderfer, 1960a; Duursma et al., 1973; Sauzay, 1973; Fullen, 1982; Ridley, 1986; Showler et al., 1988; Quine et al., 1999). Aside from the attributes of

134

Cs, and

60

Co as sediment tracers

outlined previously, reasons for their apparent under-utilisation in soil erosion investigations may relate to health and safety issues, to the legislation concerning their use and to the possible stigma associated with their deliberate release into the environment. For gamma-emitting radionuclides in particular, the main risks associated with their use arise from the internal or external over-exposure of organisms to the effects of ionizing radiation (Duursma et al., 1973; Sauzay, 1973) and for this reason, it is necessary to adhere to national legislation governing the use of such material (Fullen, 1982; Connor et al., 2007). For the UK, the 1986 Atomic Energy and Radioactive Substances Act No. 1002 places a maximum permissible activity concentration of 0.4 Bq g-1 of any solid matter which remains substantially insoluble in water (such as sediment and soil) when converted to or considered to be waste material (Ridley, 1986). Waste material exceeding this upper threshold is therefore deemed to be of a hazardous nature and the generation, storage and disposal of such material typically requires additional permits (Connor, 2007). However, problems associated with the disposal of modest quantities of relatively low-level radioactive waste soil can be legitimately avoided by incorporating it with unlabelled soil to dilute the activity concentration (Sailerova, Pers. Comm.). Other ways in which the inherent dangers associated with using radioactive material can be minimised involve maintaining the radiation exposure ‘as-low-as52


reasonably-achievable’ (ALARA) and involves following three relatively basic safety strategies that reflect shielding, time and distance (Anon., 2003; Connor et al., 2007). For shielding, exposure to radiation can be substantially reduced by enclosing and storing the radioactive source within a container made from a suitable material of an appropriate thickness until the radioactive material is required for use. With regard to time, the radiation dosage can be significantly reduced by carefully controlling the time spent in proximity to the radiation field. Based on this tenet, it is a requirement that work-schedules are carefully planned before entering the radiation field to enable the work to be undertaken as quickly and efficiently as possible. With regard to distance, radiation levels decrease by a factor of the square of the distance from a radioactive source (Anon, 2003; Connor et al., 2007) and the use of remote handling devices, used in conjunction with personal safety equipment, all contribute to achieving this. In some situations where radioactive sources are above a certain threshold, it may be necessary as part of a radiation protection strategy to monitor the radiation exposure at point locations using hand-held instruments, such as a Survey Meter, or to measure the radiation exposure to operators and personnel using personal dosimeters worn outside of protective clothing. Both approaches require a means of measuring radioactivity. For soils and sediments, the unit for measuring radioactivity recognised by the International System of Unit (SI) is the bequerel (Bq). One Bq is equal to 1 decay per second (dps). Because this unit is small, it of often convenient to use additional prefixes, such as kBq, which is equal to 103 Bq and MBq, which is equal to 106 Bq. It is often necessary to use an additional prefix, for instance where the concentration of fallout radionuclides are being determined from small soil samples at typical environmental activity levels, which enables the activity concentration to be measured in milli-bequerels (mBq), which are equal to 10-3 Bq. The following text now describes the principles of gamma spectrometry to measure levels of radioactivity in soils and also describes the equipment used during such investigations.

2.4 Gamma Spectrometry 2.4.1 The Basic Principles of Gamma Spectrometry As previously outlined, each gamma-emitting radionuclide produces one or more photo-peaks, or signatures, at known locations in the electromagnetic spectrum (EMS). This allows each species to be readily identified and their relative prevalence, or 53


radioactivity, to be quantitatively determined. One of the most common and rapid methods for accomplishing this is to use a technique known as gamma spectrometry. The following text describes the principles of gamma spectrometry and the major components within a system that are required to identify radionuclides and measure radioactivity levels directly from samples of soil and sediment. Although there are several methods of determining levels of radioactivity from soil and sediment samples, high-resolution, low-level gamma spectrometers fitted with high-purity germanium (HPGe) detectors represent the most cost-effective (Wallbrink et al., 2002). A sample of sediment is placed in close proximity to the germanium crystal, which is housed within the detector-head. Gamma photons ejected from the sample interact with the Germanium (Ge) crystal, which in-turn, emits signals or pulses which correspond to the energy of the incoming gamma-photons. These are routed through a preamplifier to an amplifier and then to a multi-channel analyser (MCA), whereupon the signals are displayed as a spectrum on a computer screen. Since the energy of the emitted photons is unique to each radionuclide, this allows those radionuclides present in each sample to be identified and their relative activities to be quantified.

2.4.2 The Major Components of a Gamma Spectrometer The flowing text lists the major components of a laboratory-based gamma spectrometer, the design of which differs slightly from that of an in-situ or field-portable gamma spectrometer, which is described later in the chapter. As previously indicated, one of the most fundamental components of a gamma spectrometer is the high-purity germanium (Ge) crystal. Germanium crystals are mechanically ground to varying dimensions (e.g. ca. 75 mm long * 75 mm dia.), depending on the required level of analytical efficiency (Miller, 1998) and available budget. At the time of their manufacture, semiconductor materials are embedded on opposing sides of the crystal and these are used to establish electrical contacts through which a high voltage is applied when the detector is in an operational mode. The characteristics and configuration of the contacts incorporated within the crystal are particularly important for they ultimately determine the minimum and maximum resolution of a detector, which dictates the energy range and ability to measure certain radionuclides, especially those located at the lower end (i.e. < 100 KeV) of the energy spectrum towards the X-ray region. Depending on the configuration of the contacts, detectors are thus classified as either an ‘N’-Type or ‘P’-Type, the former of which are typically more efficient at detecting radionuclides located at the lower end of the spectrum (i.e.

210

Pb) (Miller, 1998; Wallbrink et al., 2002). The Ge 54


crystal is housed within the detector-head in a sealed vacuum and when subjected to high-voltages, behaves like an effective semiconductor. In a charged state, gamma photons ejected by radionuclides contained in a soil sample interact with the crystal where they are regularly ‘swept’ by a high-voltage electrical field to the contacts mentioned earlier in this sub-section. Electrons, initiated by the interaction of gamma rays with the crystal and the need to subject the crystal to high voltages, both result in the generation of heat, which requires controlling, as it produces unwanted background interference within the system. The detector-head is consequently normally connected to a cryogenic cooling system which consists of a Dewar filled with liquid nitrogen (N). An outline of a basic cooling system is shown in Figure 2.1.

Figure 2.1 A simplified annotated schematic of a Canberra HPGe detector regulated by a ‘Vertical Slimline Cryostat’ cooling system (taken from Wallbrink et al., 2002).

55


This arrangement typically sits under the spectrometer and connects to the crystal located within the detector-head via a cooling-rod, or ‘coldfinger’. A cryostat regulates and maintains the components housed within the detector-head at an optimum and stable operating temperature of ~ 77o K (~ -198o C). It also serves to maintain a high quality vacuum around the crystal to prevent the adsorption of contaminants onto its surface. Other benefits associated with the cryostatic cooling system include a reduction in external vibrations, which contribute to reducing microphonic background interference created from within the system itself (Wallbrink et al., 2002). Figure 2.2 shows a simplified schematic diagram of the major components of a gamma spectrometer. Commencing through the system from the previously discussed HPGe detector-head, the first major electrical component is a charge-sensitive preamplifier. This is usually integral to and incorporated within the detector-head and is also temperature controlled. The role of a preamplifier is to receive incoming electrical pulses from the crystal. It is connected to the crystal through the vacuum housing via electric feed-throughs to produce an outgoing electrical pulse which is directly proportional (i.e. linear) to the energy of the incident gamma rays originally emitted by any radionuclides present in the soil sample (EG&G ORTEC, 1990; Rowan, 1990; Wallbrink et al., 2002).

Figure 2.2 A simplified schematic diagram showing the main electrical components of a gamma spectrometry system (adapted from EG&G ORTEC, 1990).

The second major electrical component is the amplifier, which, as its name implies, substantially increases the signal from the preamplifier and also enhances the signal-to-noise ratio by filtering and shaping the incoming pulses. The third major electrical component is a pulse height analyser and analogue to digital converter (ADC), which, as its name again suggests, converts the analogue signal from the amplifier to digital values. Due to their high cost and ability to process only single incoming pulses, 56


information from multiple detectors can be fed into an ADC via a fourth component known as a multiplexer, or mixer-router (not shown in Figure 2.2) (Wallbrink et al., 2002). The subsequent information is then fed into a multi-channel analyser (MCA), which converts the pulses, or counts, from the appropriate detector and then distributes and stores the accumulated photo-peaks from each sample over the counting period, according to the energy value at which they were emitted, to form a spectrum. The MCA is also able to compensate for any internal background interference over each count time. As the MCA is usually interfaced directly with a computer, which arguably represents a sixth electrical component and enables the spectrum to be visually displayed, both are shown as one single component in Figure 2.2. The whole system requires a stable source of high voltage electricity and the management of multiple detectors can also be facilitated via the use of a nuclear instrumentation module (NIMbin), which enables multiple amplifiers and other sundry electrical components often used in large radiometry facilities to be conveniently housed, connected to a mains supply and linked to one central MCA and computer (Wallbrink et al., 2002). With regard to external components, arguably one of the largest and most noticeable is the thick (i.e. ~ 0.1 m) lead shield which encloses both the detector head and sample and serves to reduce background interference from external sources of gamma radiation. The exterior of the shields of two detectors can be seen in Figure 2.3, each of which is located above a separate Dewar filled with liquid N, which is used as a coolant.

Figure 2.3 Two of the sixteen laboratory-based gamma spectrometers based in the School of Geography at the University of Exeter.

57


The shields and detector-heads are typically cradled at height in a robust steel frame in order to facilitate the removal and replacement of a sample centrally onto the end-cap. When in operation, sliding lead doors completely encase both the detectorhead and the sample to reduce incoming gamma radiation from external sources. In addition to the lead shielding, a 2-10 mm thick steel or copper-liner is often inserted into the interior of the shield and this attenuates any photons that may be ejected internally from the lead shield (Wallbrink et al., 2002). Such a liner can be seen, along with the end-cap of the detector-head, in Figure 2.4.

Figure 2.4 A view inside a lead shield showing the copper liner and end-cap at the end of the detector head. When in operation, a plastic pot containing a soil sample would be placed on the end cap and the sliding lead doors, parts of which can be seen in the left and top-right of the photograph, would be closed to reduce external sources of background gamma radiation.

2.4.3 Gamma Ray Interactions with Matter & Detector Efficiencies Low-energy gamma rays interact with matter in three principal ways; by Compton Scattering, by the Photoelectric Effect or by Pair Production (other processes exist but are largely irrelevant in gamma spectrometry and are consequently not discussed). All three processes produce moving electrons (or positrons) within matter that can either be detected directly, or can result in the initiation of other electron processes so that electrical pulses can be indirectly detected. Detection of incoming signals received from a theoretically perfect detector would be able to convert 100% of the energy from each gamma decay into an electric pulse. However, exceptionally high 58


efficiencies are rarely attained and even the efficiency of two detectors with identical components may be different. Such variations depend on numerous factors, and include the quality and size of the Ge crystal, the quality and thickness of the shielding material, and also the species of radionuclide(s) under investigation and their relative activities within the soil sample (Wallbrink et al., 2002). Other notable sources of variations in the efficiency frequently occur due to differences in the mass and geometry of soil samples. In order to compensate for such variations, and for the general purpose of comparing data produced by different detectors, it is necessary to calculate the absolute efficiency of each detector for each radionuclide. This value is defined as a ratio of the number of impulses recorded by the detector to the number of gamma rays produced by a known radiation source over a given unit of time (Rowan, 1990; Wallbrink et al., 2002). Efficiency values depend on the type of detector (i.e. ‘N’ or ‘P’-Type) and also on the physical characteristics of different soils, the latter of which can often result in large differences between the mass and geometry of individual soil samples and can conspire to significantly influence the accuracy of the final data (Wallbrink et al., 2002). Although, larger sediment samples tend to promote greater accuracy in most systems, photons emitted from within the central portion of the soil matrix of larger samples can often self-attenuate and thus do not interact with the crystal. Samples must therefore be placed as close to the detectorhead as possible to ensure that the maximum amount of photons emitted from the sample can interact with the crystal. The most effective way of achieving this is to ensure uniformity in the mass of each sample, and hence in their geometry, as this ensures that the distance between each sample and the detector-head remains constant.

2.4.4 The Laboratory-Based Gamma Spectrometers & Radiometric Equations In an attempt to compare data from different detectors, and due to the location of 60

Co at the higher end of the gamma spectrum (i.e. ~ 1171 KeV), all laboratory-based

radiometric assays were undertaken on three EG&G ORTEC high-purity coaxial Ge detectors based in the Geography Department at the University of Exeter. These were calibrated for the analysis of

134

Cs and

60

Co by the laboratory technician using

appropriate calibration standards and ensured that measurements of radionuclide concentrations obtained from small sediment samples and samples with relatively low activity levels were accurate. For environmental radionuclides analysed by gamma spectrometry at Exeter University, activity concentrations are typically determined from soil samples with a standard mass of 100g (± ~ 0.2 g) (Grapes, Pers. Comm.). Due to 59


the nature of some of the types of erosion investigations and the way in which soil samples were generated (i.e. removed from small areas or over very short timescales by surface runoff, or derived from thin core-sections), it was expected that a vast proportion of samples undergoing radiometric assay would weigh substantially less than the standard 100 g sample size. As mentioned earlier, analyses of samples of different mass and geometries can introduce potential errors in the results, owing to variations in the relative distance from the radioactive source to detector-head. In anticipation of this problem, it was deemed essential to calculate the efficiency of the detectors using a number of standards spanning a range of different masses in an attempt to reduce such errors. A series of 20 independently prepared standards were made for each radionuclide. The smallest standard weighed 5 g and the mass of each subsequent standard thereafter increased at 5 g intervals (i.e. 5 g, 10 g, 15 g…etc.) up to a maximum 100 g mass. The soil used for each standard was labelled at an activity concentration of 1 Bq g-1 of sediment, which approximated to the mean activity of the soils that were labelled and tested in the majority of laboratory-based investigations conducted in the UK. The efficiency of each detector was then calculated for each set of standards using Equation 1.

de = (A/T)/(dps(-Ln(2)·t/t0.5))

[1]

Where de = the detector efficiency, A = the peak area, T = the count time associated with each assay (s), dps = the number of gamma disintegrations per second, t = the amount of elapsed time (d) between when the radionuclides were received and when each measurement was recorded, and t0.5 = the radionuclide half-life. Where applicable, the standard with the mass most closely corresponding to the mass of a particular sample generated from an investigation was used as a basis for determining the most appropriate detector efficiency to incorporate in Equation 2. A Region of Interest (ROI) was manually established around the peak energy range of each radionuclide. Mean values of the first and last ca. three channels on either side of the peaks were used to determine background radioactivity levels and this was removed from the peak area under the ROI. Activity concentrations for each sample were thus calculated using Equation 2:

A = (C)/(T·M·de)

[2]

60


Where A represents the specific activity of the radionuclide (mBq g-1), C·represents the area within the ROI multiplied by 105 (to express the activity in mBq g1

), T represents the counting time for each sample (s), M represents sample mass (g),

and de represents the detector efficiency for each radionuclide at the appropriate sample mass. Owing to the different half-lives of the two radionuclides, it was necessary to decay-correct all radiometric data to the 1st June 2006, when both radionuclides were received from the manufacturer. Decay correction coefficients for each sample were calculated using Equation 3: Adcc = A(λ·t/t0.5)

[3]

Where Adcc represents the decay correction coefficient, A represents the activity concentration of the sample from Equation 2 prior to decay-correcting, λ represents a disintegration constant (Ln/(2)), t represents the amount of elapsed time (d) between the time the radionuclides were received (01/06/2006) and when each measurement was recorded, and t0.5 = the half-life of the radionuclide. Values obtained from Equations 2 and 3 were then used to decay-correct each sample using Equation 4.

ADC = A·Adcc

[4]

Where ADC represents the decay-corrected activity concentration, A represents the sample activity prior to decay-correction and Adcc represents the decay-correction coefficient.

Where appropriate, the decay-corrected activity concentration derived from Equation 4 was converted to Bq for each sample using Equation 5:

ADC = ADC(mBq)/1000·M

[5]

Where ADC represents the decay-corrected activity (Bq), ADC(mBq) represents the answer from Equation 4 (mBq), and M represents the sample mass (g).

Although nuclear decays occur randomly in time, sufficient certainty exists to predict that 50% of an initial activity decays away within the first half-life of a given 61


radionuclide. Thus, the number of disintegrations measured for a given period of time represents an average value, with a typical uncertainty associated with that value. Low activity concentrations typically result in larger error uncertainties. Since the uncertainty can be reduced by using longer counting times per assay (Rowan, 1990; Wallbrink et al., 2002), it was necessary to devote sufficient counting time to each sample in order to obtain reliable counting statistics. The statistical error associated with the spectral analysis of the photo-peaks was calculated based on the spread of the distribution around the central peak. Although the nuclear decay process conforms to a Poisson Distribution, a Gaussian Distribution satisfactorily describes the spectral function where the count time and / or counting statistics are sufficiently high (EG&G ORTEC, 1990). The measured error associated with each value was calculated at the 90% confidence level to determine the quality of the radiometric data using Equation 6:

% Error = 165· ( ( A)) / A

[6]

Where A represents the area within the predefined ROI, and 165 represents the value of A (as a percentage value) to within 1.65 standard deviations from the mean value when calculating the error at the 90% confidence level. In some instances, it was necessary to calculate areal activities (Aa) from bulk cores to derive a total inventory value per unit area (m2). This was done, where applicable, using Equation 7:

Aa = A·MT/S

[7]

Where Aa represents the inventory per unit area (Bq m2), A represents the activity of the sub-sample of the bulk core, MT represents total mass of the bulk core (kg) and S represents the surface area of the corer (m2).

2.4.5 The Field Detector Radiometric assays conducted in the field were undertaken using an EG&G ORTEC laboratory-grade (EG&G ORTEC, 1998) in-situ or field-portable gammaspectrometer, hereafter referred to as the field detector. The field-detector is a highPurity Germanium (HPGe), high resolution N-type coaxial detector with a 70 mm dia. detector head and with a relative efficiency ca. 40%. It was coupled to a NOMAD Plus power supply and MCA, interfaced with a lap-top computer (He & Walling, 2000; He et 62


al., 2002). A 5 l Dewar is incorporated within the body of the detector and this provides capacity to store 5 l of liquid N, which is typically sufficient to operate the detector for the equivalent of ca. 24-36 hrs. of constant use. Power for both the detector and lap-top was provided by a small portable Honda EM650 petrol-driven generator (output: 240V, max. 650W (5.4 amp)). A circuit-breaker and voltage stabiliser was fitted in-line between the generator and the detector to protect the system against electrical surges. A standard analytical protocol was developed for in-situ gamma spectrometry of environmental fallout radionuclides which typically involved mounting the equipment on a tripod at ca. 1 m above the soil surface of a particular study area (e.g. Andrási et al., 1987; Rybacek et al., 1992; Zombori et al., 1992a, 1992b; Tyler et al., 1996a; Miller, 1998; Tyler, 1999; He & Walling, 2000; He et al., 2002). Depending on the energy of the source and its distribution within the soil profile, this approach can provide the field-detector with a ‘field of view’ representing several hundred m2 (equivalent to a radius of ca. 8 m) (Miller, 1998). During investigations where in-situ measurements were required, the technique adopted by this study differed slightly from the standardised approach described above. Instead, the field detector was mounted very close to the ground during each assay in order to measure changes in the areal inventory emitted from small source areas where small quantities of sediment, artificially-labelled with either radionuclide, had previously been deposited (Chapter 3), or from small areas of soil directly labelled with either radionuclide (Chapter 5). For these investigations, it was necessary to develop a robust means of supporting the detector at an appropriate height above the soil surface. A support cradle measuring ca. 400 mm * 400 mm * 500 mm high was constructed from lengths of ca. 25 mm dia. steel tubing, which enabled the field detector to be securely mounted and easily repositioned as required (Figure 2.5). Four legs were used to support the cradle, each of which could be independently extended or retracted according to the topography. This design also allowed the detector to be positioned so that the angle of the detector-head was parallel with the soil surface and at a constant distance from the detector-head to the soil surface during each assay.

63


Figure 2.5 The field detector was mounted in an adjustable support cradle.

Calibration of the field detector was undertaken using a similar protocol to that described earlier for the laboratory-based detectors. Prior to undertaking any in-situ measurements, the field detector was pre-filled with an appropriate volume of N and a suitable cooling-down period of ca. 12 hrs. was provided before attempting to operate the detector. Immediately before attempting to record any in-situ measurements, the equipment was firstly made operational and left for ca. 30 mins., to allow the electrical components to attain their optimum working temperatures. An independently produced standard of known mass and activity was periodically analysed during field-work and the obtained activities were decay-corrected and compared (using an independent samples t-test (P < 0.05) with previous measurements in order to ensure the consistency of the data. The radiometric inventory was calculated using Equation 8.

A = (C)/(T·de)

[8]

Where A represents the specific activity of the radionuclide (mBq g-1), C represents the area within the ROI multiplied by 105 (to express the activity in mBq g-1), T represents the counting time during each assay (s), and de represents the detector efficiency for each radionuclide at the appropriate sample mass. Equation 8 is similar to 64


Equation 2, but the potential variability in the field of view of the detector during in-situ measurements made the sample mass (M) un-quantifiable and it was hence removed from the equation. Furthermore, the same nominal detector efficiency (de) value was attributed to all in-situ recorded radiometric data. This was justified based on the operational limitations of in-situ gamma spectrometry according to Miller (1998), who demonstrated that any effects caused by deviations in the geometry, to differences in the soil bulk density, or self-attenuation of the gamma photons would be negligible since all sources of radioactivity would be at or very near the soil surface.

The final section of this chapter now briefly outlines the provenance of the radioactive material.

2.5 The Radionuclides: Purchasing & Provenance Both radionuclides used in the soil erosion investigations were purchased from Eckert & Zeigler Isotope Products based in Berlin, Germany and were delivered to Exeter University on 1st June 2006. For

60

Co, 3.7 MBq of cobalt chloride (CoCl2)

activity was sent in 10 ml of 0.1 M HCl solution. For

134

Cs, 3.7 MBq of caesium

chloride (CsCl2) activity was also sent in 10 ml of 0.1 M HCl solution and both came with a nominal ± 15% activity uncertainty (Eckert & Zeigler, 2005). Upon delivery, legislation dictated that both were stored in a lead-lined secure facility, whereupon each was partitioned and dispensed upon request at an appropriate dilution (typically 100 Bq ml-1) in additional deionised water by the Exeter University Radiation Officer, for use in the tracing investigations.

The following five chapters now each outline a separate experiment, all of which were conducted in a variety of different erosion scenarios in order to comprehensively test the ability of both 134Cs and 60Co as tracers of fine-sediment.

65

CHAPTER 3: The Extent of Floodplain Sediment Remobilisation During Single Overbank Flood Events

CHAPTER 3: DOCUMENTING SEDIMENT REMOBILISATION ON RIVER FLOODPLAINS DURING SINGLE OVERBANK FLOOD EVENTS

3.1 Overview The aim of this component of the study was to explore the potential for documenting the extent of remobilisation of recently-deposited sediment on river floodplains over successive overbank flood events. The study was conducted on a floodplain in two different river catchments in Devon, UK and the approach involved depositing small quantities of fine sediment, pre-labelled with artificial radioisotopes, during a simulated flood event on each floodplain. The radiometric inventory associated with each deposited aliquot was recorded before and after a flood event and the reduction in inventory values were interpreted as evidence of remobilisation of the labelled sediment. Two separate flood events were recorded at each site and the percentage proportion of the inventory reduction indicated the magnitude of remobilisation of the labelled sediment at each location. This information was then correlated with two potential controlling variables, distance from the river channel and inundation depth.

3.2 Introduction River floodplains represent important stores of fine-sediment ( 0, this inferred that deposition of sediment had occurred at those locations. The second step in the accounting procedure involved calculating the ratio of the additional inventory and the total eroded sediment removed from elsewhere within the plot, which was undertaken using Equation 4b: 141

CHAPTER 5: Assessing the Extent of Soil Redistribution on Livestock-Poached Pasture

Actsed = Acttotal/Etd

[4b]

Where Actsed represents the activity to sediment ratio (Bq mm-1), Acttotal represents the total eroded inventory for each plot at each re-measurement (Bq), and Etd represents the total depth of eroded sediment (mm) for each plot at each remeasurement. Where applicable, the results from Equation 4b were then used in a final step to determine the depth of sediment deposition using Equation 4c:

Depsed = IF(ΔZ 0.3 m from the plotoutlet, 4.5% (9.12 g) of the original mass was removed by surface runoff during the 76 day operational period, whereas the remaining 95.5% was retained in the bulk soil. For 60

Co-labelled casts originally placed between < 0.3 m from the plot outlet, 14.1% (26.82

g) of the original mass was removed by surface runoff over the same time period, whereas the remaining 85.9% was retained in the bulk soil. A total of 59.51 g of unlabelled sediment was removed during the operational period, the first ca. 9% (5.19

182


g) was removed between cumulative days 5-32, and the remaining ca. 91% (54.32 g) between cumulative days 38-76. Table 6.5 A summary of the corrected partitioned data for the experiment undertaken on pasture. 134

Cs (travel distance >0.3 m )

60

Co (travel distance 0.3 m, and < 0.3 m for

60

134

Cs-labelled

Co-labelled sediment. Le Bayon & Binet

(2001) speculated that the movement of cast-derived sediment on gentle slopes likely follows cycles of erosion and deposition, with the sediment moving only relatively short distances between each cycle. This theory is supported using data presented in Figure 6.17, and relates, again, to differences in the labelled sediment concentrations recorded during the early stages of the experiment. The concentration of 134Cs-labelled sediment was extremely low for cumulative days 5 and 7, and notably lower than

60

Co values.

Values then increased at cumulative day 13 and peaked at cumulative day 20. These variable concentrations contrasted with

60

Co values, which were initially high and

remained as such until cumulative day 20. Differences in the initial concentration characteristics indicated by the two radionuclides, along with the slight time-lag in the 134

Cs peak, are interpreted as evidence of the longer travel distance of the 134Cs-labelled

sediment from its original location to the plot outlet. Since it is feasible that the 134Cslabelled sediment eroded from further upslope was subjected to more cycles of erosion and deposition before leaving the plot, a substantial proportion of the eroding sediment may have still been in-transit but not reached the outlet of the plot within the 76 day monitoring period, which may explain why markedly less

134

Cs-labelled sediment was

recovered in surface runoff than 60Co-labelled sediment.

6.5.2 Cultivated Soil As data listed in Table 6.7 revealed, 20.62 g, or 10.5% of the original 197 g of 134

Cs-labelled casts was removed in runoff from the unconsolidated soil during the first

rainfall event. A further 8.2% (16.21 g) was removed in runoff from the crusted soil during the second rainfall event, both of which cumulatively represented a total of 36.6 g of sediment, or 18.7% of the original mass of labelled casts. These values contrasted with the 73.03 g, or 36.7% of the original 198.5 g of

60

Co-labelled casts which was

removed by runoff from the crusted surface during the second event. As previously mentioned, key environmental parameters, including gradient, plot size and rainfall 189


intensity were constant over both events. However, large differences were recorded between the two data-sets, both in terms of the overall sediment yield from each event, and also in the relative proportions of labelled material removed by surface runoff. These differences are attributed to variations in soil surface conditions between the two events, but specifically to the formation of the crust during the second event. Arguably, data from Figure 6.18 reveals that the magnitude of erosion during the first event was much higher than the second event and this may have been due to the availability and physical characteristics of the unconsolidated sediment and the relative ease with which this material was entrained and transported (Singer & Shainberg, 2004). Linked to this theory is the lack of clear evidence from Figure 6.18 for the preferential removal of 134

Cs-labelled sediment during the first event. Due to the fine nature and plentiful

supply of unconsolidated material and the finite threshold transport capacity of the surface runoff, sediment from eroding casts may, in effect, have been ‘in competition’ with the host soil to be transported. This would ultimately have limited its movement and possibly also promoted the in-washing of the dispersed labelled sediment into the large interstices of the unconsolidated bulk soil (Kladivko et al., 1986). In contrast, the magnitude of erosion during the second event was considerably lower than for the first event, and this suggests that the surface crust afforded a level of protection to the soil surface which reduced the overall production of sediment during the second erosion event (Levy et al., 1986; Singer & Shainberg, 2004). Crusting controls soil surface conditions by reducing its permeability (Onofiok & Singer, 1984; Levy et al., 1986; Singer & Shainberg, 2004; Le Bissonnais et al., 2005) and under certain situations on low slopes, and in the absence of an alternative source, this may lead to a detachment-limitations in the supply of sediment (Singer & Shainberg, 2004). Although differences in the relative magnitude of erosion recorded over the two events would appear to support this claim, data from Figures 6.18 and 6.19 indicate that 134Cslabelled sediment continued to represent a notable source of eroding sediment throughout the second event. Since the casts were completely dispersed after the first rainfall event, this suggests that sediment from the

134

Cs-labelled casts must have been

incorporated into the surface crust. The relatively small amounts of sediment from the 134

Cs-labelled casts that continued to be eroded during the second event, as evidenced

from samples 9-15, along with the relatively large proportions of unlabelled sediment recorded in samples 10-15, is interpreted as evidence of the continued susceptibility of the crusted soil to erosion during the second event and was probably due to raindrop

190


splash detachment of sediment from the crusted soil surface (Bradford et al., 1986; Levy et al., 1986; Singer & Shainberg, 2004). Data presented in Figures 6.19 and 6.20 also revealed that sediment from 60Colabelled casts was initially removed in a large pulse (Sample 9). Although values thereafter decreased sharply, the rate of removal shown in Figure 6.20 provided an indication of the initial importance and potential susceptibility of sediment from these casts to erosion and transportation. It is therefore pertinent to reiterate that, in comparison to 60

134

Cs-labelled casts, approximately double the amount of sediment from

Co-labelled casts was eroded in approximately half the time, and this effect is

attributed to the presence of the surface crust.

6.6 Assessment of the Approach The assessment of the approach focuses on the design of the experiments, and in particular, to their ability to faithfully simulate rates of disintegration and dispersal of earthworm casts. Attention is also given to the application of the tracers and the ways in which they were used to determine the fate of the dispersed sediment.

6.6.1 Uncertainty in the Labelling Technique Partitioning of runoff-eroded sediment initially yielded problems for some of the samples removed from pasture in terms of overestimated activity concentrations, and these required correcting. Possible reasons for the overestimated activities may have resulted from the non-uniform labelling of the casts. Additional work could have been undertaken to verify the success of labelling technique by, for instance, repeating the labelling procedure in order to confirm the results, and also by comparing the measured activity of intact casts with the measured activity of the same casts in a dispersed state.

6.6.2 Cast Durability and Resilience to Disaggregation The inherent difficulties associated with labelling sediment (Simm & Walling, 1998), especially of the nature and form of earthworm casts, along with the need to handle the casts prior to each rainfall simulation and the overall constraints on available time and resources, rendered the labelling approach described in earlier sections as the most practicable, especially in the absence of a suitable alternative method. As previously stated, integral to the labelling process was the need to subject all casts to cycles of drying and wetting. It is acknowledged that this procedure may have altered the aggregate stability of the casts (Shipitalo & Protz, 1988) and inadvertently rendered 191


them overly resilient to erosive forces and hence, unrepresentative of casts produced and eroded under natural conditions.

6.6.3 Runoff Eroded Samples: Volume and Quantity For the experiment on cultivated soil, 15 samples of runoff were collected in 10 l volumes over a combined period of approximately 150 minutes of simulated rainfall. Although these samples collectively represented all runoff, the temporal resolution of the data-sets may have been improved by collecting runoff more frequently and in smaller (i.e. 5 l) volumes. However, since inter-rill erosion was the required mobilising process and the available literature predicted low sediment yields from small inter-rill eroded areas on low slopes (e.g. Morgan, 1993; Whiting et al., 2001; Evans & Brazier, 2005; Yang et al., 2006), it was considered prudent to collect surface runoff in larger volumes as this provided a contingency to ensure that sufficient quantities of sediment with sufficiently high activities were obtained. This problem could have been overcome relatively simply however, by labelling the casts at initially higher activity concentrations.

6.6.4 Reduced Crusting by Earthworms Under natural conditions, the appearance of casts on the crusted surface after the first rainfall event would have required earthworms to be present and active in the soil. This was not possible however, since all large soil biota were removed prior to the experiment. It is therefore acknowledged that under normal circumstances, the prevalence of earthworm activities between rainfall events and in particular, the creation of burrow-openings in the soil surface would have fractured the surface crust (Kladivko et al., 1986) and promoted the infiltration of water and the re-incorporation of castderived sediment back into the surrounding soil. Moreover, the presence of surface casts might also have increased the soil surface roughness of the crusted soil, thereby reducing the velocity of the runoff and the magnitude of erosion (Kladivko et al., 1986).

6.6.5 Scale At the scale that which both experiments in this investigation were conducted, substantial proportions of cast-derived sediment were mobilised from small plots on both pasture and cultivated land during low to medium-intensity rainfall events on gentle slopes. In terms of maximum travel distance on pasture, notable quantities of sediment from dispersed 134Cs-labelled casts were transported a distance of > 0.3 m, and 192


considerably more sediment was transported from dispersed

60

Co-labelled casts over a

distance of < 0.3 m. It remains conjectural whether the data obtained from these experiments are indicative of erosion rates obtained from larger areas with similar topography (cf. Rejman et al., 1999; Parsons et al., 2006). 6.7 Conclusion The work described in this particular chapter has outlined the development of a technique for labelling earthworm casts. This technique has then been applied to document the fate of sediment from eroded casts on pasture and on cultivated land using the radioisotopes

134

Cs and

60

Co as tracers. Within the constraints of the experimental

procedures described above, and taking account of the uncertainties associated with labelling such structures in the manner described, the results indicate that sediment from casts eroded by surface runoff do significantly contribute to soil erosion from both pasture and from cultivated land on relatively gentle slopes. For the experiment conducted on pasture, it was concluded that the transportation distance of the eroded sediment was substantially further than had previously been predicted by other authors (e.g. Le Bayon & Binet, 1999, 2001) and travel distance potentially controlled the amount of sediment removed by surface runoff. The rate of removal also appeared to be determined by topography, the availability of fresh earthworm casts and the timing and intensity of rainfall events. Furthermore, and contrary to predictions, it was also found that casts continued to represent a source of eroded sediment for a considerable period of time after their initial formation (Binet & Le Bayon, 1999; Le Bayon & Binet, 1999). For cultivated land, it was concluded that the unconsolidated nature of the soil resulted in only a relatively small amount of 134Cs-labelled sediment being removed by surface runoff. Although the surface crust resulted in a reduction in the overall sediment yield during the second rainfall event, the magnitude of unlabelled material, along with the small contribution of

134

Cs-labelled casts still continued to represent substantial

source of sediment, which demonstrated the susceptibility of sediment from the crusted surface to erosion. The initially high rate of sediment removal from the

60

Co-labelled

casts during the second rainfall event indicated their susceptibility to erosion and this was attributed to a reduction in the porosity of the bulk soil owing to the development of the surface crust. In addition, it was further concluded that erosion during the second event was probably dominated by rainsplash detachment due to the presence of the surface crust and this is thought to have limited the supply of sediment available for transport and so promoted the removal of the 193

60

Co-labelled sediment from the crusted


surface. Applying this tracing technique on longer plots, on a broader range of topographies and conducted over longer timescales would represent a focus for future work.

194

CHAPTER 7: Documenting Sediment Source Changes During Inter-Rill and Rill Erosion

CHAPTER 7: DOCUMENTING SEDIMENT SOURCE CHANGES DURING INTER-RILL TO RILL EROSION

7.1 Overview The aim of this component of the study was to explore the potential for documenting sediment source changes during the transition from inter-rill to rill erosion. An erosion plot, artificial rainfall and a layered, or stratified tracing approach were used to investigate eight different agricultural soils from the south and southwest of England. The approach involved filling a 1.0 m * 0.2 m * 0.11 m erosion plot to a depth of 0.1 m with a soil labelled with 134Cs and then spreading an additional layer of the same soil labelled with 60Co over the first layer to a depth of ca. 0.01 m. The soil surface was then lightly textured in an upslope / downslope direction in order to simulate a seedbed-finish. The plot was pitched at a ca. 17% slope (ca. 9.5o) and each soil was subjected to artificial rainfall at an intensity of between 30-40 mm hr.-1 until surface flow was initiated and a rill developed. All sediment removed in surface runoff was collected throughout each rainfall event, recovered and radiometriclly analysed. Because different radionuclides were used to label the surface and sub-surface layers, changes in the radiometric signals associated with the runoff-eroded sediment could be used to demonstrate a shift in sediment source and the transition from inter-rill erosion to rill erosion during the course of each rainfall event. This information provided a basis for determining whether differences in the grain size composition of pairs of inter-rill and rill eroded sediment samples from each soil were significant. Sediment samples recording the highest activity for each radionuclide, and hence the highest contributions of sediment from the respective mobilisation process, were selected for particle size analysis. The grain size composition of pairs of sediment samples eroded by inter-rill and rill erosion from each of the eight soils was measured at corresponding 10 percentile intervals and statistically analysed in order to determine whether the differences were significant. 7.2 Introduction The recognition that sediment delivery by hydrological processes is limited either by the rate of soil detachment or by the transport capacity of the overland flow (Meyer & Wischmeier, 1969; Toy et al., 2002) led to the process of soil erosion being sub-divided into two fundamental mechanisms; rill and inter-rill erosion (Foster & Meyer, 1975 in Martinez-Mena et al., 2002), based on the different detachment and 195


transportation processes involved in the removal and conveyance of sediment (Zartl et al., 2001). A working definition of each mobilisation process is provided below and this allows fundamental differences between the two mechanisms to be differentiated Inter-rill erosion, also frequently referred to as infiltration excess overland flow, sheetwash or surface runoff, is defined as the movement of sediment on the soil surface by a thin slow-moving layer of water between rills (Alberts et al., 1980; Morgan, 1993). Particle detachment by inter-rill erosion is almost exclusively by impacting raindrops (Parsons et al., 1998) and once detached, material is then transported to rill-systems by the surface runoff. Although the detachment of sediment by raindrop impact is indiscriminate (Hairsine et al., 1999), the weak hydraulic competence of the overland flow is thought to be selective in the size of material that can be transported. This therefore limits the overall movement of eroded material and also reportedly results in the preferential transportation of fine-sediment (Alberts et al., 1980; Miller & Baharuddin, 1987). In contrast, rills are well-defined narrow, ephemeral channels, which range in both width and depth from a few centimetres to tens of centimetres (Alberts et al., 1980; Morgan, 1993; Nearing et al., 1997; Brunton & Bryan, 2000). They form where surface runoff becomes concentrated due to the presence of topographic features such as swales, cultivation furrows or wheelings, through changes in the soil surface roughness (Toy et al., 2002), or variations in vegetation density (Meyer, et al., 1975). Rill erosion reportedly accounts for the majority of eroded sediment removed from hillslopes owing to the greater erosive power of the concentrated surface runoff (Morgan, 1993). Consequently, much attention has been given to the hydraulic and sediment transport conditions during rill erosion in an attempt to identify the dominant controls and to thereby obtain improved predictions of rates of sediment transport. Conditions influencing sediment transport include slope, rainfall intensity, unit stream power, soil surface roughness, rill geometry, wetted area and the erodibility of the soil (Govers, 1992; Nearing et al., 1997; Brunton & Bryan, 2000; Zartl et al. 2001). The most important of these, particularly in uncompacted soils, are unit stream power and rill geometry (Govers, 1992; Nearing et al., 1997; Brunton & Bryan, 2000). Based on the range of sediment transport capabilities associated with each mobilisation process (Zartl et al., 2001), variations in the hydraulic conditions during the transition from inter-rill to rill erosion are thought to be responsible for changes to the grain size selectivity and this effect is thought to result in differences in the grain size distribution of the eroded soil (Young & Onstad, 1978; Proffitt & Rose, 1991; 196


Proffitt et al., 1991; Yang et al., 2006). Fully understanding this effect has important implications from both an agricultural and water-quality perspective (Stone & Walling, 1996) as fine particles act as a vector for the movement and transfer of nutrients and contaminants (Farenhorst & Bryan, 1995; Slattery & Burt, 1995; Nakamura & Kikuchi, 1996; Walling et al., 1998; Owens et al., 1999; Hogarth et al., 2004). Over substantial periods of time, the size selective mobilisation and transport of particles and aggregates by natural hillslope processes (Farenhorst & Bryan, 1995) can result in the redistribution of nutrients and contaminants, thereby leaving eroded areas depleted in nutrients and depositional areas enriched in contaminants (Basic et al., 2002). Considerable attention has thus been devoted to establishing links between mobilisation processes and grain size selectivity (e.g. Young & Onstad, 1978; Alberts, et al., 1980; Farenhorst & Bryan, 1995; Fox & Bryan, 1999; Chaplot & Le Bissonnais, 2000; Basic et al., 2002; Leguédois & Le Bissonnais, 2004). However, obtaining empirical evidence for such changes during the transition between inter-rill and rill erosion has proved difficult and in instances where such studies have been undertaken, the results have often been contradictory. The findings from a number of selected studies are summarised below. A substantial body of literature (e.g. Alberts, et al., 1980; Abrahams, 1998; Parsons et al., 1998; Leguédois & Le Bissonais, 2004; Ahmadi, et al., 2006) supports the theory that, in areas where inter-rill erosion is dominant, the low transport capacity of the surface runoff limits the size of particles and aggregates that can be transported. Consequently, it has been argued that the grain size composition of inter-rill eroded sediment will typically be finer than both the parent and rill eroded material (Legout et al., 2005). During experiments undertaken on both cultivated and uncultivated land however, Yang et al. (2006) found no difference in the particle size distribution of eroded material from an uncultivated plot during the transition from inter-rill to rill erosion, and minimal evidence of coarsening in grain sizes after the initiation of a rill on the cultivated plot. In direct contrast however, Young & Onstad (1978) reported that the grain size distribution of sediment mobilised by inter-rill erosion from a variety of soils was coarser than both the rill eroded and parent materials. Further evidence for the preferential detachment and movement of coarse to very coarse silt fractions (i.e. 16-63 µm) on cultivated hillslopes was reported by Ampontuah et al. (2006), and this resulted in the enrichment of upslope areas in clay material in comparison with downslope or depositional areas.

197


Whilst attempts have been made to couple transport capacities to mobilisation processes and incorporate this information into transport capacity equations (Huang, et al., 1999; Ahmadi et al., 2006), a clear consensus has still not been reached regarding differences in the physical characteristics of inter-rill and rill eroded sediment, despite the generation of a considerable amount of experimental data. This uncertainty provides the justification for undertaking the investigation described in this chapter. In the light of the contrasting results reported above, a tracing technique is proposed that attempts to provide evidence of changes in the grain size selectivity during changes in the mobilisation process. All experiments were conducted on a variety of agricultural (tilled) soils using an erosion plot and simulated rainfall and a non-directional hypothesis is used to argue that the particle size distributions of material mobilised from a common soil by inter-rill and rill erosion will differ. The aim of the investigation was to thus document differences in the grain size composition of runoff eroded sediment from a selection of agricultural soils using 134Cs and

60

Co as tracers to distinguish sediment mobilised by inter-rill and rill erosion. A

series of objectives were formulated in order to achieve the aim and these are as follows:



To identify and develop an appropriate method, in order to address the primary aim of the investigation and then to implement and replicate the method on a representative range of different agricultural soils;



To use the information provided by the tracing approach to allow changes in the mobilisation process for each sediment sample to be identified and to select samples that most effectively characterise each mobilisation process; and



To undertake statistical analyses on pairs of inter-rill and rill eroded sediment samples from each soil in order to assess whether differences between their grain size compositions are statistically significant.

The following sections of the chapter describe the methodology that was developed in order to achieve the aim. The results are presented and interpreted and the overall approach is assessed. Finally, the key conclusions are assessed and the implications of the investigation are considered.

198


7.3 Method 7.3.1 Choice of Tracers Detailed information regarding the environmental characteristics and behaviour of the artificial radioisotopes

134

Cs and

60

Co, and the justification for their selection as

tracers in this investigation has been previously discussed in Chapter 2. 7.3.2 Experimental Design Agricultural soils were obtained from several different floodplain and hillslope locations in three separate river catchments in the south of England. Approximately 30 kg of soil were excavated at each location from the surface to a depth of ca. 200 mm, which was selected as being representative of typical tillage operations (van Muysen et al., 2006). Samples were then transported to the laboratory for processing and labelling with radionuclide material. Processing consisted of air-drying each soil sample to ca. 10% moisture content. Each was then passed through a 10 mm soil-sieve in order to remove large stones and large soil biota, organic material and general debris and to provide a range of aggregate sizes representative of an agricultural seedbed (Leguédois & Le Bissonnais, 2004). Calculating the bulk density for each soil and multiplying this by the dimensions of the erosion plot determined the approximate mass of soil required for each layer of soil within the erosion plot. A further 25% of soil was added to this value as a contingency in order to ensure sufficient labelled soil was available to fill the plot to the required depths. Labelling the soil with radionuclide material was achieved by firstly determining the volume of water needed to saturate 1 kg of air-dried soil (typically ca. 0.7 l kg-1) to a consistency resembling a slurry. The volume of water was then extrapolated to match the total mass of sediment required to fill the erosion plot to the desired thickness for each layer. For the lower layer, this was typically ca. 18 kg of soil requiring 12.6 l of water, and ca. 2.5 kg of soil for the upper layer, requiring 1.75 l of water. Each radionuclide solution was diluted and dispensed from a stock source at an activity of 100 Bq ml-1. The required activity concentrations of each radionuclide (i.e. Bq ml-1) were converted to a mass value (i.e. Bq g-1) and were dispensed by weight using a pipette and electric scales to an accuracy of 4 decimal places. All soils were labelled at an activity concentration equivalent to 1 Bq g-1 of sediment and each radionuclide solution was added to a quantity of domestic water which was commensurate with the required mass of sediment for each layer. The mixing procedure 199


involved placing each unlabelled soil into a large plastic container and the radionuclide solution was carefully poured onto the soil and the contents stirred continually until a slurry consistency was obtained. The mixture was re-agitated periodically for a further ca. 4 hours, then carefully transferred into shallow trays and warmed in an oven at 300 C in order to evaporate the majority of the liquid. Once the majority of the liquid had evaporated, the mixture was then removed from the oven and left to air-dry and periodically re-agitated in order to facilitate the air-drying process. This also prevented the soil from setting into solid blocks and served to maintain the structure of individual soil aggregates. The erosion plot, shown in Figure 7.1, was fabricated from 10 mm thick plexiglas material and measured 1.0 m long * 0.2 m wide * 0.11 deep and provided a surface area of 0.2 m2. The frame of the plot was mounted onto a 2 mm thick aluminium base-plate using internal brackets, which also ensured that the structure remained rigid when full of wet soil. An aluminium trough was fitted at an angle across the downslope end of the plot and was used to collect the surface flow and eroded sediment. A 1 mm thick aluminium plate was profiled to form a very shallow ‘V’-shape to represent a swale and this was fitted internally across the downslope end of the plot. Justification for using this profile is discussed later in this section.

Figure 7.1. The erosion plot showing the profiled plate and collection trough (right foreground).

200


The erosion plot was filled with soil labelled with 134Cs to a depth of ca. 0.1 m, levelled and then covered with an additional ca. 0.01 m layer of the same soil labelled with

60

Co. During inter-rill erosion, soil mobilised by sheetwash will be derived

exclusively from the top layer labelled with 60Co. As the event proceeds and the surface flow becomes concentrated, a rill will begin to form and then incise down into the host soil and begin to mobilise sediment labelled with

134

Cs. A schematic diagram of the

stratified configuration of the labelled soil in the erosion plot is provided in Figure 7.2.

Figure 7.2. The stratified configuration of layers of 134Cs and 60Co-labelled soil, along with the overall concept of the tracing approach are depicted in this schematic cross-section of the erosion plot (diagram not to scale).

The concept of the stratified configuration using different radionuclides is based on work undertaken by authors such as Wallbrink et al. (1999), Whiting et al. (2001) and Yang et al. (2006), all of whom exploited the contrasting depth distributions of varying fallout radionuclides to determine changes in the grain size selectivity during the transition from inter-rill to rill erosion and establish the depth origin of the eroded sediment. The soil surface was shaped to mirror the profile of the aluminium plate, in order to encourage the initiation and development of a rill system within the centre of the erosion plot. The soil surface was very lightly textured (Figure 7.3) using a garden rake in order to represent the surface of an agricultural seedbed after sowing. This approach also provided a similar surface roughness for each soil in an attempt to

201


maintain a constant sediment transport capacity during each experiment (Abrahams et al., 1998).

Figure 7.3. The soil surface at the lower end of the erosion plot (central background) was profiled to represent a shallow swale in order to encourage the initiation and development of a rill down the centre of the erosion plot away from the edges. The soil surface was gently textured using a garden rake to represent a seedbed.

The erosion plot was positioned under the rainfall simulator and pitched at a 17% slope, which conformed with other similar experiments (e.g. Abrahams et al., 1998; Fox & Bryan, 1999; Zartl, et al. 2001; Martinez-Mena & Albaladejo, 2002), in order to ensure the rapid initiation of a rill. Approximately 24 hours before commencing each rainfall simulation, the soil surface was gently pre-wetted with domestic water applied from a garden sprayer. This minimised the time taken to initiate overland flow and also reduced the time taken to simulate the erosive conditions during an intense rainstorm (Miller & Baharuddin, 1987), thereby facilitating the mechanical breakdown of the aggregates by impacting raindrops (Leguédois & Le Bissonnais, 2004). Bowyer-Bower & Burt (1989) summarise the advantages of using simulated rainfall over natural rainfall events and these include the ability to reproduce rainfall on demand, at the desired location and intensity and for the required duration. A spray-type rainfall simulator was used during each experiment, the design of which consists of a 202


water pump, which supplies water from a domestic source to a series of 100 mm diameter mains pipes located parallel with each other at roof level and ca. 2 m apart. Water from the mains pipes passes through junctions fitted with 90o knuckle-bends, aimed downwards, and located at opposing sides of each mains pipe at 1.5 m intervals. Water is forced under pressure though nozzles fixed to the knuckle bends which create the raindrops. The diameter of the nozzles can be varied depending on the required drop-size and the rainfall intensity can be adjusted by varying the pressure of the water pump. For this experiment, a pair of 3 mm nozzles was used to create the rainfall. The distance from the nozzle-outlet to the soil surface was approximately 5.0 m and the median drop diameter was ca. 2 mm. The rainfall intensity for all experiments was between 30-40 mm hr-1, which generated ca. 300 ml min.-1 of surface runoff and generated sufficient runoff to fill one 500 ml bottle every ca. 90 s. The duration of each rainfall event for all soils was between 1-3 hours, but ultimately depended on the time taken for a rill to form and become sufficiently developed. The surface runoff from the plot was collected continuously in 500 ml plastic sealable bottles during each experiment as soon as runoff commenced and these were sequentially numbered and allowed to stand for ca. 48 hrs. in order to promote settling of the collected sediment. The majority of the supernatant was then carefully decanted and the remaining contents of each sample was transferred into 50 ml oven-proof glass beakers, placed into an oven and heated at 30o C in order to evaporate the remaining supernatant and so retain all of the sediment. Once evaporation had removed the majority of the liquid, samples were removed from the oven were periodically agitated with a rubber spatula and left to air-dry. Each sample was then recovered from the beaker, gently disaggregated using a rubber-headed pestle and mortar, weighed to two decimal places and stored in marked polythene bags in preparation for the radiometric assay. Radiometric assays were undertaken using a portable EG&G ORTEC field detector, the specifications of which have previously been described in Chapter 2. For this particular investigation however, the position of the detector was fixed for static use and fitted with a lead shield. Samples were placed in specially designed plastic containers measuring 70 mm deep x 75 mm diameter (surface area = 44.2 cm2), which matched the diameter of the detector head. The lead shield fitted around both the detector head and the sample to reduce the likelihood of interference by background gamma-radiation. Variations in the mass, and hence the geometry of each sediment

203


samples were accounted for by appropriate detector calibration (Miller, 1998) and the mean counting time for each sample was ca. 6,000 s. A Micromeritics Saturn Digisizer ™ 5200 was used to undertake particle size analysis of selected runoff samples from each experiment, and also on an unlabelled sample of each host soil. Three particle size analyses were conducted per sample and the mean of the three results was used to indicate the percentage proportions of sand, silt and clay, as well as the relative proportion of each sample in size-fractions ranging from 0.1 µm to 1,000 µm Pre-treatment of samples followed the protocol provided in the Exeter University Physical Geography Laboratory Manual & Methods (Anon., 2005). Between one and two grams (g) of sediment from each sample were placed into numbered glass beakers. Samples weighing less than the mandatory 1 g were merged with neighbouring samples until the minimum required mass (i.e. ≥ 1 g) was obtained. Approximately 10 ml of de-ionised water and 5 ml of hydrogen peroxide (H2O2) were then added to each sample using a pipette and these were allowed to effervesce for 2-3 hrs. Depending on the volatility of the reaction after this period, a further 5 ml of hydrogen peroxide was added if necessary. Samples were then allowed to stand for ca. 12 hours, after which, the beakers were placed onto hot-blocks, initially heated at 800 C and gradually increased to 1000 C until the supernatant became clear, indicating that the reaction was complete. Samples were then transferred into numbered centrifuge tubes and spun at 2,500 rpm for 1 hour. Prior to undertaking the particle size analysis, approximately 30 ml of a 0.4% sodium hexametaphosphate (NaPO3)6 solution was added to each sample and each was then vigorously shaken in order to facilitate redispersal the aggregates into primary particles.

7.3.3 Soil Characteristics & Background Information Eight different agricultural floodplain and hillslope soils of varying characteristics were obtained from a variety of sites located in a number of different catchments and sub-catchments in the south of England (Figure 7.4). Although the majority of hillslope soils were cultivated, the investigation was extended to include uncultivated floodplain soils in order to determine whether differences between those and hillslope soils was due to the latter soils being subjected to the effects of in-situ sorting (Farenhorst & Bryan, 1995; Stone & Walling, 1996; Ampontuah et al., 2006).

204


Figure 7.4. The locations of the sites from where soil samples were collected.

The soils under investigation were selected to be representative of the range of agricultural soil-types found within the southeast and southwest regions of England. General background information for each soil is listed in Table 7.1.

Table 7.1. Background information for each site. Location

Code

OS Grid Reference

River SubCatchment

River Catchment

Landform Feature

Land Use Type

Breaky Bottom

BB

TQ 056 405

n/a

River Ouse, East Sussex

Hillslope

Cultivated

Crediton/Copplestone

CC

SS 793 013

River Yeo

River Exe, Devon

Hillslope

Cultivated

Haddon Hill

HH

SS 288 966

River Haddeo

River Exe, Devon

Hillslope

Pasture

Laughton Levels

LL

TQ 124 496

Glynde Reach


Floodplain

Pasture

Southease

OF

TQ 057 426

n/a


Floodplain

Pasture

Hill Head Farm

OL

SS 672 181

River Mole

River Taw, Devon

Hillslope

Cultivated

Hill Head Wood

OP

SS 670 180

River Mole

River Taw, Devon

Hillslope

Cultivated

River Taw Floodplain

RT

SS 657 175

n/a

River Taw, Devon

Floodplain

Pasture

Each soil was texturally characterised and classified according to Hodgson (1974, in White, 1997) based on the percentage proportions of sand (defined as the 602000 µm fraction), silt (the 2-60 µm fraction) and clay (the IR) show the reverse of this and ties (R = IR) indicate particle sizes of approximately equal diameters. Wilcoxon Signed Rank Test for Paired-Samples Soil ID

CC

HH

LL

OF

OL

OP

RT

Sig

0.012

0.007

0.008

0.308

0.008

0.110

0.005

Negative Ranks (R < IR)

8

0

0

2

0

1

0

Positive Ranks (R > IR)

0

9

9

8

9

8

10

Ties (R = IR)

2

1

1

0

1

1

0

N= Sum of Ranks

R < IR R > IR

10

10

10

10

10

10

10

36.00

0.00

0.00

17.50

0.00

9.00

0.00

37.50

45.00

36.00

55.00

Not Significant

Significant

Not Significant

Significant

0.00

45.00

45.00

Significance

Significant

Significant

Significant

Direction

(IR > R)

(R > IR)

(R > IR)

(R > IR)

(R > IR)

Sig.: P 1 therefore indicate enrichment in comparison to the parent soil, whereas values < 1 indicate depletion. Enrichment ratios are presented graphically in Figure 7.14 and this is accompanied by a very brief summary describing the most notable differences or similarities between the seven soils. Between the particle size range of ca. 125-1 µm, the ER values of both inter-rill and rill eroded samples were found to be generally similar to each other and also to each parent material for most soils. The enrichment characteristics of soil OF were the exception to this generalisation, with both samples showing substantial depletion at the coarse fraction and increasing enrichment towards the finer fractions. Most notable for all soils however, was the apparent susceptibility of the coarse (i.e. 500-250 µm) fraction to enrichment or depletion by either erosion process, a factor that was also noted, albeit to a slightly lesser degree, for fine-sediment around the ca. 1-0.1 µm sizerange.

215


Figure 7.14 A comparison of enrichment ratios obtained from pairs of inter-rill and rill eroded sediment samples for seven different agricultural soils.

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7.5 Interpretation of Results The following section interprets the results of the investigation in the context of the original aim, which was to determine whether differences in the grain size composition of sediment mobilised by inter-rill and rill erosion from a range of different agricultural soils were statistically significant using the radionuclides, 60Co and 134Cs as tracers. Slope length and gradient, rainfall intensity, soil surface roughness, infiltration rate and the antecedent soil moisture content are among a suite of key environmental parameters held responsible for generating runoff and the mobilisation of sediment on tilled land (Seeger, 2007). Under the typical range of in-situ conditions, those variables interact to create a range of parameters under which sediment will often be mobilised (Chaplot & Le Bissonnais, 2000). For the purpose of the experiment performed on each of the seven soils however, attempts were made to hold each of the above variables constant in order to provide a similar set of eroding conditions. Measurements of the grain size composition of sediment samples mobilised by inter-rill and rill erosion, documented at 10 percentile intervals, revealed that statistically significant differences were obtained for a total of five soils. For four of those soils (i.e. HH, LL, OL & RT), sediment mobilised by rill erosion was found to be coarser than sediment mobilised by inter-rill erosion, and for one soil (i.e. CC), sediment mobilised by inter-rill erosion was found to be coarser than sediment mobilised by rill erosion. No significant difference in the grain size composition of pairs of inter-rill and rill eroded samples was recorded for two soils (i.e. OF & OP). The discussion will now focus on interpreting those results.

7.5.1 General Interpretations of the Results The statistically significant differences in the grain size composition of pairs of rill and inter-rill eroded sediment samples obtained from the four soils listed in Table 7.6 are interpreted as evidence of a lower sediment transport capacity during inter-rill erosion than for rill erosion. The higher available energy associated with the concentrated surface flow during rill erosion reportedly results in the transportation of larger material and these findings confirm the conclusion of previous workers (e.g. Alberts et al., 1980; Miller & Baharuddin, 1987; Slattery & Burt, 1995; Sutherland et al., 1996), all of whom concluded that inter-rill erosion is selective due to the lower energy of the unconcentrated surface flow. These findings have important implications with regard to the off-site conveyance of soil-sorbed nutrients and contaminants, since

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their transfer is closely associated with fine-sediment (Slattery & Burt, 1995; Stone & Walling, 1996; Walling et al., 2003). The result obtained from soil CC, which showed that inter-rill eroded sediment was significantly coarser than rill eroded sediment, supports the findings of Young & Onstad (1978) who conducted similar investigations on three different soil-types and reported that sediment mobilised by inter-rill erosion was enriched in sand-sized material and depleted in fine-sediment in comparison to rill eroded and parent soil. Although these findings tend to contradict the consensus reported in the majority of the relevant literature, it was suggested that the physical characteristics of those soils under their investigation may have favoured infiltration and the eluviation of fine material during erosion events, which, they predicted, might have left the soil surface and interrill eroded sediment depleted in fine material. For the two soils OF and OP that recorded no statistical significance between the grain size composition of sediment mobilised by either erosion process, these findings agree with Yang et al. (2006) who also reported little or no substantial variation in the grain size composition of sediment mobilised by either process. Reasons for this effect are tentatively attributed to the initially narrow grain size range of soils investigated at their study sites. They further speculate that a greater contrast in the particle size composition of the two erosion processes might be obtained if subsequent investigations are conducted on soils with an initially broader grain size composition. Based on this reasoning, attempts were made to determine whether a narrow grain size range might have been responsible for the lack of statistical distinction in the grain size composition of inter-rill and rill eroded sediment recorded by soils OF and OP. Table 7.6 lists the maximum particle size range of the eroded sediment samples for all seven soils and also the number of size intervals within each range. It also reconfirms whether the rill eroded sample was coarser, finer or similar to the inter-rill eroded sample. The broadest particle size range was 500-0.1 µm, was recorded by soil CC and consisted of 12 size intervals. In contrast, the minimum size range was 250-1 µm, consisted of 8 size intervals and was recorded by one soil (LL). The particle size range for both soils OF and OP was 250-0.25 µm and the number of size intervals within this range was 10. Since this score represents the median value of all seven soils, this implies that an initially narrow grain size range was not responsible for the lack of statistical differences recorded from those soils. It was noted, however, that very coarse (i.e. 500-250 µm) and very fine (0.25-0.1 µm) sediment was absent from these samples. Since material around those size fractions has been recognised for its susceptibility to 218


erosion by this and other investigations (cf. Young & Onstad, 1978; Young, 1980; Meyer & Harmon, 1984; Miller & Baharuddin, 1987; Stone & Walling, 1996; Sutherland et al., 1996; Ampontuah et al., 2006), this may provide an adequate reason for the lack of significant difference between the pairs of inter-rill and rill eroded samples for soils OF and OP. Table 7.6 A summary of the maximum particle size range of the runoff eroded sediment samples for each soil. Soil ID CC HH LL OF OL OP RT

Maximum Particle Size Range (µm) Fine

Number of Size Intervals

Significance

Coarse 500 500 250 250 500 250 500

0.1 1 1 0.25 1 0.25 0.25

12 9 8 10 9 10 11

IR > R R > IR R > IR R = IR R > IR R = IR R > IR

A possible reason for the lack of consensus between the grain size composition of pairs of inter-rill and rill eroded samples from the seven soils may relate to the magnitude of the sub-dominant erosion process at the time each pair of eroded samples were obtained during each erosion event. To explain this more fully, it has been demonstrated that the early stages of an erosion event are typically dominated by interrill erosion through a combination of splash detachment and detachment and transport by shallow surface runoff (Yang et al., 2006). Consequently, the magnitude and potential influence of rill eroded sediment on inter-rill eroded samples will typically be low. Information derived from this particular tracing approach allowed this effect to be quantified and data listed in Table 7.5 demonstrated that the proportion of rill eroded sediment incorporated within each of the inter-rill eroded samples was indeed very low (i.e. ≤ 5.3%) for all soils, and zero for the majority of soils. For sediment mobilised by rill erosion however, although information derived from the tracing technique indicated evidence of a marked increase in the magnitude and predominance of rill erosion during each of the events, particularly during the latter stages of each, the magnitude of inter-rill erosion often remained substantial, as data again listed in Table 7.5 reveal. Consequently, the relative proportions of inter-rill eroded sediment incorporated within some of the rill eroded samples was considerably larger than for the inter-rill eroded samples and ultimately necessitated omitting one soil (BB) from any further particle size analyses. For the two soils, OF and OP, that recorded no significant difference between the grain size composition of pairs of interrill and rill eroded samples, and for soil CC that recorded coarser inter-rill than rill 219


eroded sediment, the influence of inter-rill erosion may have been sufficient to reduce any contrasting differences between the grain size characteristics of the three pairs of samples. It is tentatively suggested that rates of erosion during each event for those three soils may have been transport limited (Chaplot & Le Bissonnais, 2000) and may relate to the particular characteristics of the three soils, combined with the relatively short length of the erosion plot (i.e. 1 m), the length of which may have been insufficient to permit adequate variation in the flow velocity during inter-rill and rill erosion to fully characterise the mobilised sediment (Chaplot & Le Bissonnais, 2000). During an investigation into the characteristics of eroded sediment from erosion plots of different lengths (i.e. 0.45-4.6 m), Young (1980) noted inconsistencies in the characteristics of sediment mobilised from short plots. This was attributed to the short length of the plots, limiting the exposure of the soil to the effects of sorting during its downslope conveyance. Based on this finding, it is feasible therefore that the length of the erosion plot used in this investigation might also have been too short. Limitations on the magnitude of rill erosion during some of the erosion events may therefore have prevented this process from fully characterising (i.e. sorting) the mobilised sediment, particularly in the case of soils OF and OP. The absence of any other distinguishing characteristics that might have enabled soils CC, OF and OP to be differentiated from the other five soils prevents a reason for the lack of consensus between the results from the other four soils from being established. However, since each was subjected to an initially similar suite of environmental parameters (i.e. slope gradient and length, antecedent soil moisture, infiltration rate, rainfall intensity, maximum aggregate size, surface texture, etc.), all of which were held as constant as possible during each rainfall simulation, it is suggested that differences in the sets of results may relate to specific soil properties such as the texture and initial grain size range of each, the OM content and the stability of WSAs, all of which have the ability to control mechanisms that influence rill incision (Young & Onstad, 1978; Alberts et al., 1980; Young, 1980; Yang et al., 2006).

7.6 Assessment of the Approach The assessment of the approach focuses on the design of the experiment and in particular to its ability to simulate mobilisation processes that occur in-situ on tilled slopes as a result of surface runoff. Attention is given to the tracing approach, the way in which the radionuclides were used to indicate changes in the mobilisation process and limitations in the sampling strategy. 220


7.6.1 Mass Balance: Uncertainty A simple mass balance described in Equation 1 was used to partition the radiometric data obtained from the two radionuclides associated with the runoff eroded sediment samples to estimate the relative proportions of sediment mobilised by inter-rill and rill erosion for each soil. Equation 2 was used to determine the recovery ratio for individual samples and the mean value and standard deviation, calculated from each data-set and listed in Table 7.3, were used to indicate the overall recovery of each radionuclide and the degree of variability in each data-set and thus determine the quality of each tracing experiment (Gölz, 2002). Mean recovery ratios for the eight soils ranged from 0.92 to 1.08. Five soils (i.e. BB, HH, LL, OF & RT) recorded mean values > 1.00 and the remaining three soils (CC, OL & OP) recorded mean values < 1.00. Errors (i.e. over or under estimations) associated with the radiometric data may have been obtained due to slight inaccuracies in the calculated efficiency of the detector owing to large variations in the mass and geometry of the sediment samples. Although attempts were made to account for these variations by calculating the efficiency of the detector using a range of calibration standards corresponding to the approximate mass of each runoff sample, more accurate results may have been obtained using additional calibration standards that increased in smaller mass increments (Miller, 1998). Furthermore, the quality of the radiometric data, and thus the accuracy of the partitioning procedure, could have been indirectly estimated at the end of each experiment by measuring the dimensions of each rill system. Combined with bulk density data for each soil, this would have provided an estimate of the volume of sediment removed from the rill system by rill erosion and values could then have been compared with the predicted removal of sediment indicated by the radionuclide measurements (Yang et al., 2006).

7.6.2 Limitations of the Sampling Strategy The sampling approach was based on selecting one pair of inter-rill and rill eroded samples from each soil. Each pair of sediment samples was then statistically analysed in order to determine whether differences in their grain size composition, measured at corresponding size intervals, were statistically significant at the 95% level of confidence. This strategy highlights two main limitations. The first of those arises from the fact that the results were based solely on information from just one pair of samples from each soil. Although this may have resulted in ignoring other potentially revealing information, this strategy was considered to be the most effective and efficient use of time and resources in order to achieve the aims of the investigation. The second 221


and arguably more fundamental limitation relates to the relative magnitude of the subdominant erosion process and its overall influence on the physical characteristics of the sediment mobilised by the predominant erosion process. Whilst data listed in Table 7.4 confirmed that the contribution, and hence the potential influence of rill eroded sediment incorporated within the inter-rill eroded samples was relatively low or zero, the proportion of inter-rill eroded sediment incorporated in the rill eroded samples was more substantial. Potential ways to overcome this problem in future investigations could be to increase the magnitude of rill erosion by, for instance, increasing the length of the erosion plot, increasing the rainfall intensity and extending the duration of each rainfall simulation in order to allow the rill system to incise and develop more fully. In studies where estimates of the magnitude of rill erosion have been reported, Yang et al. (2006) found that it accounted for 54.3% and 61.4% of the total eroded sediment yield from cultivated and uncultivated plots respectively, whereas Whiting et al. (2001) reported that rill erosion accounted for ca. 96% of the total eroded sediment yield from an area of pasture. All of those estimates are considerably greater than the mean 42.8% value (Table 7.3) obtained from the eight soils studied in this investigation. This suggests that restrictions in the development of the rill system during some of the experiments may have limited the magnitude of rill erosion, potentially curtailing its ability to fully characterise the eroded sediment during its downslope conveyance (Young, 1980). A possible reason for this may have been the relatively short duration of some of the rainfall simulations. Whilst extending the duration of each rainfall simulation may have mitigated this effect, the design of the erosion plot (Figure 7.1) ultimately limited the depth of rill incision to a maximum ca. 40 mm. Although this depth of incision was attained relatively quickly during some of the rainfall events, the restriction may still have ultimately been sufficient to limit the development of some of the rill systems.

7.6.3 The Erosion Plot: Dimensions, Scale & Edge Effects The benefits of using plots in soil erosion investigations have been well documented (e.g. Young & Onstad, 1978; Bowyer-Bower & Burt, 1989; Fox & Bryan, 1999; Legout et al., 2005). Major advantages include the convenience of being able to investigate different soils whilst manipulating key environmental parameters to accord with the requirements of the investigation. Major limitations associated with their use include their size, scale and representativeness (cf. Young, 1980; Parsons et al., 2006) and the influence of edge-effects (cf. Morgan, 1993) on the quality of the data. Factors arising from those limitations were considered at the design stage of the investigation 222


and attempts to mitigate their potential effects were implemented. It was considered that the most likely and recurrent problem would be the formation and / or migration of the rill system against the edge of the erosion plot. Measures to counter this potential scenario were taken and involved placing an aluminium plate with a top-edge which was profiled to form a shallow swale. The finished soil surface was then manipulated to match the profile of the plate to thus simulate a swale in an attempt to concentrate the surface flow in the centre of the plot and so encourage the rill to initially form and develop away from the edges. The two different rill systems shown in Figures 7.9 & 7.10 both demonstrate the apparent success of this approach. 7.6.4 Reduced Aggregate Stability The chosen method of labelling soils with radionuclide material necessitated subjecting each soil to a cycle of rapid wetting (slurrifying), followed by a gradual redrying. This process may have increased the susceptibility of each soil to the effects of slaking (Barthès & Roose, 2002; Ramos et al., 2003; Greene & Hairsine, 2004), thereby diminishing the stability of the aggregates and reducing their natural ability to resist the effects of detachment and erosion during rainfall events. It is thus recognised that the manipulation of the soils in the manner outlined earlier for the purpose of labelling with radionuclide material may have influenced their behaviour, thereby making it unrepresentative of in-situ soils subjected to similar erosional processes.

7.7 Conclusion The work described in this particular chapter has outlined the development of a preliminary method for documenting differences in the grain size composition of samples of sediment mobilised by inter-rill and rill erosion using the radioisotopes, 60Co and

134

Cs as tracers. The procedure was applied to eight different agricultural soils

obtained form a variety of hillslope and floodplain locations in the south of England. Within the constraints of the experimental procedures outlined above, and taking account of the difficulties associated with attempting to recreate natural processes under experimental conditions (Morgan, 1993), the results obtained from five of the eight soils confirm the influence of inter-rill and rill erosion on the physical properties of the eroded sediment. It is not entirely clear why the grain size composition of eroded sediment from two soils was not influenced by changes during the transition from interrill to rill erosion. It is speculated however, that this may relate to the potentially low magnitude of sediment conveyance during rill erosion, which is controlled by a 223


combination of the length of the erosion plot, the duration of the rainfall simulations, and to the physicochemical characteristics which control the aggregate stability of individual soils. Re-applying this tracing approach on a wider range of agricultural soils under an amended set of eroding conditions and with more detailed information on the initial physicochemical characteristics of each of the soils under investigation would therefore represent a focus for further work.

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Chapter 8: Summary & Conclusion

CHAPTER 8: SUMMARY & CONCLUSION

8.1 Chapter Content The first section of this chapter briefly describes the principal findings of each study and relates these to the specific aim (or aims) of each of the five investigations outlined in Chapters 3-7. In addition, it also highlights the relevant limitations of each application and suggests areas where further work could be undertaken, either to clarify the findings, and / or to improve or refine each tracing approach. General conclusions are presented in the second section. These summarise the generic advantages and disadvantages associated with using artificial radionuclides as sediment tracers that were encountered during the overall investigation. The final section presents the overall conclusions of the research and these are related to the main aims of the thesis introduced in Chapter 1.

8.2 Principal Findings, Limitations & Scope to Extend the Existing Work 8.2.1 Chapter 3 Chapter 3 described a tracing study aimed at determining the extent of remobilisation of recently-deposited sediment on river floodplains. The work was conducted on river floodplains located in two different catchments in Devon, UK. The findings from each site indicated that remobilisation of recently-deposited overbankderived sediment represented an important, but hitherto largely unrecognised and unquantified process. The two sets of results indicated that several physical factors controlled the extent of remobilisation at each site and these included the floodplain geometry, topography and surface roughness and the magnitude of the inundation events. The results from the River Culm site also indicated that the effect of wind-action across the surface of the floodwater may promote the re-suspension or remobilisation of recently-deposited sediment during low-magnitude flood events or in shallow floodwaters, and this may facilitate the adsorption / desorption of sediment-bound contaminants. Based on the two sets of findings, it was concluded that the remobilisation of recently-deposited sediment could exert an important influence on the rate at which fresh material is incorporated into the longer-term sediment-column. This finding could therefore have implications for estimating medium-term rates of sediment deposition on river floodplains based on measurements from fallout radionuclides 137Cs and excess

210

Pb. The estimates obtained from such investigations will reflect the net

225


deposition (i.e. total deposition less subsequent remobilisation) and may significantly underestimate the total amount of deposition. Whilst the specific aim of developing a method for documenting the remobilisation of recently deposited sediment on river floodplains during subsequent overbank flood events was successfully achieved, the results identified two key limitations to the approach which need to be addressed in its future development. The first and less problematic of the two was the progressive reduction of the analytical precision for subsequent inundation events at both sites. This effect was attributed to a combination of the diminishing radiometric signal emitted from the labelled sediment, and the relatively short counting time (i.e. 500 s) devoted to each in-situ radiometric assay. It is proposed that these effects could be mitigated in several ways. These include extending the counting time of each assay, substantially increasing the initial quantity of labelled sediment deposited at each plot and also increasing the initial activity concentration of the radionuclides. The second and more problematic limitation involved determining whether the reduction in each AAD value was caused by remobilisation of the labelled sediment, or reflected, at least in part, burial of the labelled sediment, and hence attenuation of the radiometric signal by fresh unlabelled material. As previously described in Chapter 3, a qualitative approach was suggested, whereby a coloured marker, such as powdered gypsum placed on a section of a plot, would provide a visual indication of the presence of freshly deposited sediment. A more complex quantitative field-based approach was also proposed, to complement the above method. This involved using two radionuclides deposited at the same location over two subsequent inundation events. Changes in the AAD value of each radionuclide after the subsequent flood event would indicate the relative importance of remobilisation and deposition. The findings of this second approach could also be further corroborated under laboratory conditions by quantifying the attenuating effects of a thin layer of unlabelled sediment deposited onto a plot of labelled material.

8.2.2 Chapter 4 Chapter 4 described a tracing application for determining both the translocation distance and vertical movement of soil within the till-zone during upslope tillage on a 20% slope. The work was conducted using unique experimental tillage erosion facilities located at the U of M in Canada and exploited the ability of this equipment to maintain both the depth and speed of tillage as constant values. The principal findings indicated that the quantity of mobilised soil was compatible with the findings of other

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investigations with comparable tillage parameters. In terms of the vertical movement of soil within the till-zone, a substantial net downward movement was recorded, indicating uneven mixing of the soil within the till-zone during tillage. This finding has potential implications in agricultural areas where the conversion from mould-board to chisel-plough tillage forms part of a conservation tillage regime (Carter et al., 1998) and could promote the uneven distribution of soil amendments, thereby reducing their effectiveness, and / or result in the accumulation of sediment-bound contaminants within the till-zone. The specific aims of quantifying the soil translocation distance during upslope tillage at a constant speed and depth, and determining the extent of mixing within the till-zone were successfully achieved. However, a number of fundamental limitations were identified which would need to be addressed and overcome in the future development of the approach. The results from the tillage translocation component of the investigation generally concur with the findings from previous work using physical and other forms of tracers. This therefore usefully confirms the ability of those existing techniques to provide meaningful results. Given this particular conclusion, it is suggested that further work of a similar nature could be successfully undertaken using cheaper and more convenient types of tracer. In terms of determining the vertical mixing of soil within the till-zone, the stratified tracing approach represented an improvement on techniques that employed physical tracers, since the movement of the radionuclides mimicked the soil and was not influenced by their own physical characteristics (cf. Rahman et al., 2005; Mohler et al., 2006; Spokas et al., 2007). Based on this conclusion, the most logical and beneficial amendment to this component of the investigation would be to increase the number of different radionuclides within the soil profile in order to provide more detailed information on patterns of soil mixing. In terms of the experimental approach, the high-resolution sampling strategy employed represented the biggest limitation of the investigation, since it resulted in the generation of a prohibitively large quantity of samples (i.e. ~ 2000), each of which required sectioning, processing and analysis by gamma spectrometry. This problem was further exacerbated by an inadvertent administrative oversight associated with obtaining the necessary permits to legally permit the experiment to be undertaken using radioactive material above a certain activity threshold (i.e. 100 kBq) in an enclosed space. Since the threshold had unknowingly been exceeded during the experiment, the oversight was brought to the attention of the Canadian Nuclear Safety Commission (CNSC) who immediately secured the experimental facilities and prevented personnel

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from accessing the site until a thorough review of all aspects of the research had been completed. This necessarily halted the removal, processing and analysis of samples and ultimately resulted in the loss of all of the ‘detector time’ originally allocated to the experiment. Collectively, the significant limitations outlined above firstly clearly demonstrate the large resource input required to undertake work of this nature, and secondly, they also highlight the contentiousness associated with radioactive material per se, and especially with its deliberate introduction or release into any environment.

8.2.3 Chapter 5 Chapter 5 described a tracing application for determining the extent of sediment redistribution on livestock-poached pasture. The work was conducted on a hillslope in mid-Devon and the principal findings demonstrated that all areas of livestock-poached pasture that were monitored experienced sediment redistribution (i.e. erosion and deposition). A net loss of sediment was recorded for all plots and this demonstrated that erosion was the dominant process. A very strong and direct correlation was recorded between the mean depth of erosion for some plots and the cumulative depth of rainfall between re-measurements. In contrast, no correlation was found between the mean depth of erosion and the gradient of each plot. These results are interpreted as evidence that surface hydrology (i.e. the occurrence of rainsplash impact and surface runoff) is substantially more important in influencing the erosion and redistribution of soil and sediment on livestock-poached pasture and than the slope gradient. Based on these overall findings, it was concluded that all areas devoid or partially devoid of vegetation are vulnerable to sediment redistribution, and particularly to erosion. These findings would also suggest that this process is not necessarily exclusively linked to livestock poaching, but can increase the vulnerability of any pasture areas with sparse or reduced vegetation cover. Whilst the specific aim of the investigation to assess the extent of sediment redistribution from areas of livestock-poached pasture over a series of rainfall events was successfully achieved, a small limitation was identified in the approach which would need to be addressed in its future development. This primarily related to the limited number of plots used to explore the relationships between erosion rates and rainfall, and erosion rates and gradient. Since this particular sediment mobilisation and conveyance mechanism is probably influenced by more than the three variables that were tested, a more extensive data-set derived from a more detailed network of plots located on a wider range of gradients, and perhaps also at different hillslope sites with

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contrasting rainfall characteristics, may have provided a better basis for interpreting the results provided by this tracer application.

8.2.4 Chapter 6 Chapter 6 described a tracing application for determining the fate of earthworm casts mobilised by erosion from pasture and cultivated soils. The work on pasture was conducted in-situ and relied on natural weather events as the eroding agent, whereas the work on cultivated soil was undertaken under laboratory-conditions and used simulated rainfall as the eroding agent. The principal results obtained from the experiment conducted on pasture indicated that substantial quantities of cast-derived sediment were mobilised by surface runoff and the magnitude of sediment removal was attributed to the rate of cast production, the availability of material, and to the timing of rainfall events. With regard to differences between the relative proportions of recovered sediment from the two batches of casts, substantially more was recovered from those located nearest to the outlet, whereas less was recovered from the batch located further away. The preliminary conclusion was that longer travel distances provide more opportunity for the re-incorporation of the mobilised sediment back into the surrounding soil between erosion events. However, the results also demonstrated that dispersed material from both batches of casts continued to represent a significant source of sediment throughout the duration of the experiment. For the experiment conducted on cultivated soil, the principal results demonstrated that the total sediment yield from the unconsolidated soil during the first rainfall event was notably higher than from the crusted soil during the second event, and only a relatively small proportion of sediment from the

134

Cs-labelled casts was

removed. In contrast, the overall sediment yield from the crusted surface during the second event was considerably less than the first, but included a larger proportion of sediment from the 60Co-labelled casts. The results also demonstrated that material from both the bulk soil and from the

134

Cs-labelled casts still continued to represent a

substantial source of sediment despite the surface crust. In comparing differences in the quantities of cast-derived material removed by surface runoff during both erosion events, in approximate terms, double the amount of

60

Co-labelled sediment was

removed in half the rainfall duration. Whilst the specific aims of the investigation, which were to develop a method of labelling earthworm casts with 134Cs and 60Co and to then apply the method in field and laboratory studies, were both successfully achieved, the results identified certain

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limitations in the experimental design and in aspects of the tracing application which would need to be addressed and overcome in its future development. In terms of experimental design, more emphasis should be placed during future work on determining the travel distance of the eroded cast material. This would include conducting work on plots of different lengths in order to establish whether a relationship exists between the magnitude of erosion / recovery and travel distance. Coupled to this, more emphasis would be placed on examining spatial variations which control the reincorporation of sediment back into the surrounding soil on both land-use types. With regard to cultivated soil in particular, additional work using multiple tracing techniques may provide information on the depth of sediment in-washing before and after the formation of the crust. Such work would provide a means of confirming or refuting the inferred yet unverified theory that sediment from dispersed casts may contribute to the formation of surface crusts (Shipitalo et al., 1988). In this application, the success of the whole tracing application was underpinned by the assumption that earthworm casts of differing mass and density could be uniformly labelled using a technique of immersion into solutions containing radionuclides. Owing to probable variations of the rate of penetration of the liquid into casts of differing size and mass, along with the high affinity of the radionuclides to sediment, it was possible that sediment at the surface of the casts had higher activities per unit mass than material closer to the centre of the casts. The efficacy of this labelling technique thus represented a potential source of uncertainty that would need to explored and addressed during future work. This could be accomplished by repeating the actual labelling procedure described in Chapter 6 numerous times in order to compare the results obtained by this procedure.

8.2.5 Chapter 7 Chapter 7 described a stratified tracing application for documenting sediment source changes during the transition from inter-rill to rill erosion. The approach was applied to eight different agricultural soils collected from around the south and southwest of England and the principal findings demonstrated that the particle size distribution of rill eroded sediment, when measured at 10 percentile intervals, was significantly coarser than the grain size composition of inter-rill eroded sediment for four soils, whereas for one soil, inter-rill eroded sediment was significantly coarser than rill eroded sediment. The results collectively provide empirical evidence in support of the argument that inter-rill and rill erosion are characterised by different hydraulic

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conditions and this will typically be reflected in the grain size composition of the eroded material. These findings have important geochemical and geomorphological implications with respect to hillslope processes when investigating the mobilisation and transfer of fine-sediment and sediment-associated nutrients and contaminants (Yang et al., 2005), for estimating long-term landscape evolution (Young & Onstad, 1978; Alberts et al., 1980; Farenhorst & Bryan, 1995), and for modelling the movement of soil by surface runoff (Le Bissonnais et al., 1995, 2005; Yang et al., 2006). No significant difference in the grain size composition of sediment removed by either interrill or rill erosion was documented for two soils, and one soil was omitted from statistical analysis due to the high proportion of inter-rill eroded sediment incorporated with the rill eroded sample. Possible reasons for the conflicting sets of results are summarised over the following paragraphs. The design of the experiment involved maintaining key environmental parameters (i.e. maximum aggregate size, slope, surface texture, antecedent soil moisture content and rainfall intensity) constant during each experiment and this factor potentially contributed to narrowing the range of conditions explored, which could reduce the variability associated with the different sets of results. The variables under consideration related specifically to the properties of the soils and included differences in their physical characteristics and in particular, their initial grain size range (cf. Yang et al., 2006). With hindsight, closer attention should also have been given to other fundamental soil properties, such as the nature and relative stability of the WSAs within each of the soils under investigation. In addition, generic limitations associated with the overall design of the experiment were also considered. These included using an erosion plot of sufficient length, which some workers have argued is necessary to enable rill erosion to be simulated correctly (e.g. Young, 1980; Chaplot & Le Bissonnais, 2000). These factors may explain the lack of statistical distinction between the grain size composition of pairs of inter-rill and rill eroded samples for two soils, and for the high level of inter-rill eroded sediment within the rill eroded sample recorded from soil BB. Whilst the specific aim of the investigation to document differences in the grain size composition of pairs of inter-rill and rill eroded sediment samples from a selection of agricultural soils was successfully achieved, the different sets of results and the interpretation of those results highlighted a number of limitations of the work undertaken, both in the design of the erosion plot and in the actual tracing application that would need to be addressed and overcome in its future development. In terms of

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design adjustments, as indicated above, this would involve conducting experiments on longer, and potentially deeper, erosion plots in order to ensure that sufficient opportunity be given for the incision and development of the rill system to ensure that the rill erosion processes were correctly simulated. In terms of the limitations of the investigation, consideration was also given to the effectiveness of the tracing approach and the possibility that identical results might have been obtained by simpler means (i.e. by observing the transition from inter-rill to rill erosion during each experiment and then selecting pairs of samples accordingly). This approach would firstly be dependent on qualitative estimates of the relative contribution of sediment derived from the predominant erosion process and secondly, would be unable to estimate the relative contribution of sediment derived from the sub-dominant erosion process. By contrast, the chosen approach provided the most effective means of quantitatively estimating both of the above factors, and therefore provided a systematic approach to the selection of samples most characteristic of inter-rill and rill eroded sediment for statistical comparison.

8.3 General Conclusions This section of the chapter presents a series of general conclusions relating specifically to the use of

134

Cs and

60

Co as sediment tracers. They are based on the

experience gained in designing, implementing and undertaking the five investigations and each has been categorised and amplified under the sub-headings of Advantages or Disadvantages. Although many of these conclusions may be applicable to other radionuclides or other forms of tracer, it is stressed that the comments presented here relate specifically to 134Cs and 60Co.

8.3.1 Advantages Based on information gained from the five investigations, the main advantages of using the two radionuclides over other types of sediment tracer reflect their particular characteristics and attributes. These include their rapid and strong affinity for finesediment, their availability and relatively low cost, their relatively short half-life duration and the negligible mass, high energy and penetrative nature of the ejected gamma-photons. These fundamental characteristics all contributed to permit the rapid, easy and accurate introduction of the radionuclides into a diverse range of environments or a diverse range of erosion scenarios. Those same characteristics then facilitated the re-identification and re-measurement of the radionuclides under in- or ex-situ conditions

232


using sensitive analytical equipment that allowed subtle changes in radiometric values to be accurately estimated. Both radionuclides afforded a substantial level of convenience which permitted experiments to be repeated at different locations. This allowed the relative influence of varying environmental parameters, including different land-use and / or different soil-types, on rates, patterns and characteristics of erosion processes, to be determined at representative locations on or around particular landscape features, or on landforms with similar or contrasting topographic characteristics. Within the framework of the five soil erosion investigations undertaken, the numerous

positive

attributes

summarised

above

were

fully

exploited

and

comprehensively tested. The findings produced a package of essentially unique information which was relatively easily collated and which could be interpreted to provide accurate and reliable estimates of the movement or redistribution of small quantities of fine-sediment over the range of spatial and temporal scales at which the experiments were performed.

8.3.2 Disadvantages During the course of designing and implementing the investigations, some significant disadvantages were encountered. Among the most prominent were locating a (legitimate) supplier of radioactive material, obtaining and ensuring that the necessary licences were in-place to enable the material to be purchased, stored, used and disposed of, the need for personnel (i.e. a radiation officer) trained to partition and dispense the material at suitable activity concentrations, as well as virtually unrestricted use of specialised equipment (i.e. gamma spectrometers) which allowed in-situ and laboratorybased analyses to be undertaken. For the field-based investigations, finding landowners or farm-managers that were willing to permit work to be conducted on their land was initially difficult owing to the stigma associated with radioactivity, the deliberate release of radioactive material and the potential contamination of portions of their land as a result of undertaking the investigative work. Numerous disadvantages were also noted during the preparatory work to make individual investigations operational. Among the most prominent of those was the need to subject many of the soils under investigation to repeat cycles of rapid drying, wetting / slurrification and re-drying during the labelling procedure, and the need for sieving to remove foreign debris and to obtain soil aggregates of a certain size-range. It has thus been repeatedly recognised throughout many of the investigations that this procedure may have unduly perturbed the natural aggregate stability of the soils, thereby making them overly vulnerable to the effects of

233


slaking or altering their general ability to resist erosion. The direct labelling of soils with radionuclide material also inevitably resulted in some spatial variation of initial inventory values. Since rates of soil redistribution were based on pre- and post-event inventories, reliable information on pre-event inventories was paramount. Although this problem was relatively easily overcome by determining the initial inventories before an erosion event, this approach substantially increased the duration of each total measurement campaign.

8.4 New Directions & Potential Research Areas Given the encouraging results described in Chapters 3-7 and summarised in earlier sections of this chapter, scope potentially exists to expand the range of 134Cs and 60

Co applications in new directions and into other research areas. Granger et al. (2007)

offer an example of a potentially suitable theme by calling for the development of novel tracing techniques to enable the nature of runoff generated from intensively-managed grasslands to be more closely examined; with the ultimate aim of identifying and apportioning sources of sediment, colloids and associated pollutants to soil processes and / or agricultural amendments. A review of some of the more promising techniques was undertaken by Granger et al. (2007) and each was judged based on their advantages and disadvantages. Although, surprisingly, the review does not include the use of artificial radionuclides, the evidence associated with the behaviour of both 60

134

Cs and

Co and presented in this thesis would indicate their ability to provide information at

the required spatial and temporal scales from the aforementioned environments owing to their attributes and environmental characteristics, some of which may also allow many of the disadvantages and limitations associated with some of the techniques reviewed by Granger et al. (2007) to be circumvented. Although no individual research projects have been formulated, the limited capabilities of existing techniques in this area of research could represent an ideal opportunity to challenge and further test the tracing capabilities of 134Cs and 60Co in new directions however, and so extend their repertoire as sediment tracers.

8.5 Conclusion In the context of the original aims, five novel tracing applications using the artificial radionuclides,

134

Cs and

60

Co were designed and successfully implemented.

The findings from the five investigations confirmed the efficacy of the two radionuclides as tracers of fine-sediment and each set of results has contributed to an

234


improved understanding of the mobilisation and transfer mechanism under investigation. It should be emphasised however that each of the applications represents a preliminary approach and further work will inevitably be needed to clarify the findings, or to improve or refine each of the techniques. From the perspective of the relatively few, yet nevertheless significant, disadvantages identified in the previous section, it is acknowledged that alternative forms of tracer or tracing technique may represent a cheaper, easier and more convenient option in many instances. Nonetheless, under certain erosion scenarios where other types of tracer or tracing technique would prove to be ineffective or lack the required sensitivity, the viability of both 60

134

Cs and

Co has been confirmed. Although this work has expanded the scope of erosion

scenarios that can currently be investigated, additional work is required to fully explore, further develop and confirm their capabilities as tracers in order to enable them to be added to the suite of existing tracers and tracing techniques currently available to researchers.

235

REFERENCES

Aalto, R. & Dietrich, W. (2005) “Sediment accumulation determined with

210

Pb

geochronology for Strickland River flood plains, Papua New Guinea”. International Association of Hydrological Sciences (IAHS) Publication, 291, pp. 1-7.

Aalto, R., Dunne, T, Nittrouer, C. A., Maurice-Bourgoin, L. & Montgomery, D. R. (2002) “Fluvial transport of sediment across a pristine tropical foreland basin: channelflood plain interaction and episodic flood plain deposition”. In. The Structure, Function and Management Implications of Fluvial Sedimentary Systems. International Association of Hydrological Sciences (IAHS), Publication No. 276, pp. 339-344.

Aalto, R., Lauer, J. W. & Dietrich, W. E. (2008) “Spatial and temporal dynamics of sediment accumulation and exchange along Strickland River floodplains (Papua New Guinea) over decadal-to-centennial timescales”. Journal of Geophysical Research, 113, pp. 1-22.

Aase, J. K. & Pikul, J. L (1995) “Terrace formation in cropping strips protected by tall wheatgrass barriers”. Journal of Soil and Water Conservation, 50, pp. 110-112.

Abrahams, A. D., Li, G., Krishnan, C. & Atkinson, J. F. (1998) “Predicting sediment transport by interrill overland flow on rough surfaces”. Earth Surface Processes and Landforms, 23, pp. 1087-1099.

Ahmadi, S. H., Amin, S., Keshavarzi, A. R. & Mirzamostafa, N. (2006) “Simulating watershed outlet sediment concentration using the ANSWERS Model by applying two sediment transport capacity equations”. Biosystems Engineering, 94, pp. 615-626.

Alam, S. M., Ansari, R. & Khan, M. A. (2001) “Application of radioisotopes and radiation in the field of agriculture”. Journal of Biological Sciences, 1, pp. 82-86.

Alberts, E. E., Moldenhauer, W. C. & Foster, G. R. (1980) “Soil aggregates and primary particles transported in rill and interrill flow. Soil Science Society of America Journal, 44, pp. 590-595.

236

Allan, J. C., Hart, R. & Tranquili, J. V. (2006) “The use of passive integrated transponder (PIT) tags to trace cobble transport in a mixed sand-and-gravel beach on the high-energy Oregon coast, USA”. Marine Geology, 232, pp. 63-86.

Alldredge, A. W. & Whicker, F. W. (1972) “A method for measuring soil erosion and deposition with beta particle attenuation”. Journal of Range Management, Vol. 25, pp. 393-398.

Amador, J. A. & Görres, J. H. (2005) “Role of the anecic earthworm Lumbricus terrestris L. in the distribution of plant residue nitrogen in a corn (Zea mays)- soil system”. Applied Soil Ecology, 30, pp. 203-214.

Ampontuah, E. O., Robinson, J. S. & Nortcliff, S. (2006) “Assessment of soil particle redistribution on two contrasting cultivated hillslopes”. Geoderma, 132, pp. 324-343.

Andrási, A., Németh, I. & Zombori, P. (1987) “Application of Ge-spectrometry for rapid in-situ determination of environmental radioactivity”. Central Research Institute for Physics, Hungary.

Anonymous. (1990) “EG&G ORTEC Solid State Gamma Spectrometry – Operators Manual”. EG&G ORTEC. USA.

Anonymous. (2003) “Radiation Safety Orientation Guide”. Issued by: Environmental Health & Safety Office, University of Manitoba, Winnipeg, Canada.

Anonymous. (2005) “Laboratory Manual & Methods, Physical Laboratories, Department of Geography”, University of Exeter, UK.

Avery, S. V. (1996) “Fate of caesium in the environment: Distribution between the abiotic and biotic components of aquatic and terrestrial ecosystems”. Journal of Environmental Radioactivity, 30, pp. 139-171.

Bailly du Bois, P. & Guéguéniat, P. (1999) “Quantitative assessment of dissolved radiotracers in the English Channel: Sources, average impact of la Hague reprocessing

237

plant and conservative behaviour (1983, 1986, 1988, 1994)”. Continental Shelf Research, 19, pp. 1977-2002.

Bailly du Bois, P. (1996) “Mapping of water masses in the North Sea using radioactive tracers”. Endeavour, Vol. 20, pp. 2-7.

Bandeira, J. V., Sabino, C. S., Aun, P. E., Mendes, V. L. & Garcia Agudo, E. (2002) “Development of a technique for using 99mTc as an adsorbable tracer for hydrodynamic studies of fine sediment in suspension”. Applied Radiation and Isotopes, 57, pp. 85-92.

Banner, J. L. (2004) “Radiogenic isotopes: systematics and applications to earth surface processes and chemical stratigraphy”. Earth-Science Reviews, 65, pp. 141-194.

Barling, R. D. & Moore, I. D. (1994) “Role of buffer strips in management of waterways: A review”. Environmental Management, Vol. 18, pp. 543-558.

Barrow, C. J. (2000) Environmental Management: Principles and Practise. Routledge: London.

Barthès, B. & Roose, E. (2002) “Aggregate stability as an indicator of soil susceptibility to runoff; validation at several levels”. Catena, 47, pp. 133-149.

Basher, L. R. & Ross, C. W. (2001) “Role of wheel tracks in runoff generation and erosion under vegetable production on a clay loam soil at Pukekohe, New Zealand”. Soil & Tillage Research, 62, pp. 117-130.

Basic, F., Kisic, I., Mesic, M. & Butorac, A. (2002) “Particle size distribution (texture) of eroded soil material”. Journal of Agronomy and Crop Science, 188, pp. 311-322.

Bennett, S. J. (1999) “Effect of slope on the growth and migration of headcuts in rills”. Geomorphology, 30, pp. 273-290.

Beuselinck, L., Govers, G., Hairsine, P. B., Sander, G. C. & Breynaert, M. (2002) “The influence of rainfall on sediment transport by overland flow over areas of new deposition”. Journal of Hydrology, 257, pp. 145-163. 238

Billota, G., Brazier, R. E., Haygarth, P. M., Macleod, C. J. A., Butler, P., Granger, S., Krueger, T., Freer, J. & Quinton, J. (2008) “Rethinking the contribution of drained and un-drained grasslands to sediment-related water quality problems”. Journal of Environmental Quality, 37. pp. 906-914.

Bilotta, G., Brazier, R. E. & Haygarth, P. (2007a) “Processes affecting transfer of sediment and colloids, with associated phosphorous, from intensively farmed grasslands, erosion”. Hydrological Processes, 21, pp. 135-139.

Bilotta, G. S., Brazier, R. E. & Haygarth, P. M. (2007b) “The impacts of grazing animals on the quality of soils, vegetation and surface waters in intensively managed grasslands”. Advances in Agronomy, 94, pp. 237-280.

Bilotta, G. S., Brazier, R. E., Haygarth, P. M., MacLeod, C. J. A., Butler, P., Granger, S., Krueger, T., Freer, J. & Quinton, J. (2008) “Rethinking the contribution of drained and undrained grasslands to sediment-related water quality problems”. Journal of Environmental Quality, 37, pp. 906-914.

Binet, F. & Le Bayon, R. C. (1999) “Space-time dynamics in situ of earthworm casts under temperate cultivated soils”. Soil Biology and Biochemistry, 31, pp. 85-93.

Blake, W. H., Walling, D. E. & He, Q. (1999) “Fallout beryllium-7 as a tracer in soil erosion investigations” Applied Radiation and Isotopes, 51, pp. 599-605.

Blake, W. H., Walling, D. E. & He, Q. (2002) “Using cosmogenic Beryllium-7 as a tracer in sediment budget investigations”. Geografiska Annaler 84 A, pp. 89-102.

Blanchart, E., Albrecht, A., Brown, G., Decaens, T., Duboisset, A., Lavelle, P., Mariani, L. & Roose, E. (2004) “Effects of tropical endogeic earthworms on soil erosion”. Agriculture, Ecosystems and Environment, 104, pp. 305-315.

Boix-Fayos C., Martinez-Mena, M., Arnau-Rosalén, E., Calvo-Cases, A., Castillo, V. & Albaladejo, J. (2006) “Measuring soil erosion by field plots: Understanding the sources of variation”. Earth-Science Reviews, 78, pp. 267-285. 239

Børretzen, P. & Salbu, B. (2000) “Estimation of apparent rate coefficients for radionuclides interacting with marine sediments from Novaya Zemlya”. The Science of the Total Environment, 262, pp. 91-102.

Bottrill, L. J., Walling, D. E. & Leeks, G. J. (1999) “Geochemical characteristics of overbank deposits and their potential for determining suspended sediment provenance; an example from the River Severn, UK”. In S. B. Marriott & J. Alexander (eds.), Floodplains: Interdisciplinary Approaches. Geological Society Special Publication No 163, pp. 241-257. Geological Society: London.

Bourgoin, L. M., Bonnet, M. P., Martinez, J. M., Kosuth, P., Cochonneau, G., Moreira-Turcq, P., Guyot, J. L. Vauchel, P., Filizola, N. & Seyler, P. (2007) “Temporal dynamics of water and sediment exchanges between the Curuai floodplain and the Amazon River, Brazil”. Journal of Hydrology, 335, pp. 140-156.

Bowyer-Bower, T. A. S. & Burt, T. P. (1989) “Rainfall simulators for investigating soil response to rainfall”. Soil Technology, 2, pp. 1-16.

Bradford, J. M., Remley, P. A., Ferris, J. E. & Santini, J. B. (1986) “Effect of soil surface sealing on splash from a single waterdrop”. Soil Science Society of America Journal, 50, pp. 1547-1552. Brambilla, M., Fortunati, P. & Carini, F. (2002) “Foliar and root uptake of 134Cs, 85Sr and

65

Zn in processing tomato plants (Lycopersicon esculentum Mill.)”. Journal of

Environmental Radioactivity, 60, pp. 351-363.

Brazier, R. E. (2004) “Quantifying soil erosion by water in the UK: a review of monitoring and modelling approaches”. Progress in Physical Geography, 28, pp. 340365.

Brazier, R. E., Bilotta, G. S. & Haygarth, P. M. (2007) “A perspective on the role of lowland, agricultural grasslands in contributing to erosion and water quality problems in the UK”. Earth Surface Processes and Landforms, 32, pp. 964-967.

240

Brown, L. R. (1984) “Conserving soils”. In: State of the World, 1984 (ed. L. R. Brown), 53, pp. 53-75. Norton, New York.

Brunton, D. A. & Bryan, R. B. (2000) “Rill network development and sediment budgets”. Earth Surface Processes and Landforms, 25, pp. 783-800.

Brussard, L. (1997) “Biodiversity and ecosystem functioning in soil”. Ambio, 26, pp. 563-570.

Bryant, H. T., Blaser, R. E., and Peterson, J. R. (1972). “Effect of trampling by cattle on bluegrass yield and soil compaction of a Meadowville loam”. Journal of Agronomy, 64, pp. 331–334.

Buringh, P. (1981) “An assessment of losses and degradation of productive agricultural land in the world”. Food and Agriculture Organisation, Working Group on Soils Policy, Rome.

Burt, T. P. & Haycock, N. E. (2000) “Linking hillslopes to floodplains”. In Tracers in Geomorphology (ed. I. D. L. Foster), Chpt. 14, pp. 461-492. John Wiley & Sons Ltd. Chichester, UK.

Campbell, B. L. Loughran, R. J. & Elliott, G. L. (1988) “A method for determining sediment budgets using caesium-137”. In: Sediment Budgets. International Association of Hydrological Sciences (IAHS), Publication No. 174, pp. 171-179.

Capowiez, Y., Renault, P. & Belzunces, L. (2001) “Three-dimensional trajectories of 60

Co-labelled earthworms in artificial cores of soil”. European Journal of Soil Science,

52, pp. 365-375.

Carmi, G. & Berliner, P. (2008) “The effect of soil crust on the generation of runoff on small plots in an arid environment”. Catena, 74, pp. 37-42. Caron, F. & Mankarios, G. (2004) “Pre-assessment of the speciation of 60Co, 137

Cs and

241

125

Sb,

Am in a contaminated aquifer”. Journal of Environmental Radioactivity,

77, pp. 29-46. 241

Carter, M. R. & Sanderson, J. B. (2001) “Influence of conservation tillage and rotation length on potato productivity, tuber disease and soil quality parameters on a fine sandy loam in eastern Canada”. Soil & Tillage Research, 63, pp. 1-13.

Carter, M. R., Sanderson, J. B. & MacLeod, J. A. (1998) “Influence of time of tillage on soil physical attributes in potato rotations in Prince Edward Island”. Soil & Tillage Research, 49, pp. 127-137.

Chaplot, V, & Le Bissonnais, Y. (2000) “Field measurements of interrill erosion under different slopes and plot sizes”. Earth Surface Processes and Landforms, 25, pp. 145153.

Chen, L. & Lu, S. (2008) “Sorption and desorption of radiocobalt on montmorillinite – effects of pH, ionic strength and fulvic acid”. Applied Radiation and Isotopes, 66, pp. 288-294.

Chen, Y., Tarchitzky, J., Brouwer, J., Morin, J. & Banin, A. (1980) “Scanning electron microscope observations on soil crusts and their formation”. Soil Science, 130, pp. 49-55.

Colbach, N., Roger-Estrade, J., Chauvel, B. & Caneill, J. (2000) “Modelling vertical and lateral seed bank movements during mouldboard ploughing”. European Journal of Agronomy, 13. pp. 111-124.

Collins, A. L. & Walling, D. E. (2002) “Selecting fingerprint properties for discriminating potential suspended sediment sources in river basins”. Journal of Hydrology, 261, pp. 218-244.

Collins, A. L. & Walling, D. E. (2004) “Documenting catchment suspended sediment sources: problems, approaches and prospects”. Progress in Physical Geography, 28, pp. 159-196.

242

Collins, A. L., Walling, D. E. & Leeks, G. J. L. (1997) “Source type ascription for fluvial suspended sediment based on a quantitative composite fingerprinting technique”. Catena, 29, pp. 1-27.

Collins, A. L., Walling, D. E. & Leeks, G. J. L. (1998) “Use of composite fingerprints to determine the provenance of the contemporary suspended sediment load transported by rivers”. Earth Surface Processes and Landforms, 23, 31-52.

Connor, K. J., McLintock, I. S., Heaton, B. & Moseley T. J. (2007) “Radiation Protection: Handbook for Laboratory Workers” (ed. T. J. Moseley), Association of University Radiation Protection Officers. University of Sheffield, Sheffield.

Courtois, G. (1973) “The use of artificial radioactive tracers in France”. In. Tracer Techniques in Sediment Transport. Technical Report Series No. 145, pp. 71-80. International Atomic Energy Agency (IAEA). Vienna, Austria.

Coutts, J. R. H., Kandil, M. F., Nowland, J. L. & Tinsley, J. (1968) “Use of radioactive Fe for tracing soil particle movement”. Journal of Soil Science, 19, pp. 311341.

Cózar, A., Gálvez, J. A., Hull, V., García, C. M. & Loiselle, S. A. (2005) “Sediment resuspension by wind in a shallow lake of Esteros del Iberá (Argentina): a model based on turbidimetry”. Ecological Modelling, 186, pp. 63-76.

Cremers, A., Elsen, A., De Preter, P. & Maes, A. (1988) “Quantitative analysis of radiocaesium retention in soils”. Nature, 335, pp. 247-249.

Cundy, A. B. & Croudace, I. W. (1999) “Sediment accretion and recent sea-level rise in the Solent, Southern England: Inferences from radiometric and geochemical studies”. Estuarine, Coastal and Shelf Science, 43, pp. 449-467.

da Silva, J. M. & Alexandre, C. (2004a) “Soil carbonation processes as evidence of tillage-induced erosion”. Soil & Tillage Research, 78, pp. 217-224.

243

da Silva, J. M., Soares, J. M., C., N. & Karlen, D. L. (2004b) “Implement and soil condition effects on tillage-induced erosion”. Soil & Tillage Research, 78, pp. 207-216

Dabney, S. M., Liu, Z., Lane, M., Douglas, J., Zhu, J. & Flanagan, D. C. (1999) “Landscape benching from tillage erosion between grass hedges”. Soil & Tillage Research, 51, pp. 219-231. Dalgleish, H. Y. & Foster, I. D. L. (1996) “137Cs losses from a loamy surface water gleyed soil (Inceptisol); a laboratory simulation experiment”. Catena, 26, pp. 227-245.

Darwin, C. (1897) “The Formation of Vegetable Mould, Through the Action of Worms, with Observations on their Habits”. John Murray: London.

De Alba, S., Borselli, L., Torri, D., Pellegrini, S. & Bazzoffi, P. (2006). “Assessment of tillage erosion by mouldboard plough in Tuscany (Italy)”. Soil & Tillage Research, 85, pp. 123-142.

DEFRA (Department for Environment, Food and Rural Affaires) (2004) “The First Soil Action Plan for England: 2004-2006”. DEFRA Publication, London.

De Lima, J. L. M. P., Van Dijck, P. M., Spaan, W. P. (1992) “Splash – saltation transport under wind-driven rain”. Soil Technology, 5, pp. 151–166.

Dietrich, W. E. & Dunne, T. (1978) “Sediment budget for a small catchment in mountainous terrain”. Zeitschrift fur Geomorpologie, Supplement Band 29, pp. 191206.

Dincauze, D. F. (2000) “Environmental Archaeology: Principles and Practise”. Cambridge University Press, UK. Doering, C., Akber, R. & Heijnis, H. (2006) “Vertical distribution of 210Pb excess, 7Be and

137

Cs in selected grass covered soils in Southeast Queensland, Australia”. Journal

of Environmental Radioactivity, 87, pp. 135-147.

244

Douglas, R. W. & Rippey, B. (2000) “The random redistribution of sediment by wind in a lake”. Journal of Limnological Oceanography, 45, pp. 686-694.

Duchemin, M. &, Hogue, R. (2009) “Reduction in agricultural non-point source pollution in the first year following establishment of an integrated grass / tree filter strip system in southern Quebec (Canada)”. Agriculture, Ecosystems and Environment, 131, pp. 85-97.

Dupont, L. (2009) “Perspectives on the application of molecular genetics to earthworm ecology”. Pedobiologia, 52, pp. 191-205.

Duursma, E. K., Hours, R. & Leo, G. (1973) “Health and Safety Aspects”. In. Tracer Techniques in Sediment Transport. Technical Report Series No. 145, pp. 229-232. International Atomic Energy Agency (IAEA). Vienna, Austria.

Eckert & Zeigler (2005) “Reference & Calibration Sources Product Information”. Eckert & Zeigler Isotope Products Laboratories, Berlin, Germany. Edwards, C. A. & Bohlen, P. J. (1996) “Biology and Ecology of Earthworms, 3rd ed.”. Chapman & Hall: London. Edwards, C. A. & Lofty, J. R. (1977) “Biology of Earthworms 2nd ed.”. Chapman & Hall: London.

EG&G ORTEC (1990) “Solid-State Photon Detector GMW Series Operators Manual”. EG&G ORTEC, USA.

EG&G ORTEC (1998) “NomadTM Plus Portable Spectroscopy System: Hardware Reference Manual”. EG&G ORTEC, USA.

El-Swaify, S. A. & Flach, K. W. (1988) “Foreword”. In: Soil Erosion Research Methods (ed. R. Lal). Soil and Water Conservation Society, Iowa: USA.

245

Emadi, M., Baghernejad, M & Memarian, H. R. (2009) “Effect of land-use change on soil fertility characteristics within water-stable aggregates of two cultivated soils in northern Iran” Land Use Policy 26, pp. 452-457.

Environment Agency (2008) “North Devon Catchment Flood Management Plan: Scoping Report. Chapter. 2, Catchment Overview”, pp 9-31. Environment Agency: http://www.environmentagency.gov.uk/commondata/acrobat/catchment_overview_138 9710.pdf (accessed: 24/01/08).

Environmental Protection Agency (EPA) (1999) “Understanding variation in partition coefficient Kd values – Attachment C: Radiological Properties for SSL Development”. Office of Air & Radiation, Environmental Protection Agency, USA.

Environmental Protection Agency (EPA) (2002) “Radiation Protection: Cobalt”. Environmental

Protection

Agency,

USA.

www.epa.gov/rpdweb00/radionuclides/cobalt.html (accessed 01/12/06).

Erpul, G., Norton, L. D. & Gabriels, D. (2003) “Sediment transport from interrill areas under wind-driven rain”. Journal of Hydrology, 276, pp. 184-197.

Europa (2008) “The EU Water Framework Directive – Integrated River Basin Management for Europe: Timetable for Implementation”. http://ec.europa.eu/environment/water/water-framework/info/timetableen.htm (accessed 02/12/07).

Evans, G. V. (1983) “Tracer techniques in hydrology”. The International Journal of Applied Radiation and Isotopes, Vol. 34, pp. 451-475.

Evans, R. & Brazier, R. (2005) “Evaluation of modelled spatially distributed predictions of soil erosion by water versus field-based assessments”. Environmental Science & Policy, 8, pp. 493-501.

Farenhorst, A. & Bryan, R. B. (1995) “Particle size distribution of sediment transported by shallow flow”. Catena, 25, pp. 47-62.

246

Feller, C., Brown, G. G., Blanchart, E., Deleporte, P. & Chernyanskii, S. S. (2003) “Charles Darwin, earthworms and the natural sciences: various lessons from past to future”. Agriculture, Ecosystems and Environment, 99, pp. 29-49.

Ferro, V. & Minacapilli, M. (1995) “Sediment delivery processes at basin scale”. Hydrological Sciences, 40, pp. 703-717.

Foster, I. D. L. & Lees, J. A. (2000) “Tracers in geomorphology: Theory and application in tracing fine particulate sediments”. In: Tracers in Geomorphology (ed. I. D. L. Foster), Chpt. 1, pp. 3-21. John Wiley & Sons Ltd. Chichester, England.

Fox, D. M. & Bryan, R. B. (1999) “The relationship of soil loss by interrill erosion to slope gradient”. Catena, 38, pp. 211-222.

Fox, J. F. & Papanicolaou, A. N. (2008) “An un-mixing model to study watershed erosion processes”. Advances in Water Resources, 31, pp. 96-108.

Fredericks, D. J., Norris, V. & Perrens, S. J. (1988) “Estimating erosion using caesium-137: I. Measuring caesium-137 activity in a soil”. In: Sediment Budgets. International Association of Hydrological Sciences (IAHS), Publication No. 174, pp. 225-240.

Fullen M. A. (1982) “Laboratory and field studies in the use of the isotope

59

Fe for

tracing soil particle movement”. Earth Surface Processes and Landforms, 7, pp. 285293.

Fullen, M. A., Booth, C. A. & Brandsma, R. T. (2006) “Long-term effects of grass ley set-aside on erosion rates and soil organic matter on sandy soils in east Shropshire, UK”. Soil & Tillage Research, 89, pp. 122-128.

Galabov, M., Kovachev, S. & Mavrodiev, V. (2003) “Investigations on the migration ability of some radionuclides in groundwater”. Geology and Geophysics, 46, pp. 223226.

247

Gao, B., Walter, M. T., Steenhuis, T. S., Parlange, J. Y., Nakano, K., Rose, C. W. & Hogarth, W. L. (2003) “Investigating ponding depth and soil detachability for a mechanistic erosion model using a simple experiment”. Journal of Hydrology, 277, pp. 116-124.

Godfrey, G. K. (1954) “Tracing field voles (Microtus agrestis) with a Geiger-Muller Counter”. Ecology, 35, pp. 5-10.

Golosov, V. N., Walling, D. E., Kvasnikova, E. V., Stukin, E. D., Mikolaev, A. N. & Panin, A. V. (2000) “Application of a field-portable scintillation detector for studying the distribution of

137

Cs inventories in a small basin in Central Russia”. Journal of


Gölz, E. (2002) “Iffezheim field test – three years experience with a petrographic tracer”. In: The Structure, Function and Management Implications of Fluvial Systems. International Association of Hydrological Sciences (IAHS), Publication No. 276, pp. 417-424.

Görres, J. H., Savin, M. C. & Amador, J. A. (2001) “Soil micropore structure and carbon mineralization in burrows and casts of an anecic earthworm (Lumbricus terrestris). Soil Biology & Biochemistry, 33, pp. 1881-1887.

Govers, G. (1992) “Relationship between discharge, velocity and flow area for rills eroding loose, non-layered materials”. Earth Surface Processes and Landforms, 17, pp. 515-528.

Govers, G., Lobb, D. A. & Quine, T. A. (1999) “Tillage erosion and translocation emergence of a new paradigm in soil erosion research”. Soil Tillage Research, 51, pp. 167-174.

Govers, G., Quine, T. A., Desmet, P. J. J. & Walling, D. E. (1996) “The relative contribution of soil tillage and overland flow erosion to soil redistribution on agricultural land”. Earth Surface Processes and Landforms, 21, pp. 929-946.

248

Govers, G., Vandaele, K., Desmet, P. J. J., Poesen, J. & Bunte, K. (1994) “The role of tillage in soil redistribution on hillslopes”. European Journal of Soil Science, 45, pp. 469-478.

Granger, S. J., Bol, R., Butler, P. J., Haygarth, P. M., Naden, P., Old, G., Owens, P. N. & Smith, B. P. G. (2007) “Processes affecting transfer of sediment and colloids, with associated phosphorous, from intensively farmed grasslands: tracing and organic matter”. Hydrological Processes, 21, pp. 417-422.

Greene, R. S. B. & Hairsine, P. B. (2004) “Elementary processes of soil-water interaction and thresholds in soil surface dynamics: A review”. Earth Surface Processes and Landforms, 29, pp. 1077-1091.

Gryze de, S., Six, J, Bossuyt, H. Van Oost, K. & Merckx, R. (2008) “The relationship between landform and the distribution of soil C, N and P under conventional and minimum tillage”. Geoderma, 144, pp. 180-188.

Gudelis, A., Druteikienė, R., Lukšiene, B., Gvozdaitė, R., Nielsen, S. P. Hou, X., Mażeika, J. & Petrošius, R. (2008) “Assessing deposition levels of 55Fe, 60Co and 63Ni in the Ignalina NPP environment”. Journal of Environmental Radioactivity, (article in press).

Guiresse, M. & Revel, S. C. (1995) “Erosion due to cultivation of calcareous clay soils on hillsides in South-West France. II. Effect of ploughing down the steepest slope”. Soil & Tillage Research, 35, pp. 157-166.

Hairsine, P. B., Sander, G. C., Rose, C. W., Parlange, J. Y., Hogarth, W. L., Lisle, I. & Rouhipour, H. (1999) “Unsteady soil erosion due to rainfall impact: a model of sediment sorting on the hillslope”. Journal of Hydrology, 220, pp. 115-128.

Hart, M. R., Quin, B. F. & Nguyen, M. L. (2004) “Phosphorous runoff from agricultural land and direct fertiliser effects: A review”. Journal of Environmental Quality, 33, pp. 1954-1972.

249

Haygarth, P. (2005) “Linking landscape sources of phosphorous and sediment to ecological impacts in surface waters”. Science of the Total Environment, 344, pp. 1-3.

Haygarth, P. M., Bilotta, G. S., Bol, R., Brazier, R. E., Butler, P. J., Freer, J., Gimbert, J., Granger, S. J., Kreuger, T., Macleod, C. J. A., Naden, P., Old, G., Quinton, J. N. Smith, B. & Worsfold, P. (2006) “Processes affecting transfer of sediment and colloids, with associated phosphorous, from intensively farmed grasslands: an overview of key issues”. Hydrological Processes, 20, pp. 4407-4413.

He, Q. & Walling, D. E. (1996) “Interpreting particle size effects in the adsorption of 137

Cs and unsupported 210Pb by mineral soils and sediments”. Journal of Environmental

Radioactivity, 30, pp. 117-137.

He, Q. & Walling, D. E. (2000) “Calibration of a field-portable gamma detector to obtain in situ measurements of the

137

Cs inventories of cultivated soils and floodplain

sediments”. Applied Radiation and Isotopes, 52, pp. 865-872.

He, Q., Walling, D. E. & Wallbrink, P. J. (2002) “Alternative Methods and Radionuclides for Use in Soil-Erosion and Sedimentation Investigations”. Cp. 9, pp. 185-215. In: F. Zapata (ed.) (2002) Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides. Kluwer Academic Publishers: London.

Heathwaite, A. L., Quinn, P. F. & Hewett, C. J. M. (2005) “Modelling and managing critical source areas of diffuse pollution from agricultural land using flow connectivity simulation”. Journal of Hydrology, 304, pp. 446-461.

Heckrath, G., Djurhuus, J., Quine, T. A., Van Oost, K., Govers, G. & Zhang, Y. (2005) “Tillage erosion and its effect on soil properties and crop yield in Denmark”. Journal of Environmental Quality, 34, pp. 312-324.

Heckrath, G., Halekoh, U., Djurhuus, J. & Govers, G. (2006) “The effect of tillage direction on soil redistribution by mouldboard ploughing on complex slopes”. Soil & Tillage Research, 88, pp. 225-241.

250

Heilig, A., DeBruyn, D., Walter, M. T., Rose, C. W., Parlange, J. Y., Steenhuis, T. S., Sander, G. C., Hairsine, P. B., Hogarth, W. L. & Walker, L. P. (2001) “Testing a mechanistic soil erosion model with a simple experiment”. Journal of Hydrology, 244, pp. 9-16.

Heimsath, A. M., Dietrich, W. E., Nishilzumi, K. & Finkel, R. C. (1997) “The soil production function and landscape equilibrium”. Nature, 388, pp. 358-361.

Helalia, A. M., Letey, J. & Graham, R. C. (1988) “Crust formation and clay migration effects on infiltration rate”. Soil Science Society of America Journal, 52, pp. 251-255.

Higgitt, D. L., Rowan, J. S. & Walling, D. E. (1993) “Catchment-scale deposition and redistribution

of

Chernobyl

radiocaesium

in

upland

Britain”.

Environment

International, 19, pp. 155-166.

Hogarth, W. L., Parlange, J. Y., Rose, C. W. Sander, G. C., Steenhuis, T. S. & Barry, A. (2004) “Soil erosion due to rainfall impact with no inflow: an analytical solution with spatial and temporal effects”. Journal of Hydrology, 295, pp. 140-148.

Huang, C., Wells, L. K. & Norton, L. D. (1999) “Sediment transport capacity and erosion processes: model concepts and reality”. Earth Surface Processes and Landforms, 24, pp. 503-516.

Jouquet, P., Bottinelli, N., Podwojewski, P., Hallaire, V. & Duc, T. T. (2008a) “Chemical and physical properties of earthworm casts as compared to bulk soil under a range of different land-use systems in Vietnam”. Geoderma, 146, pp. 231-238.

Jouquet, P., Podwojewski, P., Bottinelli, N., Mathieu, J., Ricoy, M., Orange, D., Tran, T. T. & Valentin, C. (2008b) “Above-ground earthworm casts affect water runoff and soil erosion in Northern Vietnam”. Catena, 74, pp. 13-21.

Jungerius, P. D., Dekker, L. W. (1990) “Water erosion in the dunes”. Catena, 18, (supplement), pp. 185– 193.

251

Jungerius, P. D., Verheggen, A. J. T. & Wiggers, A. J. (1981) “The development of blowouts in De Blink, a coastal dune area near Noordwijkerout, The Netherlands”. Earth Surface Processes and Landforms, 6, pp. 375– 396.

Junliang, T., Peihua, Z. & Puling, L. (1994) “REE tracer method for studies on soil erosion”. International journal of Sediment Research, 9, pp. 39-46

Juzwik, J., Stenlund, D. L., Allmaras, R. R., Copeland, S. M. & McRoberts, R. E. (1997) “Incorporation of tracers and dazomet by rotary tillers and a spading machine” Soil & Tillage Research, 41, pp. 234-248.

Kauffman, J. B. & Krueger, W. C. (1984) “Livestock impacts on riparian ecosystems and streamside management implication: A review”. Journal of Range Management 3, pp. 430-438.

Kellett, A. J. (1978). ‘‘Poaching of Grassland and the Role of Drainage. Technical Report 78/1.’’

Khan, S. A. (2003) “Sorption of the long-lived radionuclides cesium-134, strontium-85 and cobalt-60 on bentonite”. Journal of Radioanalytical and Nuclear Chemistry, 258, pp. 3-6.

Kimoto, A., Nearing, M. A., Zhang, X. C. & Powell, D. M. (2006) “Applicability of rare earth element oxides as a sediment tracer for course-textured soils”. Catena, 65, pp. 214-221.

Kizilkaya, R. (2008) “Dehydrogenase activity in Lumbricus terrestris casts and surrounding soil affected by addition of different organic wastes and Zn”. Bioresource Technology, 99, pp. 946-953.

Kladivko, E. J., Mackay, A. D. & Bradford, J. M. (1986) “Earthworms as a factor in the reduction of soil crusting”. Soil Science Society of America Journal, 50, pp. 191196.

252

Köhne, J. M., Köhne, S. & Šimůnek, J. (2009) “A review of model applications for structured soils: b) pesticide transport”. Journal of Contaminant Hydrology, 104, pp. 36-60.

Kosmas, C., Gerontidis, S., Marathianou, M., Detsis, B., Zafirou, T., Muysen, W. N., Govers, G., Quine, T. & Van Oost, K. (2001) “The effects of tillage displaced soil on soil properties and wheat biomass”. Soil and Tillage Research, 58, pp. 31-44.

Kuhn, N. J. & Bryan, R. B. (2004) “Drying, soil surface condition and inter-rill erosion on two Ontario soils”. Catena, 57, pp. 113-133.

Kurz, I., Daniel, O’Reilly, C. D. & Tunney, H. (2006) “Impact of cattle on soil physical properties and nutrient concentrations in overland flow from pasture in Ireland”. Agriculture, Ecosystems & Environment, 113, pp. 378-390.

Laissaoui, A. & Abril, J. M. (1999) “A theoretical technique to predict the distribution of radionuclides bound to particles in surface sediments”. Journal of Environmental Radioactivity, 44, pp. 71-84.

Lal, R. (1988) “Preface”. In: Soil Erosion Research Methods (ed. R. Lal). Soil and Water Conservation Society, Iowa: USA.

Lambert, C. P. & Walling, D. E. (1987) “Floodplain sedimentation: A preliminary investigation of contemporary deposition within the lower reaches of the River Culm, Devon, UK”. Geografiska Annaler, 69 A, pp. 393-404.

Lang, A. (2008) “Recent advances in dating and source tracing of fluvial deposits”. In: Sediment Dynamics in Changing Environments. International Association of Hydrological Sciences (IAHS), Publication No. 325, pp. 3-12.

Langlands, J. P., and Bennett, I. L. (1973). “Stocking intensity and pastoral production 1. Changes in the soil and vegetation of a sown pasture grazed by sheep at different stocking rates”. Journal of Agricultural Science, 81, 193–204.

253

Lavelle, P. “Earthworm activities and the soil system”. Biological Fertility of Soils, 6, pp. 237-251.

Le Bayon, R. C. & Binet, F. (1999) “Rainfall effects on erosion of earthworm casts and phosphorous transfers by water runoff”. Biological Fertility of Soils, 30, pp. 7-13.

Le Bayon, R. C. & Binet, F. (2001) “Earthworm surface casts affect soil erosion by run-off and phosphorous transfer in a temperate maize crop”. Pedobiologia, 45, pp. 430-442.

Le Bayon, R.C., Moreau, S. Gascuel-Odoux, C. & Binet, F. (2002) “Annual variations in earthworm surface-casting activity and soil transport by water run-off under a temperate maize agroecosystem”. Geoderma, 106, pp. 121-135.

Le Bissonais, Y., Renaux, B. & Delouche, H. (1995) “Interactions between soil properties and moisture content in crust formation, run-off and inter-rill erosion from tilled loess soils”. Catena, 25, pp. 33-46.

Le Bissonnais, Y., Cerdan, O., Lecomte, V., Benhadra, H., Souchère, V. & Martin, P. (2005) “Variability of soil surface characteristics influencing runoff and interrill erosion. Catena, 62, pp. 111-124.

Lee, H. S., Bhang, H., Hahn, I. S., Hwang, M. J., Kim, H. J., Kim, S. C. Kim, S. K., Kim, S. Y., Kim, T. Y. Kim, Y. D., Kwak, J. W., Kwon, Y. J., Lee, J., Lee, J. I., Lee, M. J., Li, J., Myung, S. S., Park, H. & Zhu, J. J. (2007) “Development of low background CsI(T

) crystals for WIMP search”. Nuclear Instruments and Methods in

Physics Research, 571, pp. 644-650.

Lee, K. E. (1985) “Earthworms: Their Ecology and Relationships with Soils and Land Use”. Academic Press (Harcourt Brace Jovanovich Publishers): London.

Leeners, H. & Schouten, C. J. (1989) “Soil erosion and floodplain soil pollution: Related problems in the geographical context of a river basin”. In: Sediment and the Environment. International Association of Hydrological Sciences (IAHS), Publication No. 184, pp. 75-83. 254

Legout, C., Leguédois, S., Le Bissonais, Y. & Issa, O. M. (2005) “Splash distance and size distributions for various soils”. Geoderma, 124, pp. 279-292.

Leguédois, S. & Le Bissonais, Y. (2004) “Size fractions resulting from an aggregate stability test, interrill detachment and transport”. Earth Surface Processes and Landforms, 29, pp. 1117-1129.

Leibundgut, C., Speidel, U. & Wiesner, H. (1993) “Transport processes in rivers investigated by tracer experiments”. In: Tracers in Hydrology. International Association of Hydrological Sciences (IAHS), Publication No. 215, pp. 211-217.

Lettner, H., Andrasi, A., Hubmer, A. K., Lovranich, E., Steger, F. & Zombori, P. (1996) “In situ gamma-spectrometry intercomparison exercise in Salzburg, Austria”. Nuclear Instruments and Methods in Physics Research A, 369, pp. 547-551.

Levy, G., Shainberg, I. & Morin, J. (1986) “Factors affecting the stability of soil crusts in subsequent storms”. Soil Science Society of America Journal, 50, pp. 196-201.

Lewin, J. & Hughes, D. (1980) “Welsh floodplain studies II: application of a qualitative inundation model”. Journal of Hydrology, 46, pp. 35-49.

Li, S., Lobb, D. A. & Lindstrom, M. J. (2007) “Tillage translocation and tillage erosion in cereal-based production in Manitoba, Canada”. Soil & Tillage Research, 94, pp. 164-182.

Li. M., Li, Z. B., Ding, W. F., Liu, P. L. & Y, W. Y. (2006) “Using rare earth element tracers and neutron activation analysis to study rill erosion processes”. Applied Radiation and Isotopes, 64, pp. 402-408.

Lindstrom, M. J., Nelson, W. W. & Schumacher, T. E. (1992) “Quantifying tillage erosion rates due to moldboard plowing”. Soil & Tillage Research, 24, pp. 243-255.

Lindstrom, M. J., Nelson, W. W., Schumacher, T. E. & Lemme, G. D. (1990) “Soil movement by tillage as affected by slope”. Soil & Tillage Research, 17, pp. 255-264. 255

Liu, J. (2005) “Study of Tillage Tool-Soil-Crop Residue Interaction”. Unpublished PhD. Thesis, Department of Biosystems Engineering, University of Manitoba, Winnipeg, Canada.

Liu, L. L., Shi, J. J., Zhao, X. Y. & Hua, Y. J. (2005) “Dynamics of transfer and distribution of

95

Zr in the broadbean-soil ecosystem”. Journal of Environmental


Liu, P. L., Tian, J. L., Zhou, P. H., Yang, M. Y. & Shi, H. (2004) “Stable rare earth element tracers to evaluate soil erosion”. Soil & Tillage Research, 76, pp. 147-155.

Lobb, D. A. (2006) “Modelling Soil Movement by Tillage”. Form 101, graduate grant application (unpublished) to the Natural Sciences and Engineering Research Council of Canada (NSERC). Department of Biosystems Engineering, University of Manitoba, Winnipeg, Canada.

Lobb, D. A. & Kachanoski, R. G. (1994) “Quantification of tillage translocation and tillage erosion”. Canadian Journal of Soil Science, 74, pp. 353.

Lobb, D. A. & Kachanoski, R. G. (1999) “Modelling tillage translocation using step, linear-plateau and exponential functions”. Soil & Tillage Research, 51, pp. 317-330.

Lobb, D. A., Kachanoski, R. G. & Miller, M. H. (1995) “Tillage translocation and tillage erosion on shoulder slope landscape positions measured using 137Cs as a tracer”. Canadian Journal of Soil Science, 75, pp. 211-218.

Lobb, D. A., Kachanoski, R. G. & Miller, M. H. (1999) “Tillage translocation and tillage erosion in the complex upland landscapes of southwestern Ontario, Canada”. Soil & Tillage Research, 51, pp. 189-209.

Lobb, D. A., Quine, T. A., Govers, G. & Heckrath, G. R. (2001) “Comparison of methods used to calculate tillage translocation using tracers”. Journal of Soil & Water Conservation, 56, pp. 321-328.

256

Loch, R. J. & Foley, J. L. (1992) “Effects of plot size distributions of water-stable material at the soil surface under simulated rainfall”. Journal of Australian Soil Research, 30, pp. 113-118.

Loughran, R. J. (1989) “The measurement of soil erosion”. Progress in Physical Geography, 13, pp. 216-233. Lowe, J. J. & Walker, M. J. C. (1997) “Reconstructing Quaternary Environments 2nd ed.”. Prentice Hall: London.

Lu, N. & Mason, C. F. V. (2001) “Sorption-desorption behavior of strontium-85 onto montmorillonite and silica colloids”. Applied Geochemistry, 16, pp. 1653-1662.

Mabit, L., Benmansour, M. & Walling, D. E. (2008) “Comparative advantages and limitations of the fallout radionuclides

137

Cs,

210

Pbex and 7Be for assessing soil erosion

and sedimentation”. Journal of Environmental Radioactivity, 99, pp. 1799-1807. MacKenzie, A. B. (2000) “Environmental radioactivity: experience from the 20th century – trends and issues for the 21st century”. The Science of the Total Environment, 249, pp. 313-329.

Manuwa, S. I. (2009) “Performance evaluation of tillage tines operating under different depths in a sandy clay loam soil”. Soil & Tillage Research, 103, pp. 399-405.

Mariani, L., Jiménez, J. J., Torres, E. A., Amézquita, E. & Decaëns, T. (2007) “Rainfall impact effects on ageing casts of a tropical anecic earthworm”. European Journal of Soil Science, 58, pp. 1525-1534.

Martinez-Mena, M., Castillo, V. & Albaladejo, J. (2002) “Relations between interrill erosion processes and sediment particle size distribution in a semi-arid Mediterranean area of SE Spain”. Geomorphology, 45, pp. 261-275.

Martinez-Mena, M., Lopez, J., Almagro, M., Boix-Fayos, C. & Albaladejo, J. (2008) “Effect of water erosion and cultivation on the soil carbon stock in a semiarid area of South-East Spain”. Soil & Tillage Research, 99, pp. 119-129. 257

Massas, I., Skarlou, V. & Haidouti, C. (2002) “Plant uptake of 134Cs in relation to soil properties and time”. Journal of Environmental Radioactivity, 59, pp. 245-255.

Mathier, L. & Roy, A. G. (1996) “A study on the effect of spatial scale on the parameters of a sediment transport equation for sheetwash”. Catena, 26, pp. 161-169.

Matishov, G. G., Matishov, D. G., Namjatov, A. A., Carroll, JL. & Dahle, S. (1999) “Anthropogenic radionuclides in Kola and Motovsky Bays of the Barents Sea, Russia”. Journal of Environmental Radioactivity, 43, pp. 77-88.

McCubbin, D. & Leonard, K. S. (1995) “Use of radiotracers for studies of metal sorption behaviour”. Science of the Total Environment, 173/174, pp. 259-266.

McKenna, P. & Longworth, R. D. (1995) “Residual Chernobyl fallout and Sellafield pollutants found on the Isle of Man”. The Science of the Total Environment, 173 / 174, pp. 7-14.

McKenzie, A. L., Kenwick, S. D. & Boddy, K. (1989) “Analytical survey of body radioactivity in the general public”. Journal of Radiological Protection, 9, pp. 247-250.

Mech, S. J. & Free, G. R. (1942) “Movement of soil during tillage operations”. Agricultural Engineering, 23, pp. 379-382.

Meyer, L. D. (1988) “Rainfall simulators for soil conservation research”. Chpt. 4. In: Soil Erosion Research Methods (ed. R. Lal). Soil and Water Conservation Society, Iowa: USA.

Meyer, L. D. & Harmon, W. C. (1984) “Susceptibility of agricultural soils to interrill erosion”. Soil Science Society of America Journal, 48, pp. 1152-1157.

Meyer, L. D., Foster, G. R. & Nikolov, S. (1975) “Effect of flow rate and canopy on rill erosion”. Transactions of the American Society of Agricultural Engineers, 18, pp. 905-911.

258

Middelkoop, H. & Van Der Perk, M. (1998) “Modelling spatial patterns of overbank sedimentation on embanked floodplains”. Geografiska Annaler, 80A, pp. 95-109.

Miller, K. M. (1998) “Notes describing the use and applications of in-situ gamma spectrometry”. EG&G ORTEC Nuclear Instruments, USA.

Miller, W. P. & Baharuddin, M. K. (1987) “Particle size of interrill-eroded sediments from highly weathered soils”. Soil Science Society of America Journal, 51, pp. 16101615.

Millington, A. C. (1981) “Relationship between three scales of erosion measurement on two small basins in Sierra Leone”. In: Erosion and Sediment Transport Measurement. International Association of Hydrological Sciences (IAHS) Publication, No. 133, pp. 485-492.

Moeyersons, J. (1983) “Measurements of splash – saltation fluxes under oblique rain”. Catena, 4 (supplement), pp. 19– 31.

Mohler, C. L., Frisch, J. C. & McCulloch, C. E. (2006) “Vertical movement of weed seed surrogates by tillage implements and natural processes”. Soil & Tillage Research, 86, pp. 110-122.

Montgomery, D. R. (2007) “Dirt: The Erosion of Civilisations”. University of California Press: London.

Morgan, R. P. C. (1993) “Soil Erosion and Conservation”. (ed. D. A. Davidson), John Wiley & Sons Inc., New York.

Morris, P., Biggs, J. & Brookes, A. (2001) “Water”. Chpt. 10. In: Methods of Environmental Impact Assessment 2nd ed. (eds. P. Morris & R. Therivel). Spon Press: London.

Mulholland, B., and Fullen, M. A. (1991). “Cattle trampling and soil compaction on loamy sands”. Soil Use and Management, 7, pp. 189–193.

259

Müller-Lehmans, H. & Van Dorp, F. (1996) “Bioturbation as a mechanism for radionuclide transport in soil: relevance of earthworms”. Journal of Environmental Radioactivity, 31, pp. 7-20.

Murray, A. S., Marten, R., Johnston, A. & Martin, P. (1987) “Analysis for naturally occurring radionuclides at environmental concentrations by gamma spectrometry”. Journal of Radiology and Nuclear Chemistry Articles, 115, pp. 263-288.

Mutchler, C. K., Murphree, C. E. & McGregor, K. C. (1988) “Laboratory and field plots for soil erosion studies”. Chpt. 2. In: Soil Erosion Research Methods (ed. R. Lal). Soil and Water Conservation Society, Iowa: USA.

Nakamura, F. & Kikuchi, S. (1996) “Some methodological developments in the analysis of sediment transport processes using age distribution of floodplain deposits”. Geomorphology, 16, pp. 139-145.

Nash, D. & Murdoch, C. (1997) “Phosphorous in runoff from a fertile dairy pasture”. Australian Journal of Soil Research, 35, pp. 419-429.

National Water Archives (2008) “Concise Register of Gauging Stations, Environment Agency

–

Southwest:

50001

–

Taw

at

Umberleigh”.

http://www.nwl.ac.uk/ih/nrfa/station_summaries/050/001.html (accessed 21/02/08).

Nearing, M. A., Norton, L. D., Bulgakov, D. A., Larionov, G. A., West, L. T. & Dontsova, K. M. (1997) “Hydraulics and erosion in eroding rills”. Water Resources Research, 33, pp. 865-876.

Nicholas, A. P. & Walling, D. E. (1997) “Investigating spatial patterns of mediumterm overbank sedimentation on floodplains: a combined numerical modelling and radiocaesium-based approach”. Geomorphology, 19, pp. 133-150.

Nicholas, A. P., Walling, D. E., Sweet, R. J. & Fang, X. (2006) “New strategies for upscaling high-resolution flow and overbank sedimentation models to quantify floodplain sediment storage at the catchment scale”. Journal of Hydrology, 329, pp. 577-594. 260

Oldeman, L. R. (1998) “Soil Degradation: A Threat to Food Security?” Report 98/01. International Soil Reference and Information Centre, Wageningen.

Oliver, M. A. & Webster, R. (1990) “Kriging: a method of interpolation for Geographical

Information

Systems”.

International

Journal

of

Geographical

Information Science, 4, pp. 313-332.

Onofiok, O. & Singer, M. J. (1984) “Scanning electron microscope studies of surface crusts formed by simulated rainfall”. Soil Science Society of America Journal, 48, pp. 1137-1143.

Onwuala, A. P. & Watts, K. D. (1998) “Draught and vertical forces obtained from dynamic soil cutting by plane tillage tools”. Soil & Tillage Research, 48, pp. 239-253.

Oostwoud Widjenes, D. J. & Poesen, J. (1999) “The effect of soil moisture on the vertical movement of rock fragments by tillage”. Soil & Tillage Research, 49, pp. 301312.

Ordnance Survey (OS) (1980) “Geological Survey of Great Britain, England & Wales. Chulmleigh, Sheet 309”. Scale 1:50,000 Series. Institute of Geological Sciences. Natural Environment Research Council, UK.

Owens, P. N., Duzant, J. H., Deeks, L. K., Wood, G. A., Morgan, R. P. C. & Collins, A. J. (2007) “Evaluation of contrasting buffer features within an agricultural landscape for reducing sediment and sediment-associated phosphorous delivery to surface waters”. Soil Use & Management, British Society of Soil Sciences, 23 (Supplement 1), pp. 165-175.

Owens, P. N. & Walling, D. E. (1998) “The use of a numerical mass-balance model to estimate rates of soil redistribution on uncultivated land from

137

Cs measurements”.

Journal of Environmental Radioactivity, 40, pp. 185-203.

Owens, P. N. & Walling, D. E. (2002) “Changes in sediment sources and floodplain deposition rates in the catchment of the River Tweed, Scotland, over the last 100 years: 261

The impact of climate and land use change”. Earth Surface Processes and Landforms, 27, pp. 403-423.

Owens, P. N., Walling, D. E. & He, Q. (1996) “The behaviour of bomb-derived caesium-137 fallout in catchment soils”. Journal of Environmental Radioactivity, 32, pp. 169-191.

Owens, P. N., Walling, D. E. & Leeks, G. J. L. (1999) “Deposition and storage of fine-grained sediment within the main channel system of the River Tweed, Scotland”. Earth Surface Processes and Landforms, 24, pp. 1061-1076.

Owens, P. N., Walling, D. E. & Leeks, G. J. L. (1999) “Use of floodplain sediment cores to investigate recent historical changes in overbank sedimentation rates and sediment sources in the catchment of the River Ouse, Yorkshire, UK”. Catena, 36, pp. 21-47. Owens, P. N., Walling, D. E. & Leeks, G. J. L. (1999b) “Deposition and storage of fine-grained sediment within the main channel system of the River Tweed, Scotland”. Earth Surface Processes and Landforms, 24, pp. 1061-1076.

Pahlke, H. (1973) “Some applications using artificial radioactive tracers in Germany”. In. Tracer Techniques in Sediment Transport. Technical Report Series No. 145, pp. 8197. International Atomic Energy Agency (IAEA). Vienna, Austria.

Pan, C. & Shangguan, Z. (2006) “Runoff hydraulic characteristics and sediment generation in sloped grassplots under simulated rainfall conditions”. Journal of Hydrology, 331, pp. 178-185.

Papiernik, S. K., Schumacher, T. E., Lobb, D. A., Lindstrom, M. J., Lieser, M. L., Eynard, A. & Schumacher, J. A. (2009) “Soil properties and productivity as affected by topsoil movement within an eroded landform”. Soil & Tillage Research, 102, pp. 6777.

Parsons, A. J., Abrahams, A. D. & Wainwright, J. (1994) “Rainsplash and erosion rates in an interrill area on semi-arid grassland, Southern Arizona”. Catena, 21, pp. 215226. 262

Parsons, A. J., Brazier, R. E., Wainwright, J. & Powell, D. M. (2006) “Scale relationships in hillslope runoff and erosion”. Earth Surface Processes and Landforms, 31, pp. 1384-1393.

Parsons, A. J., Stromberg, S. G. L. & Greener, M. (1998) “Sediment-transport competence of rain-impacted interrill overland flow”. Earth Surface Processes and Landforms, 23, pp. 365-375.

Parsons, A. J. & Wainwright, J. (2006) “Depth distribution of interrill overland flow and the formation of rills”. Hydrological Processes, 20, pp. 1511-1523.

Parsons, A. J., Wainwright, J. & Abrahams, A, D. (1993) “Tracing sediment movement in interrill overland flow on a semi-arid grassland hillslope using magnetic susceptibility”. Earth Surface Processes and Landforms, 18, pp. 721-732.

Patterson, J. B. E. (1939) “Soil erosion in England”. Nature, 3611, p.79.

Patto, P. M., Clement, C. R., and Forbes, T. J. (1978). ‘‘Grassland Poaching in England and Wales.’’ Grassland Research Institute, Maidenhead.

Payne, T. E., Hatje, V., Itakura, T., McOrist. G. D. & Russell, R. (2004) “Radionuclide applications in laboratory studies of environmental surface reactions”. Journal of Environmental Radioactivity, 76, pp. 237-251.

Payne, T. E., Itakura, T., Comarmond, M. J. & Harrison, J. J. (2009) “Environmental mobility of cobalt – Influence of solid phase characteristics and groundwater chemistry”. Applied Radiation and Isotopes, 67, pp. 1269-1276.

Pennock, D. J. & Appleby, P. G. (2002) “Site selection and sampling design”, Chpt. 2, pp. 15-40. In: Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides (ed. F Zapata). Kluwer Academic Publishers, London, UK.

263

Phillips, J. D. (1993) “Pre- and post-colonial sediment sources and storage in the Lower Neuse basin, North Carolina”. Physical Geography, 14, pp. 272-284.

Pietola, L. Horn, R. & Yli-Halla, M. (2005) “Effects of trampling by cattle on the hydraulic and mechanical properties of soil”. Soil & Tillage Research, 82, pp. 99-108.

Pizzuto, J. E. (1987) “Sediment diffusion during overbank flows”. Sedimentology, 34, pp. 301-317.

Plante, A. F., Duke, M. J. M. & McGill, W. B. (1999) “A tracer detectable by neutron activation for soil aggregation and translocation studies”. Soil Science Society of America Journal, 63, pp. 1284-1290.

Poesen, J. W., Van Wesmael, B., Bunte, K. & Benet, A. S. (1998) “Variation of rock fragment cover and size along semiarid hillslopes: a case study from southeast Spain”. Geomorphology, 23, pp. 323-335.

Polyakov, V. O. & Nearing, M. A. (2004) “Rare earth oxides for tracing sediment movement”. Catena, 55, pp. 255-276.

Poręba, G. J. (2006) “Caesium-137 as a soil erosion tracer: a review” Geochronometria, 25, pp. 37-46.

Proffitt, A. P. B. & Rose, C. W. (1991) “Soil erosion processes. I. The relative importance of rainfall detachment and run-off entrainment”. Australian Journal of Soil Research, 29, pp. 671-683.

Proffitt, A. P. B., Rose, C. W. & Hairsine, P. B. (1991) “Rainfall detachment and deposition: Experiments with low slopes and significant water depths”. Soil Science Society of America Journal, 55. pp. 325-332.

Quine, T. A., Govers, G., Poesen, J., Walling, D. E., van Wesemael, B. & MartinezFernandez, J. (1999) “Fine-earth translocation by tillage in stony soils in the Guadalentin, south-east Spain: an investigation using caesium-134”. Soil & Tillage Research, 51, pp. 279-301. 264

Quine, T. A. & Walling, D. E. (1991) “Rates of soil erosion on arable fields in Britain: quantitative data from caesium-137 measurements”. Soil Use and Management, 7, pp. 169-176.

Quine, T. A. & Zhang, Y. (2002) “An investigation of spatial variation in soil erosion, soil properties, and crop production within an agricultural field in Devon, United Kingdom”. Journal of Soil and Water Conservation, 57, pp. 55-65.

Rahman, S. & Chen, Y. (2001) “Laboratory investigation of cutting forces and soil disturbance resulting from different manure incorporation tools in a loamy sand soil”. Soil & Tillage Research, 58, pp. 19-29.

Rahman, S., Chen, Y. & Lobb, D. (2005) “Soil movement resulting from sweep type liquid manure injection tools”. Biosystems Engineering, 91, pp. 379-392.

Rahman, S., Chen, Y., Zhang, Q. & Lobb, D. (2005) “Evaluation methods on manure exposure from liquid manure injection tools”. Canadian Biosystems Engineering, 47, pp. 6.9-6.16.

Ramos, M. C., Nacci, S. & Pla, I. (2003) “Effect of raindrop impact and its relationship with aggregate stability to different disaggregation forces”. Catena, 53, pp. 365-376.

Real, J., Persin, F. & Camarasa-Claret, C. (2002) “Mechanisms of desorption of 134

Cs and

85

Sr aerosols deposited on urban surfaces”. Journal of Environmental


Reeder, J. D., Schuman, G. E. & Bowman, R. A. (1998) “Soil C and N changes on conservation reserve program lands in the Central Great Plains”. Soil & Tillage Research, 47, pp. 339-349.

Reicosky, D. C. & Archer, D. W. (2007) “Moldboard plow depth and short-term carbon dioxide release”. Soil & Tillage Research, 94, pp. 109-121.

265

Rejman, J., Usowicz, B. & Dębicki, R. (1999) “Source of errors in predicting silt soil erodibility with USLE”. Polish Journal of Soil Science, 32, pp. 13-22.

Ridley, N. (1986) “The Radioactive Substances (Substances of Low Activity) Exemption Order 1986”. Statutory Instruments, No 1002. Atomic Energy and Radioactivity

Substances.

http://www.ionactive.co.uk/pdfs/Substances_Low_Activity_Exemption_Order_111901 8891.pdf (accessed 02/11/06).

Riebe, B. (1995) “Monitoring the translocation of soil particles using a neutron activated tracer”. In: Soil Structure: Its Development and Function (eds. K. H. Hartge & B. A. Stewart), pp. 277-294. CRC Press Inc., Florida, USA.

Riley, H., Pommeresche, R., Eltun, R., Hansen, S. & Korsaeth, A. (2008) “Soil structure, organic matter and earthworm activity in a comparison of cropping systems with contrasting tillage, rotations, fertilizer levels and manure use”. Agriculture, Ecosystems and Environment, 124, pp. 275-284.

Ristic, M., Degetto, S., Ast, T. & Cantallupi, C. (2002) “Sample preparation for the determination of

241

Am in sediments utilizing γ-spectrometry”. Journal of


Ritchie, J. C. & McHenry, J. R. (1990) “Application of radioactive fallout cesium-137 for measuring soil erosion and sediment accumulation rates and patterns: A review”. Journal of Environmental Quality, 19, pp. 215-233.

Ritchie, J. C., Spraberry, J. A. & McHenry, J. R. (1974) “Estimating soil erosion from the redistribution of fallout

137

Cs”. Soil Science Society of America Proceedings,

38, pp. 137-139. Roberson, J. A. & Crowe, C. T. (1997) “Engineering Fluid Mechanics (6th ed)”. John Wiley & Sons, Inc. Chichester.

Robinson, D. A. & Boardman, J. D. (1989) “Soil erosion, soil conservation and agricultural policy for arable land in the UK”. Geoforum, 20, pp. 83-92. 266

Robinson, D. A. & Phillips, C. P. (2001) “Crust development in relation to vegetation and agricultural practise on erosion susceptible, dispersive clay soils from central and southern Italy”. Soil & Tillage Research, 60, pp. 1-9.

Roger-Estrade, J., Colbach, N., Leterme, P. Richard, G. & Caneill, J. (2001) “Modelling vertical and lateral weed movements during mouldboard ploughing with a skim-coulter”. Soil & Tillage Research, 63, pp. 35-49.

Roth, C. H. (1997) “Bulk density of surface crusts: depth functions and relationships to texture”. Catena, 29, pp. 223-237.

Rowan, J. S. (1990) “The Sediment-Associated Transport and Redistribution of Chernobyl-Derived Radiocaesium in Fluvial Systems”. Unpublished PhD. Thesis, Department of Geography, University of Exeter, UK.

Rowan, J. S. & Walling, D. E. (1992) “The transport and fluvial redistribution of Chernobyl-derived radiocaesium within the River Wye basin, UK”. The Science of the Total Environment, 1211, pp. 109-131.

Rowan, J. S., Goodwill, P. & Greco, M. (1995) “Temporal variability in catchment sediment yield determined from repeated bathymetric surveys: Abbeystead Reservoir, U.K.”. Physics and Chemistry of the Earth, 20, pp. 199-206.

Rowan, J. S., Price, L. E., Fawcett, C. P. & Young, P. C. (2001) “Reconstructing historic reservoir sedimentation rates using data-based mechanistic modelling”. Physics and Chemistry of the Earth, 26, pp. 77-82. Rühm, W., Kammerer, L. Hiersche, L. & Wirth, E. (1996) “Migration of 137Cs and 134

Cs in different forest soil layers”. Journal of Environmental Radioactivity, 33, pp. 63-

75.

Russell, M. A., Walling, D. E. & Hodgkinson, R. A. (2001) “Suspended sediment courses in two small lowland agricultural catchments in the UK”. Journal of Hydrology, 252, pp. 1-24. 267

Rybacek, K., Jacob, P. & Meckbach, R. (1992) “In-situ determination of deposited radionuclide activities: improved method using derived depth distributions from the measured photon spectra”. Health Physics, 62, pp. 519-528.

Sanchez, A. L., Smolders, E., Van den Brande, K,. Merckx, R., Wright, S. M. & Naylor, C. (2002) “Predictions of in situ solid/liquid distribution of radiocaesium in soils”. Journal of Environmental Radioactivity, 63, pp. 35-47. Sandeep, S. & Manjaiah, K. M. (2008) “Transfer factors of 134Cs to crops from Typic Haplustept under tropical region as influenced by potassium application”. Journal of Environmental Radiation, 99, pp. 349-358.

Sattar, A., Traniello, J. F. A., Salihah, Z. & Farid, A. (2002) “Determining different parameters of Co-60 bait as a radiotracer for three species of subterranean termites”. The Nucleus, 39, pp. 89-93.

Sauzay, G. (1973) “Tracer techniques in sediment transport: Report of the panel”. In. Tracer Techniques in Sediment Transport. Technical Report Series No. 145, pp. 3-8. International Atomic Energy Agency (IAEA). Vienna, Austria.

Saynor, M. J., Loughran, R. J., Erskine, W. D. & Scott, P. F. “Sediment movement on hillslopes measured by caesium-137 and erosion pins”. In: Variability in Stream Erosion and Sediment Transport. International Association of Hydrological Sciences (IAHS), Publication No. 224, pp. 87-93.

Scherr, S. J. & Yadav, S. (1996) “Land degradation in the developing world: Implications for food, agriculture and the environment to 2020”. Food, Agriculture and the Environment Discussion Paper 14. International Food Policy Research Institute, Washington DC, USA.

Schumacher, T. E., Lindstrom, M. J., Schumacher, J. A. & Lemme, G. D. (1999) “Modelling spatial variation in productivity due to tillage and water erosion”. Soil & Tillage Research, 51, pp. 331-339.

268

Sear, D. A., Lee, M. W. E., Oakey, R. J., Carling, P. A. & Collins, M. B. (2000) “Coarse sediment tracing technology in littoral and fluvial environments: A review”. In Tracers in Geomorphology (ed. I. D. L. Foster), Chpt. 2, pp. 21-55. John Wiley & Sons Ltd. Chichester, UK.

Seeger, M (2007) “Uncertainty of factors determining runoff and erosion processes as quantified by rainfall simulations” Catena, 71, pp. 56-67. Sepulveda, A., Schuller, P., Walling, D. E. & Castillo, A. (2008) “Use of 7Be to document soil erosion associated with a short period of extreme rainfall”. Journal of Environmental Radioactivity, 99, pp. 35-49.

Sharifat, K. (1999) “Soil Translocation With Tillage Tools”. Unpublished PhD. Thesis, Department of Agricultural and Bioreseource Engineering, University of Saskatchewan, Saskatoon, Canada.

Sharma, P. P., Gupta, S. C. & Rawls, W. J. (1991) “Soil detachment by single raindrops of varying kinetic energy” Soil Science Society of America Journal, 55, pp. 301-307.

Sharpley, A. N., Syers, J. K. & Springett, J. A. (1979) “Effect of surface-casting earthworms on the transport of phosphorous and nitrogen in surface runoff from pasture”. Soil Biology and Biochemistry, 11. pp. 459-462. Shi, J. & Guo, J. (2006) “Distribution and migration of 95Zr in a tea plant/soil system”. Journal of Environmental Radioactivity, 87, pp. 170-174.

Shields, P. M. & Tajalli, H. (2006) “Intermediate theory: the missing link in successful student scholarship”. Department of Political Science Faculty Publications, Texas State University, USA. http://ecommons.txstate.edu/polsfacp/39 (accessed 13/03/2009).

Shinonaga, T., Schimmack, W. & Gerzabek, M. H. (2005) “Vertical migration of 60

Co, 137Cs and 226Ra in agricultural soils as observed in lysimeters under crop rotation”.


269

Shipitalo, M. J. & Protz, R. (1988) “Factors influencing the dispersibility of clay in worm casts”. Soil Science Society of America Journal, 52, pp. 764-769.

Shmulevich, I., Asaf, Z. & Rubinstein, D. (2007) “Interaction between soil and a wide cutting blade using the discrete element method”. Soil & Tillage Research, 97, pp. 3750.

Shor, J. T. & Dial, C. (2000) “Volatilization of cobalt, technetium, and uranium isotopes from soils through salt amendment and calcination”. Journal of Environmental Radioactivity, 48, pp. 35-48.

Showler, A. T., Knaus, R. M. & Reagan, T. E. (1988) “The versatility of radiotracer methods for studying insect ethology and ecology”. Florida Entomologist, 71, pp. 554580.

Shuster, W. D., Subler, S. & McCoy, E. L. (2000) “Foraging by deep-burrowing earthworms degrades surface soil structure of a Fluventic Hapludoll in Ohio”. Soil & Tillage Research, 54, pp. 179-189.

Silver, L. R. (1987) “Fallout from Chernobyl”. Deneau Publishers & Company Ltd. Ontario, Canada.

Simm, D. J. (1993) “The deposition and storage of suspended sediment in contemporary floodplain systems: a case study of the River Culm, Devon, UK”. Unpublished PhD. Thesis, Department of Geography, University of Exeter, UK.

Simm, D. J. & Walling, D. E. (1998) “Lateral variability of overbank sedimentation on a Devon floodplain”. Hydrological Sciences, 43, pp. 715-732.

Simm, D. J., Walling, D. E., Bates, P. D. & Anderson, M. G. (1997) “The potential of finite element modelling of flood plain inundation to predict patterns of overbank deposition”. Hydrological Sciences, 42. pp. 859-875.

Singer, M. J. & Shainberg, I. (2004) “Mineral soil surface crusts and wind and water erosion”. Earth Surface Processes and Landforms, 29, pp. 1065-1075. 270

Six, J., Elliott, E. T. & Paustian, K. (2000) “Soil macroaggregate turnover and microaggregate formation: a mechanism for C sequestration under no-tillage agriculture”. Soil Biology and Biochemistry, 32, pp. 2099-2103.

Slattery, M. C. & Burt, T. P. (1995) “Size characteristics of sediment eroded from agricultural soil: Dispersed versus non-dispersed, ultimate versus effective”. River Geomorphology, Chpt. 1, pp. 1-17. John Wiley & Sons Ltd., Chichester.

Slattery, M. C., Gares, P. A. & Phillips, J. D. (2002) “Slope-channel linkage and sediment delivery on North Carolina coastal plain cropland”. Earth Surface Processes and Landforms, 27, pp. 1377-1387.

Smith, J. T., Appleby, P. G., Hilton, J. & Richardson, N. (1997) “Inventories and fluxes of 210Pb,

137

Cs and 241Am determined from the soils of three small catchments in

Cumbria, UK”. Journal of Environmental Radioactivity, 37, pp. 127-142.

Souchère, V., Cerdan, O., Ludwig, B., Le Bissonnais, Y., Couturier, A. & Papy, F. (2003) “Modelling ephemeral gully erosion in small cultivated catchments”. Catena, 50, pp. 489-505.

Spargo, J. T., Alley, M. M., Follett, R. F. & Wallace, J. V. (2008) “Soil carbon sequestration with continuous no-till management of grain cropping systems in the Virginia coastal plain”. Soil & Tillage Research, 100, pp. 133-140. Sparks, D. L. (2003) “Environmental Soil Chemistry 2nd ed.”. Academic Press Publishing: London.

Spezzano, P. (2005) “Distribution of pre- and post-Chernobyl radiocaesium with particle size fractions of soils”. Journal of Environmental Radioactivity, 83, pp. 117127.

Spokas, K., Forcella, F., Archer, D. & Reicosky, D. (2007) “SeedChaser: vertical soil tillage distribution model”. Computers and Electronics in Agriculture, 57, pp. 62-73.

271

Spoor, G. & Godwin, R.J. (1978) “An experimental investigation into the deep loosening of soil by rigid tines”. Journal of Agricultural Engineering Research, 23, pp. 243–258.

Stone, P. M. (1996) “The Particle Size Selectivity of Suspended Sediment Delivery from Drainage Basins”. Unpublished PhD. Thesis, Department of Geography, University of Exeter, UK.

Stone, P. M. & Walling, D. E. (1996) “The particle-size selectivity of sediment mobilization from Devon hillslopes”. In: Advances in Hillslope Processes (eds. M. G. Anderson & S. M. Brooks), 1, pp. 507-527. John Wiley & Sons Ltd., Chichester.

Stroosnijder, L. (2005) “Measurement of erosion. Is it possible”? Catena, 64, pp. 162173.

Strudley, M. W., Green, T. R. & Ascough, J. C. (2008) “Review: Tillage effects on soil hydraulic properties in space and time: State of the science”. Soil & Tillage Research, 99, pp. 4-48.

Suárez, J. A., Rodríguez, M., Espartero, A. G. & Piña, G. (2000) “Radiochemical analysis of 41Ca and 45Ca”. Applied Radiation and Isotopes, 52, pp. 407-413.

Summerfield, M. A. (1991) “Global Geomorphology: An Introduction to the Study of Landforms”. Prentice Hall: London.

Sutherland R. A., (1991) “Examination of caesium-137 areal activities in control (uneroded) locations”. Soil Technology, 4, pp. 33-50. Sutherland, R. A. (1994) “Spatial variability of 137Cs and the influence of sampling on estimates of sediment redistribution”. Catena, 21, pp. 57-71.

Sutherland, R. A. (1996) “Caesium-137 soil sampling and inventory variability in reference locations: A literature survey”. Hydrological Processes, 10, pp. 43-53.

272

Sutherland, R. A., Wan, Y., Lee, C. T. & Zeigler, A. D. (1996) “Aggregate enrichment ratios for splash and wash transported sediment from an Oxisol”. Catena, 26, pp. 187-208.

Syers, J. K. & Springett, J. A. (1983) “Earthworm ecology in grassland soils”. Chpt. 7, In: Earthworm Ecology: From Darwin to Vermiculture” (ed. J. E. Satchell). Chapman & Hall: London.

Syversen, N., Øygarden, L. & Salbu, B. (2001) “Cesium-134 as a tracer to study particle transport processes within a small catchment with a buffer zone”. Journal of Environmental Quality, 30, pp. 1771-1783.

Thompson, J., Dyer, F. M. & Croudace, I. M. (2002) “Records of radionuclide deposition in two salt marshes in the United Kingdom with contrasting redox and accumulation conditions”. Geochimica et Cosmochimica Acta, 66, pp. 1011-1023.

Thorén, M., Rähn, T., Guo, W. Y. & Werner, S. (1991) “Stereotactite radiosurgery with the cobalt-60 gamma unit in the treatment of growth hormone pituitary tumors”. Neurosurgery, 29, pp. 663-668.

Tiessen, K. H. D., Mehuys. G. R., Lobb, D. A. & Rees, H. W. (2007a) “Tillage erosion within potato production systems in Atlantic Canada I. Measurement of tillage translocation by implements used in seedbed preparation”. Soil & Tillage Research, 95, pp. 308-319.

Tiessen, K. H. D., Lobb, D. A., Mehuys, G. R. & Rees, H. W. (2007b) “Tillage erosion within potato production in Atlantic Canada: II Erosivity of primary and secondary tillage operations”. Soil & Tillage Research, 95, pp. 320-331.

Tiwari, S. C., Tiwari, B. K. & Mishra, R. R. (1989) “Microbial populations, enzyme activities and nitrogen-phosphorous-potassium enrichment in earthworm casts and in the surrounding soil of a pineapple plantation”. Biology & Fertility of Soils, 8, pp. 178182.

273

Toth, S. J. & Alderfer, R. B. (1960a) “A procedure for tagging water-stable aggregates with Co-60”. Soil Science, 89, pp. 36-37.

Toth, S. J. & Alderfer, R. B. (1960b) “Formation and breakdown of Co-60-tagged water-stable aggregates in a Norton Silt Loam Soil. Soil Science, 90, pp. 232-238.

Toy, T. J., Foster, G. R. & Renard, K. G. (2002) “Soil Erosion: Processes, Prediction, Measurement and Control”. John Wiley and Sons Ltd., New York.

Trimble, S. W. (1999) “Decreased rates of alluvial sediment storage in the Coon Creek, Wisconsin, 1975-93”. Science, 285, pp. 1244-1246.

Tsara, M., Gerontidis, S., Marathianou, M. & Kosmas, C. (2001) “The longterm effect of tillage on soil displacement of hilly areas used for growing wheat in Greece”. Soil Use and Management, 17, pp. 113-120.

Turkelboom, F., Poesen, J., Ohler, I., Van Keer, K., Ongprasert, S. & Vlassak, K. (1997) “Assessment of tillage erosion rates on steep slopes in northern Thailand”. Catena, 29, pp. 29-44.

Twining, J. R., Payne, T. E. & Itakura, T. (2004) “Soil-water distribution coefficients and plant transfer factors for

134

Cs,

85

Sr and

65

Zn under field conditions in tropical

Australia”. Journal of Environmental Radioactivity, 71, pp. 71-87.

Tyler, A. N. (1999) “Monitoring anthropogenic radioactivity in salt marsh environments through in situ gamma-ray spectrometry”. Journal of Environmental Radioactivity, 45, pp. 235-252.

Tyler, A. N., Sanderson, D. C. W. & Scott, E. M. (1996a) “Estimating and accounting for

137

Cs source burial through in-situ gamma spectrometry in salt marsh

environments”. Journal of Environmental Radioactivity, 33, pp. 195-212.

Tyler, A. N., Sanderson, D. C. W., Scott, E. M. & Allyson, J. D. (1996b) “Accounting for spatial variability and fields of view in environmental gamma ray spectrometry”. Journal of Environmental Radioactivity, 33, pp. 213-235. 274

Tyler, A. N., Carter, S., Davidson, D. A., Long, D. J. & Tipping, R. (2001) “The extent and significance of bioturbation in upland soils”. Catena, 43, pp. 81-99.

Tyson, M. J., Sheffield, E. & Callaghan, T. V. (1999) “Uptake, transport and seasonal recycling of

134

Cs applied experimentally to bracken (Pteridium aquilinum L. Kuhn).


Ussiri, D. A. N. & Lal, R. (2009) “Long-term tillage effects on soil carbon storage and carbon dioxide emissions in continuous corn cropping system from an Alfisol in Ohio”. Soil & Tillage Research, 104, pp. 39-47.

Valentin, C. & Bresson, L. M. (1992) “Morphology, genesis and classification of surface crusts in loamy and sandy soils”. Geoderma, 55, pp. 225-245.

Van Hoof, P. (1983) “Earthworm activity as a cause of splash erosion in a Luxembourg forest”. Geoderma, 31, pp. 195-204.

Van Muysen, W., Van Oost, K. & Govers, G. (2006) “Soil translocation resulting from multiple passes of tillage under normal field operating conditions”. Soil & Tillage Research, 87, pp. 218-230.

Van Muysen, W., Govers, G., Bergkamp, G., Roxo, M. & Poesen, J. (1999) “Measurement and modeling of the effects of initial soil conditions and slope gradient on soil translocation by tillage”. Soil & Tillage Research, 51, pp. 303-316.

Van Muysen, W., Govers, G., Van Oost, K. & Van Rompaey, A. (2000) “The effect of tillage depth, tillage speed and soil condition on chisel tillage erosivity”. Journal of Soil and Water Conservation, 55, pp. 355-364.

Van Oost, K., Govers, G., de Alba, S. & Quine, T. A. (2006) “Tillage erosion: a review of controlling factors and implications for soil quality”. Progress in Physical Geography, 30, pp. 443-466.

275

Van Oost, K., Govers, G., Van Muysen, W. & Quine, T. A. (2000) “Modeling translocation and dispersion of soil constituents by tillage on sloping land”. Soil Science Society of America Journal, 64, pp. 1733-1739.

Van Oost, K., Quine, T. A., Govers, G., De Gryze, S., Six, J., Harden, J. W., Ritchie, J. C., McCarty, G. W., Heckrath, G., Kosmas, C., Giraldez, J. V., Marques da Silva, J. R. & Merckx, R. (2007) “The impact of agricultural soil erosion on the global carbon cycle”. Science, 318, pp. 626-629.

Van Oost, K., Van Muysen, W., Govers, G., Deckers, J. & Quine, T. A. (2005) “From water to tillage erosion dominated landform evolution”. Geomorphology, 72, pp. 193-203.

Ventura, E., Nearing, M. & Norton, L. D. (2001) “Developing a magnetic tracer to study soil erosion”. Catena, 43, pp. 277-291.

Ventura, E., Nearing, M. A., Amore, E. & Norton, L. D. (2002) “The study of detachment and deposition on a hillslope using a magnetic tracer”. Catena, 48, pp. 149161.

Wallbrink, P. J. & Murray, A. S. (1993) “Use of fallout radionuclides as indicators of erosion processes”. Hydrological Processes, 7, pp. 297-304.

Wallbrink, P. J., Murray, A. S. & Olley, J. M. (1999) “Relating suspended sediment to its original soil depth using fallout radionuclides”. Soil Science Society of America Journal, 63, pp. 369-378.

Wallbrink, P. J., Olley, J. M. & Hancock, G. (2002) “Estimating residence times of fine sediment in river channels using fallout

210

Pb”. In: The Structure, Function and

Management Implications of Fluvial Sedimentary Systems. International Association of Hydrological Sciences (IAHS), Publication No. 276, pp. 425-432.

Wallbrink, P. J., Walling, D. E. & He, Q. (2002) “Radionuclide Measurement Using HPGe Gamma Spectrometry” Cp. 5, pp. 67-96. In: F. Zapata (ed.) (2002) Handbook for

276

the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides. Kluwer Academic Publishers: London.

Walling, D. E. (1983) “The sediment delivery problem”. Journal of Hydrology, 65, pp. 209-237.

Walling, D. E. (1990) “Linking the field to the river: Sediment delivery from agricultural land”. In: Soil Erosion on Agricultural Land (eds. J. Boardman, I. D. L. Foster, & J. A. Dearing). John Wiley & Sons Ltd. Chichester, pp. 129-152.

Walling, D. E. (1999) “Linking land use, erosion and sediment yields in river basins”. Hydrobiologia, 410, pp. 223-240.

Walling, D. E. (2004) “Quantifying the fine sediment budgets of river basins”. National Hydrology Seminar, pp. 9-20.

Walling, D. E. (2006) “Tracing versus monitoring: new challenges and opportunities in erosion and sediment delivery research”, Chpt. 2. In. Soil Erosion and Sediment Redistribution in River Catchments: Measuring, Monitoring and Management” (eds. P. N. Owens & A. J. Collins). CABI Publishing, Oxfordshire, UK.

Walling, D. E. (2008) “The changing sediment loads of the world’s rivers”. In: Sediment Dynamics in Changing Environments. International Association of Hydrological Sciences (IAHS), Publication No. 325, pp. 323-338.

Walling, D. E. & Bradley, S. B. (1988) “The use of caesium-137 measurements to investigate sediment delivery from cultivated areas in Devon, UK”. In: Sediment Budgets. International Association of Hydrological Sciences (IAHS), Publication No. 174, pp. 325-335.

Walling, D. E. & Bradley, S. B. (1989) “Rates and patterns of contemporary floodplain sedimentation: A case study of the River Culm, Devon, UK”. GeoJournal, 19, pp. 5362.

277

Walling, D. E. & Collins, A. L. (2008) “The catchment sediment budget as a management tool”. Environmental Science & Policy, 11, pp. 136-143.

Walling, D. E., Collins, A. L., Jones, P. A., Leeks, G. J. L. & Old, G. (2006) “Establishing fine-grained sediment budgets for the Pang and Lambourn LOCAR catchments, UK”. Journal of Hydrology, 330, pp. 126-141.

Walling, D. E., Golosov, V. N., Panin, A. V. & He, Q. (2000) “Use of radiocaesium to investigate erosion and sedimentation in areas with high levels of Chernobyl fallout”. In Tracers in Geomorphology (ed. I. D. L. Foster), Chpt. 9, pp. 183-200. John Wiley & Sons Ltd. Chichester, UK.

Walling, D. E. & He, Q. (1993) “Use of cesium-137 as a tracer in the study of rates and patterns of floodplain sedimentation”. In: Tracers in Hydrology. International Association of Hydrological Sciences (IAHS), Publication No. 215, pp. 319-328.

Walling, D. E. & He, Q. (1994) “Rates of overbank sedimentation on the floodplains of several British rivers during the past 110 years”. In: Variability in Stream Erosion and Sediment Transport. International Association of Hydrological Sciences (IAHS) Publication, No. 224, pp. 77-86.

Walling, D. E. & He, Q. (1997) “Investigating spatial patterns of overbank sedimentation on river floodplains”. Water, Air and Soil Pollution, 99, pp. 9-20.

Walling, D. E. & He, Q. (1998) “The spatial variability of overbank sedimentation on river floodplains”. Geomorphology, 24, pp. 209-223.

Walling, D. E. & He, Q. (1999) “Improved models for estimating soil erosion rates from cesium-137 measurements”. Journal of Environmental Quality, 28, pp. 611-622.

Walling, D. E., He, Q. & Appleby, P. G. (2002) “Conversion models for use in soilerosion, soil-redistribution and sedimentation investigations”, Chp. 7, pp. 111-164. In: Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides (ed. F Zapata). Kluwer Academic Publishers, London, UK.

278

Walling, D. E., He. Q. & Blake, W. (1999) “Use of 7Be and

137

Cs measurements to

document short- and medium-term rates of water-induced soil erosion on agricultural land”. Water Resources Research, 35, pp. 3865-3874.

Walling, D. E. & Owens, P. N. (2002) “The role of flood plain sedimentation in catchment sediment and contaminant budgets”. In: The Structure, Function and Management Implications of Fluvial Sedimentary Systems. International Association of Hydrological Sciences (IAHS), Publication No. 276, pp. 407-416.

Walling, D. E., Owens, P. N., Carter, J., Leeks, G. J. L., Lewis, S., Meharg, A. A. & Wright, J. (2003) “Storage of sediment-associated nutrients and contaminants in river channel and floodplain systems”. Applied Geochemistry, 18, pp. 195-220.

Walling, D. E., Owens, P. N. & Leeks, G. J. L. (1997). “The characteristics of overbank deposits associated with a major flood event in the catchment of the River Ouse, Yorkshire, UK”. Catena, 31, pp. 53-75.

Walling, D. E., Owens, P. N. & Leeks, G. J. L. (1998) “The role of channel and floodplain storage in the suspended sediment budget of the River Ouse, Yorkshire, UK”. Geomorphology, 22, pp. 225-242.

Walling, D. E. & Quine, T. A. (1990) “Calibration of caesium-137 measurements to provide quantitative erosion rate data”. Land Degradation & Rehabilitation, 2, pp. 161175. Walling, D. E. & Quine, T. A. (1991) “Use of 137Cs measurements to investigate soil erosion on arable fields on the UK: potential applications and limitations”. Journal of Soil Science, 42, pp. 147-165.

Walling, D. E. & Quine, T. A. (1992) “The use of caesium-137 measurements in soil erosion surveys”. In: Erosion and Sediment Transport Monitoring Programmes in River Basins. International Association of Hydrological Sciences (IAHS), Publication No. 210, pp. 143-152.

279

Walling, D. E., Rowan, J. S. & Bradley, S. B. (1989) “Sediment-associated transport and redistribution of Chernobyl fallout radionuclides”. In: Sediment and the Environment. International Association of Hydrological Sciences (IAHS), Publication No. 184, pp. 37-45.

Walling, D. E., Schuller, P., Zhang, Y. & Iroumé, A. (2009) “Extending the timescale for using beryllium 7 measurements to document soil redistribution by erosion”. Water Resources Research, 45, W02418.

Walling, D. E. & Woodward, J. C. (1992) “Use of radiometric fingerprints to derive information on suspended sediment sources”. In: Erosion and Sediment Transport Monitoring Programmes in River Basins. International Association of Hydrological Sciences (IAHS), Publication No. 210, pp. 153-164.

Wang, C. H., Willis, D. L. & Loveland, W. D. (1975) “Radiotracer Methodology in the Biological, Environmental, and Physical Sciences”, (eds. W. D. McElroy & Swanson, C. P.). Prentice-Hall, New Jersey: USA.

Warren, S. D., Nevill, M. B., Blackburn, W. H. & Garza, N. E. (1986a) “Soil response to trampling under intensive rotation grazing”. Soil Science Society of America Journal, 50, pp. 1336-1341.

Warren, S. D., Thurow, T. L., Blackburn, W. H. & Garza, N. E. (1986b) “The influence of livestock trampling under intensive rotation grazing on soil hydraulic characteristics”. Journal of Range Management, 39, pp. 491-495.

Wentworth, C. K. (1922) “A scale of grade and class terms for clastic sediments”. Journal of Geology, 30, pp.377–392.

Weyhenmeyer, G. A., Håkanson, L. & Meili, M. (1997) “A validated model for daily variations in the flux, origin, and distribution of settling particles within lakes”. Journal of Limnological Oceanography, 42, pp. 1517-1529.

White, R. E. (1997) “Principles and Practice of Soil Science: The soil as a natural resource 3rd ed”. Blackwell Science, London, UK. 280

Whiting, P. J., Bonniwell, E. C. & Matisoff, G. (2001) “Depth and areal extent of sheet and rill erosion based on radionuclides in soils and suspended sediment”. Geology, 29, pp. 1131-1134.

Wicker, R. D. & Ibrahim, S. A. (2006) “Vertical migration of

134

Cs bearing soil

particles in arid soils: implications for plutonium redistribution”. Journal of Environmental Radioactivity, 88, pp. 171-188.

Willatt, S. T., and Pullar, D. M. (1984). “Changes in soil physical properties under grazed pastures”. Australian Journal of Soil Research, 22, pp. 343–348.

Wilson, C. G., Matisoff, G. & Whiting, P. J. (2003) “Short-term erosion rates from a 7

Be inventory balance”. Earth Surface Processes and Landforms, 28, pp. 967-977.

Wolman, M. G. (1977) “Changing needs and opportunities in the sediment field”. Water Resources Research, 13, pp. 50-54.

Wood, F. L., Heathwaite, A. L. & Haygarth, P. M. (2005) “Evaluating diffuse and point phosphorous contributions to river transfers at different scales in the Taw catchment, Devon, UK”. Journal of Hydrology, 304, pp. 118-138.

Wooldridge, D. D. (1965) “Tracing soil particle movement with Fe-59”. Soil Science Society of America Journal, 29, pp. 469-472.

Yair, A. & Lavee, H. (1981) “An investigation of source areas of sediment transport by overland flow along arid hillslopes”. In: Erosion and Sediment Transport Measurement. International Association of Hydrological Sciences (IAHS) Publication, No. 133, pp. 433-446.

Yang, M. Y., Walling, D. E., Tian, J. L. & Liu, P. L. (2006) “Partitioning the contributions of sheet and rill erosion using Beryllium-7 and Cesium-137”. Soil Science Society of America Journal, 70, pp. 1579-1590.

281

Yenish, J. P., Doll, J. D. & Buhler, D. D. (1992) “Effects of tillage on vertical distribution and viability of weed seed in soil”, Weed Science, 40, pp. 429-433.

Young, R. A. & Onstad, C. A. (1978) “Characterization of rill and interrill eroded soil”. Transactions of the American Society of Agricultural Engineers, 21, pp. 11261130.

Young, R. A. (1980) “Characteristics of eroded sediment” Transactions of the American Society of Agricultural Engineers, 23, pp. 1139-1142.

Zapata, F. (2003) “The use of environmental radionuclides as tracers in soil erosion and sedimentation investigations: recent advances and future developments”. Soil & Tillage Research, 69, pp. 3-13.

Zapata, F. & Garcia-Agudo, E. (2000) “Future prospects for the

137

Cs technique for

estimating soil erosion and sedimentation rates”. Acta Geologica Hispanica, 35, pp. 197-205

Zapata, F., Garcia-Agudo, E., Ritchie, J. C. & Appleby, P. G. (2002) “Introduction”, Chp. 1. pp. 1-13. In: Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides (ed. F Zapata). Kluwer Academic Publishers, London, UK.

Zartl, A. S., Klik, A. & Huang, C. (2001) “Soil detachment and transport processes from inter-rill and rill areas”. Physics and Chemistry of the Earth (B), 26, pp 2.

Zehnder, H. J., Kopp, P., Eikenberg, J., Feller, U. & Oertli, J. J. (1995) “Uptake and transport of radioactive cesium and strontium into grapevines after leaf contamination”. Radiation and Physical Chemistry, 46, pp. 61-69.

Zhang, X. C., Friedrich, J. M., Nearing, M. A. & Norton, L. D. (2001) “Potential use of rare earth oxides as tracers for soil erosion and aggregation studies”. Soil Science Society of America Journal, 65, pp. 1508-1515.

282

Zhang, X., Higgitt, D. L. & Walling, D. E. (1990) “A preliminary assessment of the potential for using caesium-137 to estimate rates of soil erosion in the Loess Plateau of China”. Hydrological Sciences Journal, 35, pp. 243-251.

Zhang, X. C., Nearing, M. A., Polyakov, V. O. & Friedrich, J. M. (2003) “Using rare-Earth oxide tracers for studying soil erosion dynamics”. Soil Science Society of America Journal, 67, pp. 279-288.

Zhang, H. & Schrader, S. (1993) “Earthworm effects on selected physical and chemical properties of soil aggregates”. Biological Fertility of Soils, 15, pp. 229-234.

Zhang, X., Walling, D. E., Quine, T. A. & Wen, A. (1997) “Use of reservoir deposits and caesium-137 measurements to investigate the erosional response of a small drainage basin in the rolling Loess Plateau region of China”. Land Degradation & Development, 8, pp. 1-16.

Zombori, P., Andrási, A. & Németh, I. (1992a) “A new method for the determination of radionuclide distribution in the soil by in situ gamma ray spectrometry”. Central Research Institute for Physics, Institute for Atomic Energy Research, Hungary.

Zombori, P., Németh, I., Andrási, A. & Lettner, H. (1992b) “In-situ gammaspectrometric measurement of the contamination in some selected settlements of Byelorussia (BSSR), Ukraine (UkrSSR) and the Russian Federation (RSFSR)”. Journal of Environmental Radioactivity, 17, pp. 97-106.

283