THE ERODIBILITY AND PHOSPHORUS LOSS POTENTIAL OF A SELECTION OF IRISH

THE ERODIBILITY AND PHOSPHORUS LOSS POTENTIAL OF A SELECTION OF IRISH
NATIONAL UNIVERSITY OF IRELAND, GALWAY
THE ERODIBILITY AND PHOSPHORUS LOSS
POTENTIAL OF A SELECTION OF IRISH
TILLAGE SOILS
John T. Regan B.E.
Research Supervisors:
Dr. Mark G. Healy, Civil Engineering, NUI Galway
Dr. Owen Fenton, Teagasc Johnstown Castle, Wexford
Professor of Civil Engineering: Padraic E. O‟Donoghue
Thesis submitted in fulfilment of the requirements for the degree of Doctor of
Philosophy.
August 2012
The National University of Ireland requires the signatures of all persons using or
photocopying this thesis. Please sign below, and give the address and date.
i
For my parents
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“Tell me and I'll forget; show me and I may remember; involve me and I'll understand.”
Chinese proverb
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ABSTRACT
Arable cropping, due to its intensive nature, can leave soil with reduced ground cover and
impaired soil structure, making it vulnerable to erosion under heavy rainfall. Runoff
containing suspended sediment (SS) and nutrients, particularly phosphorus (P), from
agricultural fields is considered to be one of the main causes of water quality impairment.
To date, in Ireland, no study has investigated erosion and associated P loss from tillage
soils when subjected to high intensity rainfall, even though international research
indicates significantly higher P export coefficients from this land use than from grassland.
As a result, only agronomic nutrient advice is available and this has been adopted into
current legislation. Research was therefore necessary to assess the potential P losses
arising from complying with the legislation. This objective was addressed in the first part
of the study using two simulated rainfall experiments. A related objective involving the
development of a screening methodology to identify tillage fields with erosion risk and
soil quality problems was addressed in the second part.
The aim of the first part of the study was to quantify the amount of dissolved reactive
phosphorus (DRP), total phosphorus (TP), particulate phosphorus (PP) and SS released in
runoff from five tillage soils with varying soil test P (STP) when subjected to a rainfall
intensity of 30 mm hr -1 applied in three successive events. Soil physical and chemical
parameters, slope, and time interval between storm events were ranked in order of
importance for the prediction of P and SS releases, and a runoff dissolved phosphorus
risk indicator (RDPRI) was developed to identify the STP for Irish tillage soils above
which there may be a potential threat to surface water quality. The effect of soil type on
the flow weighted mean concentration (FWMC) of DRP (p = 0.013) depended on both
slope and time between rainfall events. The effect of soil type depended only on surface
slope for the FWMCs of SS (p = 0.044), TP (p = 0.014) and PP (p = 0.022) in surface
runoff. Increasing the overland flow rate over the soil surface in the presence of rainfall
had the effect of increasing the concentrations of SS, PP and TP (but not DRP) in surface
runoff (p < 0.05) across all soils. An increase in extractable soil P resulted in an increase
in concentrations of DRP in surface runoff (p < 0.05) across all soils. Of the five methods
used to extract soil P in these experiments, water extractable P (WEP) was identified as
having the greatest potential to be used as an indicator of the risk of P movement from
soil into runoff water. However, despite its apparent advantage over Morgan‟s
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Phosphorus (Pm) in determining environmental risk, it would appear to be impractical and
costly to run two soil P tests side by side given that P m gives a good approximation for
both agronomic and environmental purposes.
Combining the results of both experiments (rainfall only and rainfall and overland flow)
indicated that if the current agronomic optimum P m range for tillage soils of 6.1 - 10 mg
L-1 is maintained by tillage farmers through adherence to recommended P application
rates, then the risk of runoff with DRP concentrations in excess of the level at which
eutrophication is likely to occur (0.03 mg molybdate reactive phosphorus (MRP) L -1)
should be minimal.
Ireland has a valuable resource in terms of its land and soil quality, and promoting
sustainable soil management is one of the areas of action included in Food Harvest 2020,
the national strategy for the development of the agri-food sector. Agricultural activities
that can negatively impact on soil quality must be tackled if Ireland is to meet the
ambitious growth targets set out in this vision. Therefore, in the second part of the study,
the five soils above, and a sixth soil, were assessed in their natural field conditions to
determine their erosion risk and soil quality status. At each study site, detailed soil
classification results and visual soil assessments were used in conjunction with observed
erosion levels to select the most appropriate indicators for assessing erosion risk and soil
quality status. Parameters selected include: texture, slope, erosion features, structure,
ponding, potential rooting depth, soil organic matter (SOM), average annual rainfall and
current land use. Assessment of these indicators at each study site using the user friendly
grading system developed here, made it possible to correctly identify the sites where the
erosion risk was high. This work showed that adoption of an erosion and soil quality
screening method by tillage farmers and advisory specialists in Ireland will enable the
quantification of the extent of erosion risk and soil quality degradation on farms. This
will allow remediation measures to be prioritised for the most vulnerable sites, which is
likely to result in cost and resource savings for farmers and advisory services.
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DECLARATION
This dissertation is the result of my own work, except where explicit reference is made to
the work of others, and has not been submitted for another qualification to this or any
other university.
John Regan
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ACKNOWLEDGEMENTS
I wish to thank an tOllamh Padraic O‟Donoghue for his encouragement and help during
the research study and during the preparation of this thesis.
I am grateful to Teagasc for providing a Walsh fellowship and funding for this project,
and to the authorities in the National University of Ireland, Galway for providing the
facilities to carry out the research work.
I would like to express my most sincere gratitude to my NUI, Galway supervisors, Dr.
Mark Healy and Dr. Michael Rodgers and my Teagasc supervisors, Dr. Owen Fenton and
the late Dr. John Mulqueen, who were always there to guide me, and who were always
generous with their time. I owe them my deepest gratitude.
I have been very fortunate and received help from many people along the way. I would
however like to express a special thanks to Drs. Michael Walsh, Jim Grant and Laura
Kirwan for their guidance and patience. This work would not have been possible without
tremendous cooperation and help from many people, in particular technical staff in NUI
Galway. A special thanks to Mary O‟Brien, Maja Drapiewska, Peter Fahy, and Gerry
Hynes.
I would like to thank my friends for all their encouragement throughout these years.
I would like to thank my girlfriend, Chloé for her support, encouragement and for being
there when I needed her.
Finally, I would like to thank my parents, Sean and Anne, and my brother, Colum, who
have helped me in more ways than they will ever know.
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ABBREVIATIONS
AAR
Average annual rainfall
ACP
Agricultural catchments programme
AgSt
Aggregate stability
AIC
Aikake information criterion
Al
Aluminium
Alox
Ammonium oxalate-oxalic acid extractable aluminium
Be
Beryllium
C
Carbon
Ca
Calcium
CaCO3
Calcium carbonate
CEC
Cation exchange capacity
CORINE
CO-oRdination of INformation on the Environment
CO 2
Carbon dioxide
CSA
Critical source area
Cs
Caesium
CSO
Central Statistics Office
DAFF
Department of Agriculture, Fisheries and Food
Defra
Department for Environment, Food and Rural Affairs
DEHLG
Department of the Environment, Heritage and Local Government
DRP
Dissolved reactive phosphorus
EEA
European Environment Agency
EPA
Environmental Protection Agency
EQS
Environmental quality standards
Fe
Iron
Feox
Ammonium oxalate-oxalic acid extractable iron
FWMC
Flow-weighted mean concentration
GIS
Geographical information systems
GLMM
Generalized linear mixed model
GOPC
Grid orientated P component
ha
Hectare
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HSPF
Hydrological simulation program - FORTRAN
ISIS
Irish soil information system
ISO
International organization for standarization
LMM
Linear mixed model
Mg
Magnesium
MRP
Molybdate reactive P
M3-P
Mehlich-3 phosphorus
N
Nitrogen
Na
Sodium
NH4-N
Ammonium nitrogen
NO3-N
Nitrate
NPRP
National phosphorus research project
OM
Organic matter
OSR
Oilseed rape
P
Phosphorus
Pb
Lead
Pcacl2
Calcium chloride extractable P
Pm
Morgan‟s P
POM
Programme of measures
Pox
Ammonium oxalate-oxalic acid extractable P
PP
Particulate P
PRD
Potential rooting depth
Psatox
Soil P saturation
PSC
Phosphorus sorption capacity
PSD
Particle size distribution
RBMP
River basin management plans
RDPRI
Runoff dissolved P risk indicator
RUSLE
Revised Universal Soil Loss Equation
SAS
Statistical analysis software
SDR
Sediment delivery ratio
SEDD
Sediment distribution delivery
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SERBD
South eastern river basin district
SFD
Soil framework directive
SHETRAN
Système hydrologique Européen TRANsport
SI
Statutory instrument
SOC
Soil organic carbon
SoCo
Sustainable agriculture and soil conservation
SOM
Soil organic matter
Sq
Soil structure quality as scored by VESS
SS
Suspended sediment
STP
Soil test P
SWAT
Soil water assessment tool
TDP
Total dissolved P
TRP
Total reactive P
TP
Total P
T1/2
Half-life
USLE
Universal soil loss equation
VESS
Visual evaluation of soil structure
VSA
Visual soil assessment
WEP
Water extractable phosphorus
WFD
Water Framework Directive
α
Empirical parameter used to relate P sorption capacity to (Alox + Feox)
β
Catchment specific modeling parameter
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TABLE OF CONTENTS
ABSTRACT ................................................................................................................ IV
DECLARATION ........................................................................................................ VI
ACKNOWLEDGEMENTS ...................................................................................... VII
ABBREVIATIONS.................................................................................................. VIII
LIST OF TABLES.................................................................................................. XVII
TABLE OF FIGURES ............................................................................................. XIX
CHAPTER 1 INTRODUCTION .................................................................................. 1
1.1 OVERVIEW ............................................................................................................ 1
1.2 OBJECTIVES .......................................................................................................... 3
1.3 PROCEDURE........................................................................................................... 3
1.4 STRUCTURE OF DISSERTATION .............................................................................. 5
CHAPTER 2 LITERATURE REVIEW ....................................................................... 7
2.1 INTRODUCTION...................................................................................................... 7
2.2 SOIL EROSION ...................................................................................................... 11
2.2.1 Factors effecting soil erosion rates in Ireland ............................................. 13
2.2.1.1 Soil type .......................................................................................... 14
2.2.1.2 Precipitation (amount, duration and intensity) ................................ 15
2.2.1.3 Management practices .................................................................... 17
2.2.2 Field indicators of a soil‟s susceptibility to runoff and erosion ................... 18
2.2.2.1 Soil texture ..................................................................................... 18
2.2.2.2 Soil colour ...................................................................................... 18
2.2.2.3 Mottling .......................................................................................... 19
2.2.2.4 Soil structure and consistence ......................................................... 20
2.2.2.5 Soil porosity and root development ................................................. 20
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2.2.3 Main soil properties controlling soil erosion at field-scale .......................... 22
2.2.3.1 Particle size distribution ................................................................. 22
2.2.3.2 Soil organic carbon/matter ............................................................. 24
2.2.3.3 Bulk density .................................................................................... 25
2.2.3.4 Aggregate stability .......................................................................... 26
2.2.4 Impact of tillage farming ............................................................................ 28
2.2.5 Processes and mechanics of erosion and deposition .................................... 32
2.2.5.1 Raindrop erosion ............................................................................ 32
2.2.5.2 Flow-driven erosion........................................................................ 36
2.2.6 Measuring and quantifying soil erosion on arable land ............................... 38
2.3 PHOSPHORUS TRANSFER FROM AGRICULTURAL SOILS TO SURFACE WATER
BODIES ...................................................................................................................... 39
2.3.1 Sensitivity of surface waters to eutrophication ........................................... 39
2.3.2 Soil test phosphorus ................................................................................... 41
2.3.3 Phosphorus use on tillage land ................................................................... 42
2.3.4 Critical source areas of phosphorus ............................................................ 44
2.3.5 Phosphorus mobilisation ............................................................................ 46
2.3.6 Hydrological pathways of phosphorus transport ......................................... 48
2.3.7 Rainfall simulation as phosphorus research tool ......................................... 50
2.4 PHOSPHORUS LOSS RISK ASSESSMENT TOOLS ...................................................... 51
2.4.1 Phosphorus risk index ................................................................................ 51
2.4.2 SCIMAP (Sensitive Catchment Integrated Modelling and Analysis
Platform) ............................................................................................................ 52
2.5 MITIGATION MEASURES TO PREVENT SEDIMENT AND PHOSPHORUS LOSS FROM
TILLAGE SOILS .......................................................................................................... 55
2.5.1 Soil and land management to prevent erosion............................................. 55
2.5.2 Developing phosphorus management guidelines for water quality
protection ........................................................................................................... 58
2.5.3 Catchment-scale research ........................................................................... 59
2.6 FUTURE RESEARCH DIRECTION IN THE QUANTIFICATION OF PHOSPHORUS AND
SEDIMENT LOSS FROM IRISH TILLAGE SOILS ............................................................. 61
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2.6.1 Sediment provenance ................................................................................. 61
2.7 SUMMARY............................................................................................................ 63
CHAPTER 3 DETERMINING PHOSPHORUS AND SEDIMENT RELEASE
RATES FROM FIVE IRISH TILLAGE SOILS WHEN SUBJECTED TO
SIMULATED RAINFALL AND INCREASING OVERLAND FLOW RATES ..... 64
3.1 INTRODUCTION.................................................................................................... 64
3.2 THE TILLAGE SOILS, LABORATORY FLUME SET-UP, AND ANALYSIS METHODS
USED IN THIS STUDY................................................................................................... 66
3.2.1 Soil collection and preparation ................................................................... 66
3.2.2 Simulated rainfall experiment .................................................................... 70
3.2.3 Overland flow experiment .......................................................................... 74
3.2.4 Soil analysis ............................................................................................... 75
3.2.5 Runoff analysis .......................................................................................... 76
3.2.6 Data analysis .............................................................................................. 76
3.2.7 Statistical methods ..................................................................................... 77
3.3 EXPERIMENTAL RESULTS FROM SIMULATED RAINFALL/OVERLAND FLOW
EXPERIMENTS............................................................................................................ 78
3.3.1 Characteristic properties of the study soils ................................................. 78
3.3.2 Suspended sediment and phosphorus concentrations in runoff from
simulated rainfall events ..................................................................................... 79
3.3.3 Suspended sediment and phosphorus concentrations in runoff from
simulated rainfall/overland flow events............................................................... 83
3.3.4 Effect of slope, time between events and overland flow rate on
concentrations in surface runoff .......................................................................... 89
3.4 SUMMARY............................................................................................................ 92
CHAPTER 4 THRESHOLD VALUES AND POTENTIAL RISK INDICATORS
FOR ESTIMATING PHOSPHORUS AND SEDIMENT RELEASE FROM IRISH
TILLAGE SOILS ........................................................................................................ 93
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4.1 INTRODUCTION.................................................................................................... 93
4.2 DATA ANALYSIS METHODS................................................................................... 95
4.3 ANALYSIS OF EXPERIMENTAL RESULTS FROM SIMULATED RAINFALL/OVERLAND
FLOW EXPERIMENTS.................................................................................................. 95
4.3.1 Soil properties affecting phosphorus release from soil to water .................. 95
4.3.2 A method for identifying the critical soil test phosphorus threshold above
which simulated rainfall-induced surface runoff may pose a threat to surface
water quality....................................................................................................... 98
4.3.3 Investigating the effect of increasing overland flow rates on the critical soil
test phosphorus threshold above which runoff may pose a threat to surface water
quality .............................................................................................................. 103
4.3.4 Phosphorus and sediment loss risk indicators for Irish tillage soils subjected
to simulated rainfall .......................................................................................... 109
4.3.5 Phosphorus loss risk indicators for Irish tillage soils subjected to simulated
rainfall/overland flow ....................................................................................... 110
4.4 SUMMARY.......................................................................................................... 111
CHAPTER 5 PHYSICAL, CHEMICAL AND VISUAL EVALUATION OF SIX
IRISH TILLAGE SOILS TO ASSESS SOIL QUALITY AND SUSCEPTIBILITY
TO EROSION............................................................................................................ 113
5.1 INTRODUCTION.................................................................................................. 113
5.2 SITE INFORMATION AND SOIL ASSESSMENT METHODS ....................................... 117
5.2.1 Site selection ............................................................................................ 117
5.2.2 Management history of the sites ............................................................... 118
5.2.3 Methods used to assess erosion risk and soil quality at the study sites ...... 120
5.2.3.1 Field-Scale Defra (2005) Assessment ............................................ 121
5.2.3.2 Visual Soil Assessment .................................................................. 124
5.2.3.3 Visual Evaluation of Soil Structure ............................................... 125
5.2.3.4 Soil Survey of Ireland assessment ................................................. 125
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5.3 EXPERIMENTAL RESULTS FROM EROSION RISK AND SOIL QUALITY ASSESSMENTS
................................................................................................................................ 128
5.3.1 Field-Scale Defra Assessment .................................................................. 128
5.3.2 Visual Soil Assessment ............................................................................ 130
5.3.3 Visual Evaluation of Soil Structure .......................................................... 134
5.3.4 Soil Survey of Ireland assessment ............................................................ 135
5.4 AN EROSION AND SOIL QUALITY SCREENING TOOLKIT FOR IRISH TILLAGE SOILS
................................................................................................................................ 139
5.4.1 Development of the screening toolkit ....................................................... 139
5.4.2 Toolkit testing .......................................................................................... 147
5.5 SUMMARY.......................................................................................................... 150
CHAPTER 6 CONCLUSIONS AND RECOMMENDATIONS ............................. 151
6.1 OVERVIEW ........................................................................................................ 151
6.2 INTEGRATION OF RDPRI FINDINGS WITH RESULTS FROM THE SCREENING
TOOLKIT.................................................................................................................. 151
6.3 CONCLUSIONS ................................................................................................... 152
6.4 RECOMMENDATIONS FOR FUTURE WORK .......................................................... 153
REFERENCES .......................................................................................................... 155
APPENDICES ........................................................................................................... 203
APPENDIX A LIST OF PUBLICATIONS .............................................................. 204
APPENDIX B MODELLING SOIL EROSION AND P AND SS DELIVERY TO
SURFACE WATERS AT THE CATCHMENT SCALE ................ 207
APPENDIX C RELATIONSHIPS BETWEEN DRP IN RUNOFF AND SOIL
EXTRACTABLE P METHODS ...................................................... 214
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APPENDIX D SOIL TEXTURE CLASSIFICATION (DEFRA, 2005;
ENVIRONMENT AGENCY, 2007) ................................................. 218
APPENDIX E VISUAL SOIL ASSESSMENT SCORE CARD, TABLES AND
VISUAL AIDS (SHEPHERD, 2009). ............................................... 219
APPENDIX F FIELD PROCEDURE FOR VISUAL EVALUATION OF SOIL
STRUCTURE (BALL ET AL., 2011; GUIMARÃES ET AL.,
2011) .................................................................................................. 224
APPENDIX G SOIL CLASSIFICATION DATA .................................................... 225
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LIST OF TABLES
Table 2.1 Key tillage operations/practices that may impact on soil and water
quality and possible mitigation options ...................................................... 30
Table 2.2 Phosphorus Index System (from SI 610 of 2010 and adapted from
Schulte et al., 2010a).................................................................................... 42
Table 3.1 Chemical and physical properties of selected Irish tillage soils ................. 69
Table 3.2 Overall Anova for responses from GLMM analyses (rainfall only) .......... 90
Table 3.3 Overall ANOVA for responses from LMM analyses (rainfall and
overland flow).............................................................................................. 91
Table 4.1 Comparing RDPRI outputs (Section 4.3.4) with new model outputs (this
section) to identify thresholds for each of Morgan’s P, Pm; water
extractable phosphorus, WEP; Mehlich-3, M3-P; calcium chloride
extractable phosphorus, Pcacl2; and Soil P saturation, Psatox, above
which DRP in surface runoff may exceed 0.03 mg L-1 (old limit - SI 258
of 1998) and 0.035 mg L-1 (new limit - SI 272 of 2009)............................. 106
Table 5.1 Information on selected tillage sites. ......................................................... 119
Table 5.2 Methods used in visual and tactile assessment ......................................... 120
Table 5.3 Advantages and limitations of respective methods .................................. 121
Table 5.4 Water erosion risk-assessment (adapted from Defra, 2005) .................... 122
Table 5.5 Results of the Defra (2005) erosion risk assessment ................................. 129
Table 5.6 Results of Visual Soil Assessments ............................................................ 131
Table 5.7 Results of Visual Evaluation of Soil Structure ......................................... 135
Table 5.8 K-factor erodibility values and associated risk. ....................................... 138
Table 5.9 Integration of Defra (2005) erosion risk rankings with selected soil
quality indicators ...................................................................................... 140
Table 5.10 Proposed toolkit indicators for use in screening sites for erosion risk
and reduced soil quality. .......................................................................... 142
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Table 5.11 Application of the methodology for upgrading/downgrading Defra
risk classes to the study soils. ................................................................... 149
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TABLE OF FIGURES
Figure 2.1 Polluted river sites surveyed in 2007-2009 grouped by severity of
pollution and by suspected cause (from McGarrigle et al., 2010). ............. 9
Figure 2.2 Precipitation depth-duration-frequency for Valentia for the complete
(1940-1993) and partial time series (post-1975) (from Kiely, 1999). ........ 13
Figure 2.3 Soil erodibility (ton ha h ha-1 MJ-1mm-1) across Europe based on the
nomograph of Wischmeier and Smith (1978) (from Panagos et al.,
2012). .......................................................................................................... 23
Figure 2.4 Raindrop falling on a bare unprotected soil surface and detaching soil
particles from the original soil matrix: a) raindrop velocity, b) raindrop
impact producing a disrupting force in the form of laterally flowing jets
(Saavedra, 2005) ........................................................................................ 34
Figure 2.5 Schematic showing the distinction between rainfall detachment of the
original soil matrix, and re-detachment of previously eroded and
deposited sediment (from Rose, 1993)....................................................... 36
Figure 2.6 Processes in the transfer of P from terrestrial to aquatic ecosystems
(from Sharpley et al., 1995). ...................................................................... 47
Figure 2.7 Field-by-field soil P status (left) collated by index and showing below
optimum to the north and above optimum to the south. The SCIMAP
connectivity map (centre) on a 5 m pixel scale is also collated as a mean
score per field (right) and indicates CSA potential in the south where
the propensity for runoff (score closer to 1) is higher (adapted from
Wall et al. (2011)). ...................................................................................... 54
Figure 3.1 Arable land use in Ireland (CORINE, 2006) and selected/rejected soil
sampling sites. ............................................................................................ 68
Figure 3.2 Laboratory flume set-up for rainfall simulator/overland flow
experiment. ................................................................................................ 70
Figure 3.3 Rainfall Simulator (isometric drawing and photo of underside) ............. 71
Figure 3.4 Soil in laboratory flume before and after rainfall simulation. ................. 72
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Figure 3.5 Phosphorus and sediment losses over time from tillage soils inclined at
a 10 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody
(○), Fermoy ( ). ......................................................................................... 80
Figure 3.6 Phosphorus and sediment losses from selected tillage soils inclined at a
15 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody
(○), Fermoy ( ). ......................................................................................... 81
Figure 3.7 Phosphorus and sediment concentrations in runoff water from tillage
soils subjected to rainfall and overland flow (225 ml min -1) when
inclined at a 10 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦),
Bunclody (○), Fermoy ( ). ........................................................................ 85
Figure 3.8 Phosphorus and sediment concentrations in runoff water from tillage
soils subjected to rainfall and overland flow (450 ml min -1) when
inclined at a 10 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦),
Bunclody (○), Fermoy ( ). ........................................................................ 86
Figure 3.9 Mass loss of phosphorus and sediment in runoff from selected tillage
soils subjected to rainfall and two overland flow rates (225 and 450 ml
min-1) while inclined at a 10 degree slope. ................................................ 88
Figure 4.1 Runoff Dissolved Phosphorus Risk Indicator using Morgan’s P for
selected tillage soils at 10 and 15 degree slopes under a 30 mm hr-1
rainfall. Rainfall 1 - 1st rainfall event; Rainfall 2 - 1hr after Rainfall 1;
Rainfall 3 - 24 hr after Rainfall 2. ............................................................. 99
Figure 4.2 Runoff Dissolved Phosphorus Risk Indicator using water extractable P
for selected tillage soils at 10 and 15 degree slopes under a 30 mm hr-1
rainfall. Rainfall 1 - 1st rainfall event; Rainfall 2 - 1 hr after Rainfall 1;
Rainfall 3 - 24 hr after Rainfall 2. ........................................................... 101
Figure 4.3 Runoff Dissolved Phosphorus Risk Indicator using Mehlich-3
extractable P for selected tillage soils at 10 and 15 degree slopes under
a 30 mm hr-1 rainfall. Rainfall 1 - 1st rainfall event; Rainfall 2 - 1hr
after Rainfall 1; Rainfall 3 - 24 hr after Rainfall 2. ................................ 102
xx
Figure 4.4 Log FWMDRP against Morgan’s P for selected tillage soils at a 10
degree slope under a 30 mm hr-1 rainfall and subjected to two distinct
overland flow rates (225 and 450 ml min-1)............................................. 104
Figure 4.5 Runoff Dissolved Phosphorus Risk Indicator using Morgan’s P, water
extractable P, and Mehlich-3 P for selected tillage soils at a 10 degree
slope, under a 30 mm hr-1 rainfall and subjected to two distinct
overland flow rates (225 and 450 ml min-1)............................................. 108
Figure 4.6 Runoff Dissolved Phosphorus Risk Indicator using calcium chloride
extractable P and soil P saturation for selected tillage soils at a 10
degree slope, under a 30 mm hr-1 rainfall and subjected to two distinct
overland flow rates (225 and 450 ml min-1)............................................. 109
Figure 5.1 Equipment used to carry out visual soil assessments. ............................ 124
Figure 5.2 Methodology for determining soil erodibility using Soil Survey of
Ireland data.............................................................................................. 127
Figure 5.3 Visual scoring of soil structure during visual soil assessments at the 6
study sites. ................................................................................................ 132
Figure 5.4 Methodology for determining erosion risk (adapted from Defra
(2005))....................................................................................................... 144
Figure 5.5 Methodology for upgrading/downgrading Defra risk class using soil
quality indicators and average annual rainfall. ...................................... 145
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Chapter 1
Chapter 1 Introduction
1.1 Overview
In the Republic of Ireland, agriculture accounts for 60.8% (approx. 4.2 million ha) of the
total land area (CSO, 2010) - well above the European average of approximately 40%. Of
this agricultural land, approximately 80% is devoted to grass (silage, hay and pasture),
10% to rough grazing, and 10% to arable cereal and crop production, with barley as the
most important cereal crop representing 4.4% (185,900 ha). Crop production is
concentrated in the south of the country (Schulte et al., 2010a), where the soils are highly
suitable for tillage, having a light-to-medium texture and free drainage (Gardiner and
Radford, 1980). Arable farming has been identified as one of the principal factors
impacting on the ecological status of streams and rivers in Ireland (Donohue et al., 2006).
Arable land has higher phosphorus (P) application rates than grassland due to higher
offtakes and the need for new seeding each year, and can lead to a build-up of soil test P
(STP) over time. Despite having a significantly higher P export coefficient (Doody et al.,
2012) and greater susceptibility to erosion (Boardman and Poesen, 2006), arable land has
not received the same level of research as the more dominant grassland. Therefore, its
current degradation status and contribution to surface water impairment in Ireland is as
yet unknown.
A literature review highlighted knowledge gaps in relation to two key areas where tillage
has the potential to impact negatively on water and soil quality: (1) Nutrient loss, in
particular P, via surface runoff needs to be quantified. There is a need for risk indicators
that can identify critical source areas (CSAs) of P loss, so that current agronomic
guidelines on P application can be assessed with respect to their potential to impact on
water quality. Remediation measures can then be focused in these areas. (2) Sediment
loss from fields with reduced ground cover and impaired soil structure. There is currently
no methodology available in Ireland with which to quantify the risk of erosion. There is a
need for a screening method so that high risk sites can be identified and prioritised for
1
Chapter 1
further investigation. While the two areas required separate examination and are reported
on in separate chapters, they are intrinsically linked in their potential to impact on water
and soil quality. Combining both parameters helps to build a clearer picture of the overall
risk to the environment. The outcome will also be more useful to farmers and advisors,
who will ultimately utilise the knowledge generated.
Nutrient enrichment causing eutrophication is the principal and most widespread pressure
on the aquatic environment in Ireland. In freshwaters this enrichment is attributed
primarily to excessive inputs of P (DEHLG, 2009). Overland flow events resulting from
intense rainfalls could potentially transport P and sediment from vulnerable tillage soils
to surface water bodies. In addition, chronic P losses from the soil as a result of a buildup of STP in excess of crop requirements can contribute to losses. As a result of the
European Union (EU) Water Framework Directive (WFD) (2000/60/EC: Council of the
European Union, 2000), there is increasing pressure in Europe and Ireland to develop Pbased management practices that will reduce the risk of diffuse losses from agricultural
land to surface waters. Furthermore, the emergence of policies such as the proposed EU
Draft Soil Framework Directive (SFD) (COM (2006) 231), which deals with concern
over soil degradation and anthropogenic impacts to soil is likely to increase the
requirement for assessment of soil quality and identification of soils at risk from
degradation (Bone et al., 2012).
The overarching objective of the present body of work was to develop two methods
which can be used in parallel to screen tillage land for likelihood of erosion and P loss to
surface waters. The first method uses the relationship between STP and dissolved reactive
P (DRP) measured in surface runoff water to define a threshold STP level, above which
DRP concentrations in runoff from tillage soils subject to agronomic guidelines have the
potential to cause eutrophication. The second method uses a novel soil structure and
quality assessment system to provide farmers with the necessary tools to identify tillage
fields where the risk of erosion is high. This screening method will allow mitigation
strategies, such as buffer zones and minimum tillage, to be targeted on a smaller number
of tillage sites, leading to cost and resource efficiencies.
2
Chapter 1
1.2 Objectives
The specific objectives of this study were:
1. To review the current state of research and the regulatory regime relating to
diffuse P and suspended sediment (SS) loss for tillage soils in Ireland. Following
this, the study aimed to identify the key factors contributing to surface runoff and
erosion at catchment-scale.
2. To acquire experimental data on the loss of P and SS in surface runoff from a
selection of tillage soils when subjected to high intensity, low energy simulated
rainfall and inclined at various slopes.
3. To examine the effect of increasing overland flow rates on the mobilisation and
transport of DRP, particulate P (PP) and SS by flowing water.
4. To quantify the STP threshold above which there may be a potential threat to
surface water quality in Ireland.
5. To develop a novel screening toolkit for use by farmers/specialist advisors in
assessment of tillage fields for likelihood of erosion and reduced soil quality.
1.3 Procedure
A literature review of P and SS release from Irish tillage soils and the methods used to
quantify and reduce losses to surface waters was undertaken. Critical source areas of P
were chosen for further study as the identification of CSAs, where specific mitigation
measures can be targeted, has significant implications for the WFD river basin
management plans (RBMP). The identification of an environmental soil P threshold,
above which surface runoff from tillage soils will have a negative impact on water
quality, will ensure that mitigation strategies employed in Ireland to meet the
3
Chapter 1
requirements of the WFD are targeted in those areas where they will be most cost
effective.
Following this, a runoff experiment was designed to compare the P and SS releases from
5 Irish tillage soils when subjected to high intensity (30 mm hr -1), low energy rainfall. In
this experiment, each rainfall simulation comprised 3 successive 1-hr rainfall events
applied at intervals of 1 hr and 24 hr to determine the effect of storm interval on surface
runoff. As the velocity of surface runoff increases with slope, each soil was examined at 2
slopes, 10 and 15 degrees, to investigate the effect of slope on the magnitude of measured
losses in the runoff. Using the relationship between STP in study soils and the DRP
measured in the surface runoff, a runoff dissolved P risk indicator (RDPRI) was
developed to identify the STP threshold above which there is a potential threat to surface
water quality. This was achieved by constructing 95% confidence limits around the DRP
relationships using the upper and lower confidence bands for the linear predictor.
It is critical that the RDPRI be tested under increasing overland flow rates, because the
potential for pollutant transport to surface waters increases with flow rate. To determine
the effect of overland flow rate on P and SS released in surface runoff, each soil was
subjected to two distinct overland flow rates of 225 and 450 ml min-1 (possible worst case
scenarios in fields where the soil has become saturated due to high rainfall intensities),
which were introduced at the top of the flume in the presence of rainfall.
The majority of P lost in runoff during the rainfall and overland flow simulation
experiments was in particulate form and therefore was not accounted for in the RDPRI.
To address the issue of PP loss in surface runoff from the study soils, a field-based
erosion risk assessment approach was adopted with the aim of developing a screening
toolkit with which, farmers/specialist advisors can assess tillage fields for likelihood of
erosion and loss of soil quality. At each study site, different methods of erosion risk, and
soil structure and quality assessment were conducted. Comparisons of erosion risk
indicators and soil quality indicators with observed erosion levels in the study fields
allowed selection of a strategic set of indicators for inclusion in a screening toolkit, which
4
Chapter 1
can be used by Irish tillage farmers and specialist advisors to identify, expeditiously,
fields with high erosion risk and poor soil quality.
1.4 Structure of dissertation
In Chapter 2, a review of the environmental risks posed by poorly managed cultivated
soils due to weakened structure and high P status and the methods used to assess and
eliminate these risks are presented. Chapter 3 details the results of a laboratory rainfall
simulation study, which was used to determine the effects of land slope, overland flow
rate and storm interval on P and SS losses measured in surface runoff. Chapter 4 uses the
data compiled in Chapter 3 to develop a RDPRI for Irish tillage soils. Chapter 4 also
ranks soil extractable P methods with respect to their potential to be used as P loss risk
indicators, and identifies important parameters for which to test when attempting to
predict SS, TP, and PP loss from tillage soils. Chapter 5 details the results of erosion risk,
and soil structure and quality assessments carried out at each study site. Chapter 5 also
details the development and testing of an erosion and soil quality screening toolkit for
farmers and specialist advisors. Finally, in Chapter 6, conclusions from the study are
presented and recommendations for future research work are made.
To date, two peer-reviewed papers have been published from this work:
Regan, J.T., Fenton, O. and Healy, M.G. 2012. A review of phosphorus and sediment
release from Irish tillage soil, the methods used to quantify losses and the current state of
mitigation practice. Biology and Environment: Proceedings of the Royal Irish Academy
112B: 157 - 183.
Regan, J.T., Rodgers, M., Healy, M.G., Kirwan, L. and Fenton, O. 2010. Determining
phosphorus and sediment release rates from five Irish tillage soils. Journal of
Environmental Quality 39: 185-192.
5
Chapter 1
In addition, this work contributed to a review of erosion in arable soils in Atlantic
Europe, contained in the following publication:
Creamer, R.E., Brennan, F., Fenton, O., Healy, M.G., Lalor, S.T.J., Lanigan, G.J., Regan,
J.T., and Griffiths, B.S. 2010. Implications of the proposed Soil Framework Directive on
agricultural systems in Atlantic Europe – a review. Journal of Soil Use and Management
26: 198-211.
A number of international and national conference papers have also been published
describing this work. A list of outputs and manuscripts in preparation for submission to
international journals can be found in Appendix A.
6
Chapter 2
Chapter 2 Literature Review
Overview
The environmental risks posed by cultivated soils due to susceptibility to erosion and
high P status, and the methods used to quantify and eliminate these risks are reviewed in
this chapter. The contents of this chapter are published in Biology and Environment:
Proceedings of the Royal Irish Academy (112B: 157 - 183, 2012).
2.1 Introduction
In Ireland, agriculture is an important national industry accounting for 60.8% (approx 4.2
million ha) of the total land area (CSO, 2010). Of this agricultural land, approximately
0.4 million ha is devoted to arable cereal and crop production. Barley, wheat and oats are
the main cereal crops representing 72% of this area. The cereal sector is small in relative
terms representing < 1% of total EU production and approximately 3% of national gross
agricultural output, with an annual cereal output of approximately 2 million tonnes in
2010 (Eurostat, 2012). In the southeast, cereals alone account for 17% of farmed land in
County Carlow and 23% in County Wexford (Hooker et al., 2008). In the south of the
country and the southeast in particular, the favourable climate provides better
opportunities for seedbed preparation and harvesting. Here, there are fewer wet days,
higher temperatures, less chance of frost, higher radiation receipts and more hours of
bright sunshine (Collins and Cummins, 1996). Ireland‟s cereal yields are among the
highest in the world largely due to suitable soils and climate and the technical ability of
the farmers.
In Europe, de Wit et al. (2002) observed that agriculture contributed approximately onethird of all P to pollution in rivers. A biological survey of 13,177 km of Irish river and
stream channels from 2007 to 2009 (McGarrigle et al., 2010) estimated that 20.7% were
slightly polluted, 10% were moderately polluted, and 0.4% were seriously polluted.
7
Chapter 2
However, when assessed for ecological status, according to the requirements of the EU
WFD (2000/60/EC: Council of the European Union, 2000), based on the various
biological and supporting physico-chemical quality elements for individual river water
bodies on a one-out all-out basis, a different picture emerges, with just 52% of rivers
achieving „good status‟ (McGarrigle et al., 2010). Of the 2,515 sites surveyed in this
period, the percentage of pollution attributed to agriculture was approximately 54% and
39% in rivers and streams that were slightly or moderately polluted, respectively, but
only 15% in those that were seriously polluted. This data indicates that diffuse
agricultural pollution causing eutrophication, accounted for 47% of the number of
polluted river sites recorded over this period (Figure 2.1).
Algal growth, due to excessive P inputs, is often the primary reason for failing to achieve
good ecological quality (Jeppesen et al., 2003). The WFD aims to restore polluted water
bodies to „at least‟ good ecological and chemical status (≤ 0.035 mg MRP L-1 in rivers)
by 2015 and prevent any further deterioration in the status of surface waters, transitional
waters, groundwater and water-dependent terrestrial ecosystems and wetlands. Key to the
WFD is the adoption and implementation of RBMP and Programme of Measures (POM)
by 2012. These set out the actions required within each major river basin to achieve set
environmental quality objectives, which will be reviewed on a six-yearly basis. These
plans must include basic measures and, where necessary, supplementary measures to be
implemented for a specific water body to help achieve prescribed water quality standards.
The RBMP have identified agriculture as one of the main physico-chemical pressures
affecting a water body. The basic regulation for agriculture in the Republic of Ireland
(henceforth referred to as Ireland) is the Nitrates Directive (91/676/EEC: Council of the
European Union, 1991), and is given statutory effect in the European Communities
(Good Agricultural Practice for Protection of Waters) Regulations 2010 (SI 610 of 2010).
The latter sets out detailed nutrient management controls for farming, including P
application rates for crop production.
8
Chapter 2
Figure 2.1 Polluted river sites surveyed in 2007-2009 grouped by severity of
pollution and by suspected cause (from McGarrigle et al., 2010).
9
Chapter 2
Implicit in these directives and management plans is the protection and preservation of
soils. The EU Draft SFD identifies the following threats to soil quality: erosion, decline
of soil organic carbon (SOC), compaction, contamination, sealing, salinisation, landslides
and desertification. However, to date, the directive has not been ratified. If ratified,
member states will have to identify areas where soil degradation processes have occurred,
or are likely to occur in the future. The identification of areas at risk of erosion will take
account of the following parameters: soil type, texture and density, hydraulic properties,
topography (including slope gradient and length), land cover and use, climate (including
rainfall distribution and wind characteristics), hydrological conditions and agroecological zones. Once risk areas have been identified, member states will be required to
draw up a POM, including a timetable for implementation. Ratification of the SFD will
result in the unification of soil measures under one directive and provide a common
approach and level playing field for member states with regard to soil protection
(Creamer et al., 2010).
As part of the WFD, new environmental quality standards (EQS) for chemical
compounds, and biological and hydromorphological classification systems for surface
waters in Ireland have been developed by the Environmental Protection Agency (EPA)
(Bowman, 2009). The proposed EQS values for chemical parameters in rivers are derived
using a WFD-compliant methodology, using 3 years of biological and physicochemical
data collected in Irish rivers considered to be of high and good biological status. The EQS
values are necessary to define the high/good and good/moderate boundaries for the
purposes of ecological classification. Rivers with molybdate reactive P (MRP)
concentrations ≤ 0.025 and 0.035 mg P L-1 are classified as having high and good status,
respectively (Bowman, 2009).
In agriculture, diffuse SS and P loss from soils occur primarily as a result of overland
flow (Sharpley and Rekolainen, 1997; Sharpley et al., 2003a; Vadas et al., 2004)
following hydrological (storm) events, and are the primary non-point-sources of pollution
degrading the quality of surface waters (Daniel et al., 1998). However, P loss can also
occur through drainage, with the most significant instances of downward movement of P
10
Chapter 2
through the soil profile being associated with excessive application of P in manure and
fertiliser (Sims et al., 1998), in particular, on soils with low P-retention properties and/or
significant preferential flow pathways (e.g. cracking clay soils) (Hart et al., 2004). Recent
extensive water quality surveys in Ireland revealed that diffuse P pollution originating
from agricultural land, and transported by runoff and subsurface flows (e.g. leaching and
artificial drainage systems) is the primary cause of the deterioration of surface water
quality (Nasr et al., 2007).
2.2 Soil erosion
In Ireland, tillage land (cereals and root crops) only accounts for 9.6% of agricultural area
utilised (CSO, 2010), but it accounts for a higher amount of fertiliser use than pastoral
land (Lee, 1986). Therefore, the majority of research on the island of Ireland has focused
on quantifying nutrient and, to a lesser extent, sediment losses from permanent grassland
at laboratory- (Doody et al., 2006; Murphy, 2007; Murphy and Stevens, 2010), plot/field(Tunney et al., 2007; Kurz et al., 2000, 2005 and 2006; Douglas et al., 2007; Doody et al.,
2010 and 2011) and catchment-scales (Smith et al., 1995 and 2005; Scanlon et al., 2004;
Jordan et al., 2005 a,b; Jordan et al., 2007). Modelling of diffuse P loss from grassland
catchments has also been undertaken by Jordan et al. (2000), Daly et al. (2002), Scanlon
et al. (2005), and Nasr et al. (2007) with the aim of improving management strategies to
minimise P loss. Arable land is, however, more susceptible to water erosion than
grassland (Van Oost et al., 2009) due to greater soil surface exposure to erosive forces
during fallow and planting periods (Lal, 2001) and soil disturbance by tillage operations
(Lal, 2001), which alters its structure. Furthermore, in grassland soils, higher infiltration
rates can lower runoff rates (Fullen, 1991) and higher soil organic matter (SOM) levels
can reduce erodibility (Fullen, 1998).
Increasing eutrophication of many surface water bodies in arable regions of the UK has
been linked with increasing rates of soil erosion causing sediment and P loss from fields
cropped with winter cereals and with an accumulation of soil P through continuous
application of fertiliser and manures (Catt et al., 1998). Research to establish the
11
Chapter 2
circumstances leading to sediment and P losses from arable land and to quantify these
losses has been carried out in the UK (Speirs and Frost, 1987; Chambers et al., 1992; Catt
et al., 1998; Chambers and Garwood, 2000) and throughout Europe (Kronvang et al.,
1997; Verstraeten and Poesen, 2001; Miller et al., 2009) at multiple scales. Reported
sediment and P losses from arable sites in these and other similar studies were
significantly higher than losses from grassland, and were high enough to cause concern
over eutrophication of surface water bodies in arable areas. Given the susceptibility of
tillage land to erosion in general and high P applications associated with this land use in
Ireland, there is a need to quantify the sediment and P losses from tillage soils to surface
water bodies and monitor the effects of improving tillage practices.
Climate change will continue to make agricultural land more susceptible to P and
sediment loss, with more extreme weather conditions forecast to increase the amounts of
surface runoff, the detachment and transport of soil particles, and P released from
fertiliser and manure amendments (Mainstone et al., 2008). Summer storm events,
coupled with impervious agricultural soils, can lead to P addition to waterways during the
growing season. By analysing long-term, hourly precipitation data from eight sites and
daily streamflow data from four rivers in Ireland, Kiely (1999) identified a climatic and
hydrologic change in 1975, after which both precipitation and streamflow increased. The
study found that, since 1975, there has been an increase of about 10% in precipitation on
the west-coast. Furthermore, analysis of the twenty most extreme 24-hr-duration rainfall
events at the Valentia observatory for the period 1940-1993 identified 75% of them as
having occurred since 1975, even though the period length since 1975 was 18 years by
comparison with a length of 36 years before 1975. Kiely (1999) also noted that increases
in extremes (frequency and/or magnitude) are likely to lead to increased incidence of
flooding, which may also lead to more frequent incidences of deterioration in river water
quality due to more frequent diffuse surface runoff. The immediate concern for river
basin managers is highlighted by the precipitation depth-duration-return period curves for
Valentia Observatory (Figure 2.2), which show that a storm depth that only occurred once
every 30 years prior to 1975, is likely to occur approximately once every 10 years after
1975. The trend since 1975 of increasing precipitation and more extreme events is set to
12
Chapter 2
continue. In the UK, in recent years, it has been observed that heavy rainfall events
during the summer months are increasing (Maran et al., 2008).
Figure 2.2 Precipitation depth-duration-frequency for Valentia for the complete
(1940-1993) and partial time series (post-1975) (from Kiely, 1999).
2.2.1 Factors affecting soil erosion rates in Ireland
There is greater pressure on soils in Ireland due to increased erosion under modern
intensive arable production systems and disposal of organic wastes to soils. Very few
studies in Ireland have looked at erosion rates on tillage soils despite international
research by Kronvang et al. (1997), Chambers et al. (2000), Deasy et al. (2009), Stevens
et al. (2009), and Van Oost et al. (2009) concentrating specifically on such soils due to
their erosion propensity. Greater demand for food and advances in farm machinery has
resulted in intensified crop production, leading to higher tillage and water erosion rates in
Europe. Lindstrom et al. (2001) defined tillage erosion as „the net movement of soil
13
Chapter 2
downslope through the action of mechanical implements‟. Both types of erosion can have
a negative impact on productivity, with the most severe impact occurring due to a loss of
topsoil depth in soils with a root restrictive layer (Lal, 2001). It is estimated that 115
million ha, or 12% of Europe‟s total land area, is affected by water erosion (EEA, 1995).
Soil water erosion in the UK is primarily a regional phenomenon associated with sandy
tillage soils in the southwest and southeast of England (Chambers et al., 2000). In Ireland,
soil erosion is primarily a phenomenon associated with tillage soils and periods of intense
rainfall (Fay et al., 2002). The potato growing area of Donegal is a good example. The
main drivers predisposing arable soils to water erosion in Ireland and the UK are: soil
type, precipitation (amount, duration and intensity), and management practice.
2.2.1.1 Soil type
Soil type is important when determining the erosion risk from an arable field. The texture
of a soil strongly influences SOM storage (Fullen et al., 2006). Soil organic matter breaks
down faster in sandy soils than in fine-textured soils due to: (1) a lack of clay for physicochemical binding with SOM (Fullen et al., 2006) and (2) greater oxygen availability for
decomposition by microorganisms in the former. Disturbance of topsoil by tillage
operations further aerates the soil which, in turn, increases soil SOM decomposition.
Fullen et al. (2006) also reported that silt can play an important role in influencing
organic matter (OM) storage in clay-deficient sandy soils. Sandy soils are particularly
vulnerable to erosion due to low SOM content and poor structural stability, which
predisposes the soil to: disaggregation under raindrop impact and a subsequent
development of a surface crust, reduction of infiltration rate, and surface runoff (Quinton
and Catt, 2004). Long-term arable use and modern cultivation methods can result in light
textured soils capping (surface crust caused by heavy raindrops on finely cultivated soils)
under rain impact (Fraser et al., 1999).
Increasing SOM content improves the cohesiveness of the soil, reduces the risk of surface
crusting, lowers the risk of soil compaction, increases its water holding capacity and
promotes soil aggregate formation, thereby improving structural stability and reducing
14
Chapter 2
erosion. As an EU Member state, Ireland is required to monitor SOC levels in long-term
(6 years or more) continuous tillage soils in order to ensure that sustainable management
practices are put in place to reduce any further decline in SOC (Department of
Agriculture, Fisheries and Food (DAFF) 2009a). In contrast to the UK, no work to date
has been carried out in Ireland to determine the susceptibility of sandy soils to erosion
under arable cropping. Findings in the UK may be indicative of potential erosion
problems with sandy tillage soils in Ireland, but there is a need for Irish-specific data to
establish if there is an erosion problem.
2.2.1.2 Precipitation (amount, duration and intensity)
Rainfall characteristics influence processes affecting infiltration, runoff, soil detachment,
and sediment and chemical transport (Truman, 2007). The risk of nutrient loss is
generally greatest when soils are near field-capacity or saturation, and any further
precipitation leads to water surpluses and either sub-surface drainage or overland flow
(Schulte et al., 2006a). In unsaturated soils, erosion and P loss is mainly governed by the
occurrence, frequency and timing of intense storm events that result in intense overland
flow events. Rainfall intensity is generally considered to be one of the most important
factors influencing soil erosion by overland flow and rills because it affects the
detachment of soil particles by raindrop impact and enhances their transport by runoff.
A study by Chambers et al. (2000) of 13 erosion-susceptible arable catchments
(containing medium silt, sand/light loam, silty clay loam, loamy sand and sandy loam
soils) in England and Wales between 1989 and 1994 revealed that soil erosion can occur
at any time of the year, provided the conditions suitable for erosion are present. These
include: lack of ground cover vegetation (< 15%); loose, fluffy and very fine seed bed
conditions; heavy rainfall (> 15 mm day-1) with a high intensity (> 4 mm hr -1) in the
presence of high winds; steep slopes; presence of valley floor features that concentrate
surface runoff; and compacted tramlines (unseeded wheeling areas used to facilitate
spraying operations in cereal crops). The incidence of severe erosion resulting in
transport of SS and P, in particular, tends to be highly dependent on hydrological storm
15
Chapter 2
events (Edwards and Withers, 2008), and it has been shown that approximately 90 to
95% of soil erosion occurs during the most severe 2% of storms (Winegardner, 1996).
Erosion also occurs over periods of prolonged lower-intensity rainfall (Robinson, 1999;
Fraser et al., 1999).
Mean annual precipitation for Ireland ranges from 750-1000 mm on the greater part of
the east coast to between 1000-1250 mm on the greater part of the west coast. The highest
annual rainfall of between 1600-2800 mm occurs when Atlantic rain-bearing storms
encounter landfall and mountainous terrain on the west coast. Any change in precipitation
(amount, duration and intensity) over Ireland as a result of climate change, is likely to
impact directly on P and sediment losses in surface runoff from agricultural soils. The
10-yr moving average for Ireland shows that rainfall amounts increased from 800 mm in
the 1890s to 1100 mm in the 1990s (McElwain and Sweeney, 2006).
The general consensus from numerous climate change studies in Ireland is that winter
rainfall will increase as will the frequency of intense rainfall events during summer. An
EPA modelling report by Sweeney et al. (2008) on the impacts of climate change for
Ireland projected an increase of 10% in winter rainfall by 2050, while reductions in
summer of 12-17% are projected by the same time. Spatially, the largest percentage
winter increases are forecast for the midlands, while summer reductions of 20-28% are
forecast for the southern and eastern coasts. Sweeney et al. (2008) also predicted more
frequent intense rainfall events during the summer. These projected changes will
necessitate adaptation in water resources management in Ireland. For example, increased
levels of erosion and greater suspended sediment loads will have to be managed for all
Irish rivers (Sweeney et al., 2008). Increased P export in summer, resulting from high
intensity rainfall events has been reported in numerous Irish grassland studies by Lennox
et al. (1997), Tunney et al. (2000), Kurz (2000), Morgan et al. (2000), Kiely (2000), and
Irvine et al. (2001). Overland flow events resulting from intense summer rainfalls could
potentially transport P and sediment from vulnerable tillage soils to surface water bodies
during the growing season.
16
Chapter 2
2.2.1.3 Management practices
Erosion driven by management practices such as the decisions made by farmers as to
what crops to grow, how to manage and prepare the land, and when to sow are also very
important and easier to change in the short-term. Best management practices on steep,
vulnerable slopes aims to minimise soil erosion losses, which, in turn, limit nutrient
losses to a water body (Creamer et al., 2010). Research on cultivation practice in the UK
by Chambers and Garwood (2000) identified valley features, lack of crop cover,
wheelings (the passage over soil by wheels of a vehicle) and tramlines as the main
contributors to erosion. Research by Silgram et al. (2010) on sandy loam and silty clay
loam soils on 4-degree slopes in England has shown that tramlines can represent the most
important pathway for P and sediment loss from moderately sloping fields. In a study of
mitigation options for sediment and P loss from winter-sown arable crops on three soil
types (sandy, silty and clay), Deasy et al. (2009) showed that compared to losses from
cropped areas without tramlines, losses of sediment and P were between 2 and 230 and 2
and 293 times greater from tramline areas, respectively. The increase in losses due to
tramlines was lower for the clay soils and greater for the silty soils, largely due to the
cohesiveness of the clay soil. However, it is important to note that tramline areas
normally only account for about 5% of the field area. Accelerated rates of soil erosion
within agricultural landscapes are causing major modifications to terrestrial carbon (C),
nitrogen (N) and P cycling (Quinton et al., 2010). Measures that can help maintain or
increase SOC include: adoption of reduced tillage; straw incorporation; use of organic
manures; use of cover crops; and adoption of mixed rotations (Hackett et al., 2010).
Increases in SOC resulting from management changes are slow and reversible (Hackett et
al., 2010).
Other contributors to erosion under modern intensive arable production systems are: ditch
removal and field enlargement; use of high-powered modern traction systems, which can
plough up and down slopes rather than contour ploughing (Quinton and Catt, 2004); use
of heavy rollers after sowing (Boardman, 1990); and loose, fluffy and very fine seed bed
conditions (Speirs and Frost, 1987; Catt et al., 1998). The removal of hedgerows, ditches
17
Chapter 2
and open drains is now prohibited as part of EU Cross Compliance. Key tillage
operations/practices that may impact on soil and water quality and possible mitigation
options for Ireland are outlined in Section 2.2.4, Table 2.1.
2.2.2 Field indicators of a soil’s susceptibility to runoff and erosion
2.2.2.1 Soil texture
Soil texture and aggregate stability are the two main factors which make a soil more or
less erodible (Coles and Moore, 1998). Soil texture also provides an indication of a soil‟s
susceptibility to structural decline and compaction. While field textures, determined by
hand, can approximate particle size distribution determined in the laboratory, they are
influenced by cementing materials such as SOM, clay mineralogy, CaCO 3 content and by
iron and aluminium oxides and hydroxides, and as such must be carefully interpreted, and
results treated with due caution. In general, field texture measurements are a good guide
to soil behaviour (Moore, 1998) because texture controls water movement, affects
chemical reactivity and nutrient availability, and is a factor in the erosion potential of a
soil. As such, they can be used to infer a soil‟s susceptibility to erosion, nutrient leaching,
waterlogging, subsurface compaction, and structural decline. While little can be done to
change or improve the texture of a soil, it is an important property for guidance on how
best to manage the soil.
2.2.2.2 Soil colour
Soil colour is mainly due to the presence of SOM (dark), iron oxides (yellow, brown,
orange and red) and manganese oxides (black), or it may be due to the colour of the
parent rock (for example, old red sandstone -red). A Munsell Soil Colour Chart (Munsell,
2009) is used in soil classification to describe the colour of each soil horizon using the
notations for hue, value and chroma. Hue is the colour frequency which, for most soils,
ranges from red to yellow; value refers to the lightness from white to black; and chroma
defines the degree of colour saturation or intensity of hue (Fitzpatrick et al., 1999). Soil
18
Chapter 2
colour is a good indicator of soil quality because it can provide an indirect measure of
other more useful properties that are not so easily and accurately assessed (Shepherd,
2009). Change in soil colour can be a useful indicator of (1) changes in OM under a
particular land use or management; (2) soil drainage class; (3) the amount and oxidation
state of soil iron oxides; and (4) degree of soil aeration, all of which have management
implications. By informing on soil drainage class, soil colour can be indicative of a
particular soil‟s associated runoff risk.
2.2.2.3 Mottling
Mottles are spots or blotches of different colours, or shades of colour interspersed with
the dominant colour of the soil (FAO, 2006). Their presence in a soil indicates that it is
less permeable than a whole coloured soil (Northcote, 1983), and therefore poses a
greater risk of erosion arising from infiltration excess overland flow. The number and
colour of soil mottles present in a soil horizon are important measurements related to soil
aeration and hence waterlogging. They are also early warning signs of a decline in soil
structure due to compaction under wheel traffic and over-cultivation (Shepherd, 2009).
Mottling is caused by changes in the distribution, concentration, and state of oxidation of
iron and manganese compounds, which are present in many minerals in the soil (Batey,
2000). The changes are caused by reducing conditions, which, in turn, result from
microbial activity during anoxic conditions (Batey, 2000). Under anaerobic conditions,
any iron (Fe) and manganese (Mn) present in their brown/orange oxidised ferric (Fe 3+)
and manganic (Mn3+) forms are reduced to grey ferrous (Fe2+) and manganous (Mn2+)
oxides. If air re-enters the reduced soil, the process is reversed. As oxygen depletion
increases, orange, and ultimately grey, mottles predominate. The abundance of grey
mottles indicates the soil is poorly drained for a large part of the year (Shepherd, 2009),
and is therefore prone to flooding during wet periods. These soils require careful
management to ensure that the soil is not left vulnerable to erosion by surface runoff.
19
Chapter 2
2.2.2.4 Soil structure and consistence
Soil structure refers to the arrangement of soil particles into discrete soil units (aggregates
or peds), which are separated from each other by pores or voids. Consistence refers to the
degree of cohesion or adhesion of the soil mass. A field description of soil structure
enables its classification in terms of grade (degree of distinctness and durability), class
(size) and type (shape and arrangement) of aggregates. These are important soil structural
characteristics that have major influence on the pathway of water movement (drainage or
runoff), root penetration and development, soil trafficability, and resistance of soil to
structural degradation. Soil structure is vulnerable to change by compaction and erosion,
and its preservation is key to sustaining soil function (Mueller et al., 2010). Soil structure
influences soil water movement and retention, erosion, crusting, nutrient recycling, root
penetration and crop yield (Bronick and Lal, 2005). Soils with good structure have
friable, fine, porous, sub-angular and sub-rounded (nutty) aggregates, while those with
poor structure have large, dense, very firm, angular or sub-angular blocky clods that pack
closely together (Shepherd, 2009). The decline in soil structure is increasingly seen as a
form of soil degradation and is often related to land use and soil/crop management
practices (Bronick and Lal, 2005). The influence of soil structure on soil erodibility is
primarily a result of reduced infiltration of rainwater in poorly structured soils, leading to
erosion caused by runoff from higher ground. Soil management practices that can
negatively affect soil structure include: (1) use of power harrows and heavy machinery to
produce a fine tilth in weakly structured fine sands and light silts; and (2) high axle loads
and incorrect tyre pressures resulting in soil compaction (Creamer et al., 2010).
Externalities such as runoff, surface- and ground-water pollution, and CO2 emissions are
influenced by soil structure (Bronick and Lal, 2005).
2.2.2.5 Soil porosity and root development
Porosity provides an indication of the amount of soil pores, which can exist in a wide
range of shapes and sizes between and within soil aggregates, and are occupied by air and
water. The type, size, abundance, continuity and orientation of pores can be determined
20
Chapter 2
using a hand lens as part of a detailed soil classification. Roots and macropores orientated
vertically are most important for water infiltration. The permeability of soil during
infiltration is mainly controlled by coarse pores, in which the water is not held under the
influence of capillary forces (Beven and Germann, 1982), whereas fine pores control
plant-available water and water storage capacity (Lal et al., 1999). Any reduction in the
size of the coarse pores as a result of hydro-mechanical processes like slaking of
aggregates and dispersion of clays or gross mechanical stresses like raindrop impact,
compaction and tillage, may reduce infiltration, thereby increasing the risk of surface
runoff. A reduction in voids or structural porosity will also affect root growth (Moore,
1998), and the absence of roots below the plough layer can be indicative of a plough pan.
For agricultural purposes, a soil should ideally be able to withstand stress induced by
vehicle traffic and to remain in a porous structural state, permitting water and air
transmission for crop growth (Zhang and Hartge, 1995). Soil structure degradation
following intensive agricultural activities, soil compaction, loss of structural stability and
the formation of surface crusts gives rise to the loss of continuity of elongated
transmission pores, thereby reducing water transport and increasing runoff and soil
erosion (Pagliai et al., 2004).
Soil porosity is influenced largely by type of structure; but it is also influenced by rooting
and by the activity of earthworms and other soil macro-organisms (Gardiner and Ryan,
1964). Plant root systems help to bind the soil by releasing a variety of compounds,
which have a cementing effect on soil particles (Bronnick and Lal, 2005), and can play an
important role in preventing soil erosion after the crop has been harvested. Soils with
good structure have a high porosity between and within aggregates, but soils with poor
structure may not have macropores or coarse micropores within the large clods, thus
restricting their drainage and aeration (Shepherd, 2009). The act of cultivating the soil
causes short-term increases in porosity but long-term decreases in aggregation (Bronnick
and Lal, 2005).
21
Chapter 2
2.2.3 Main soil properties controlling soil erosion at field-scale
Soils identified as potentially susceptible to runoff and erosion by visual and tactile
assessment of the field indicators just discussed can be further assessed by analytical
methods. Results obtained by these methods must be relayed to the farmer, so that he/she
is aware of the soil‟s vulnerability to specific threats and can then manage the soil
accordingly.
2.2.3.1 Particle size distribution
Particle size analysis has an advantage over texture determinations carried out by hand in
the field, in that it enables the classification of sandy loams as either very fine, fine,
medium, or coarse based on the exact particle size range within the sand fraction, which
is difficult to determine accurately by hand estimation. The fineness of the sand particles
within a sandy loam is important when assessing its susceptibility to erosion. Resistance
to erosion is lowest for small non-cohesive grains, particularly silt and fine sand-sized
particles with low clay content (Grimm et al., 2002). The most erodible soils identified by
Romero et al. (2007) when measuring interrill and rill erodibility were those with the
greatest amount of silt and very fine sand, while the most resistant to erosion were clayey
soils. Fine sandy loam and silty loam textures can be highly susceptible to water erosion
and may also be hardsetting (Moody and Cong, 2008).
Particle size distribution is one of the major soil properties governing soil erodibility and
is one of the crucial factors required to assess soil erodibility in terms of the K-factor of
Wischmeier and Smith (1978). The K-factor is a lumped parameter that represents an
integrated average annual value of the soil profile reaction to the processes of soil
detachment and transport by raindrop impact and surface flow (Renard et al., 1997). Most
recently, Panagos et al. (2012) produced a K-factor erodibility map of Europe using data
from 22,000 soil samples, collected during the first conducted LUCAS (Land Use and
Cover Area Frame Survey; 2009) pan-European soil sampling campaign. Inverse distance
weighted interpolation with a power parameter of 2 performed best (R2 adjusted = 0.81)
22
Chapter 2
to interpolate LUCAS point data to a soil erodibility map of Europe (Figure 2.3). This
represents an enormous improvement in the precise estimation of K-factor at European
level compared with past methodologies, which derived this attribute based only on 5
textural classes and relatively coarser scale (Panagos et al, 2012).
Figure 2.3 Soil erodibility (ton ha h ha-1 MJ-1mm-1) across Europe based on the
nomograph of Wischmeier and Smith (1978) (from Panagos et al., 2012).
23
Chapter 2
During the survey, 4,165 geo-referenced points from a standard 2 km x 2 km grid, were
visited in Ireland. Soil samples (0-30 cm) were collected at 389 of these survey points for
determination of sand, silt, clay and OM content, amongst other parameters. The mean Kfactor determined for the land area of Ireland was 0.039, just below the European average
of 0.041. Ireland had the lowest dispersion of K-factor values of all the countries
surveyed, as measured by the coefficient of variation. Based on the K-factor distribution
within Ireland (Figure 2.3), more than half the soils of Ireland can be considered to be at
least moderately susceptible (K-factors between about 0.032 and 0.052) to detachment
and transport by raindrop impact and surface flow. Soils having the highest K-factor
erodibility values are located mainly in the potato growing areas of Donegal and in the
tillage areas of the South of the country. These soils, if occurring on slopes > 3 and left
bare or degraded by compaction, may pose a significant erosion risk.
2.2.3.2 Soil organic carbon/matter
Soil organic carbon also serves to bind individual soil particles into larger aggregates and
is important in maintaining the aggregate stability. By keeping the aggregates in soil
stable (by combining with soil minerals), it promotes infiltration, the movement and
retention of water, helps develop and stabilise soil structure, cushions the impact of wheel
traffic and cultivators, and reduces the potential for wind and water erosion (Shepherd,
2009). Generally, OM can hold up to 20 times its weight in water and can, therefore,
directly affect soil water retention, which makes soil more resistant to drought and
erosion, as well as indirectly through its positive effects on soil structure (Dick and
Gregorich, 2004). The texture of a soil strongly influences SOM storage (Fullen et al.,
2006), with coarse textured soils being particularly vulnerable to SOM decline.
Furthermore, the decline of SOC levels has been highlighted in numerous legislative
reports and scientific papers as contributing to a decline in soil quality and can result in
increased soil erosion, loss of nutrients and an increased susceptibility to compaction
(Van-Camp et al., 2004).
24
Chapter 2
In a review of critical levels of SOC in tillage land in Ireland, Spink et al. (2010)
concluded that soil function is unlikely to be adversely affected when SOC is above a
threshold of 2% (equivalent to c. 3.4% SOM). Soils having less SOM than this threshold
should be further assessed to see if they are in good environmental and agronomic
condition. These further measures could include observation of: erosion, gullies in the
field, compaction and capping (Spink et al., 2010). Whitmore et al. (2004) developed a
unified framework of measurement to assess rapidly the stability of soil surface structures
in England and Wales. Classification of 120 soils using this protocol, found that all
aggregates classified as unstable or very unstable were from arable, light and medium
soils (< 35% clay) that had low SOC (< 1.5%). In addition, practically all grassland soils,
regardless of their SOC or clay content, were categorised as stable or very stable. Soil
organic matter is an important component of both managed and unmanaged terrestrial
ecosystems, but is especially important in influencing soil erodibility (Jankauskas et al.,
2007). As an EU Member state, Ireland is required to monitor SOC levels in long-term
tillage soils in order to ensure that sustainable management practices are put in place to
reduce any further decline in SOC (DAFF 2009). Measurement of soil organic carbon in
1310 soil samples in Ireland showed that it varies from 1.4% to 55.8%, with a median
value of 7% (Zhang et al., 2008). Most recently, Zhang et al. (2011) produced a spatial
distribution map of SOC in Ireland created using geographically-weighted regression.
2.2.3.3 Bulk density
The natural rate of soil erosion is accelerated by increased soil bulk density (BD)
resulting from vehicular traffic-induced compaction (Lal, 2001). A healthy soil bulk
density is important in terms of sustainable soil productivity and environmental wellbeing (Merrington et al., 2006). High BD values indicate a poorer environment for root
growth, reduced aeration and undesirable changes in hydrologic function, such as reduced
water infiltration (FAO, 2006). Since erosion will not occur without surface runoff, soil
infiltration rate is important in relation to erosion. The higher the BD in surface or subsurface layers, the lower will be the total porosity and hence the greater the risk of surface
runoff. Packing density (PD) - a derivative of BD calculated as PD = BD + 0.009(% clay)
25
Chapter 2
(Hawes et al., 2010) - is a better parameter than BD for comparison of physical structure
between different soils and, based on derived statistics, affords an excellent indirect
estimation of soil porosity (Hall et al., 1977).
Surface compaction along tramlines and wheelings can trigger erosion problems in winter
cereal crops (Deasy et al., 2009; Batey, 2009). In well-graded loam and sandy loam soils,
the susceptibility to compaction can be very high because smaller particles can fit into
spaces between larger particles, thus providing the ideal proportions of particles of
different sizes to achieve the densest packing arrangements. In clayey soils, BD normally
ranges from 1.2 to 1.5 ton m-3 and in sandy soils from 1.6 to 1.9 ton m-3 (Needham et al.,
1998). Irrespective of soil textural group, arable soils consistently have a higher BD than
other land uses. Merrington et al. (2006) proposed critical BD thresholds for broad soil
and habitat groupings in England and Wales derived from a range of data sources. The
proposed BD threshold for arable and horticultural function on mineral soil was 1.3 ton
m-3. They also noted that for UK conditions one would expect severe hydrological
degradation of a soil with bulk density > 1.5 ton m-3. Packing density classes according to
Van Ranst et al. (1995) are: low < 1.40 ton m-3, medium 1.4 - 1.75 ton m-3 and high >
1.75 ton m-3.
2.2.3.4 Aggregate stability
Aggregate stability is one of the most complex and dynamic soil properties affecting
principal physical and hydraulic soil characteristics, such as infiltration rate, hydraulic
conductivity and erodibility (Dimoyiannis, 2011). Crop management practices such as
application of organic fertilisers, liming, incorporation of stubble and min-till or no-till
cultivation, can improve aggregate stability. Aggregation is essentially the flocculation
and cementation of individual soil particles to form aggregates. The primary soil
properties influencing aggregation and aggregate stability are texture, clay mineralogy,
SOM, cations, sesquioxides and calcium carbonate (Le Bissonnais, 1996). Clay
flocculation promotes soil aggregation and structural stability. The general consensus is
that the clay fraction has a positive effect on structural stability compared to the silt and
26
Chapter 2
sand fractions. Cations such as Fe3+, Al3+, Mg2+ and Ca2+ serve to: (1) stimulate the
precipitation of compounds that act as bonding agents for primary soil particles and (2)
form bridges between clay and SOM particles resulting in aggregation (Bronick and Lal,
2005). Dontsova et al. (2001) showed that Ca2+ ions were more effective than Mg2+ in
aggregating soil clays and that if soils are prone to surface sealing, it is beneficial to
manage them to high Ca:Mg ratios. High concentrations of Al and Fe oxides and
hydroxides, often referred to as cementing agents, have the effect of increasing aggregate
stability (Needham et al., 1998). The presence of Al and Fe oxide in soils also has a
favourable effect on soil structure. Evidence of soil structural improvement is provided
by increased aggregate stability, permeability, friability, porosity, and hydraulic
conductivity (Goldberg, 1989). In general, Al oxides have a greater stabilising effect than
spherical Fe oxides on structure because of their platy morphology (Goldberg, 1989).
Dimoyiannis (2011) found that the aggregate stability of calcareous surface soils in
Greece was positively affected by the presence Al oxides, whereas Fe oxides had no
significant effect. The same study found that the presence of free carbonates in soils as an
excess negatively affected aggregate stability.
Water is the main cause of aggregate breakdown in most soils, either directly by rainfall
or by surface runoff. Water-stable aggregation on the soil surface is therefore important
when considering the inherent erodibility of a soil and its susceptibility to soil structural
decline. There are four main mechanisms for aggregate breakdown: (1) slaking; (2)
differential swelling; (3) mechanical; and (4) physicochemical dispersion (Le Bissonnais,
1996). Slaking is the spontaneous disintegration of a soil aggregates which have
insufficient strength to withstand the stresses induced by rapid water intake. Differential
swelling depends on the same properties as slaking and produces similar
microaggregates, but normally occurs in soils with higher clay contents. Mechanical
breakdown is a result of the kinetic energy imparted to the soil by rain drop impact.
Physicochemical dispersion is the complete breakdown of aggregates into primary
particles because the attractive forces holding the particles together are lessened by
wetting. The Emerson aggregate test (Emerson, 1967) enables visual assessment of how
aggregates breakdown in water (slaking or dispersion), and is useful for broadly defining
27
Chapter 2
the stability of soils. Recommendations about how a soil should be managed are made
based on the Emerson score.
2.2.4 Impact of tillage farming
Concern with the effects on water quality and river health of non-point-source delivery of
nutrients and pollutants has given added impetus to research on the dynamics of erosion
and erosion-driven processes, and, in particular, the size distribution of transported
sediment (Hogarth et al., 2004). Soil erosion is a gravity-driven process by which soil
particles are first detached from the soil mass by rainfall and runoff, and then transported
by rainfall and runoff until they settle out of suspension. Concerns about present-day soil
erosion are related to accelerated erosion, where the natural rate has increased as a result
of human activities, which leave the land unprotected and vulnerable. These activities
predominate on agricultural soils and, in particular, on tillage soils where intensification
of land management, change in land cover through cultivation, and inappropriate land use
(cultivation of steep slopes), coupled with naturally occurring erosive rainfalls, can lead
to accelerated erosion rates. Excessive erosion can have a negative impact on-site
(deterioration in soil structure, decrease in crop yield, ecosystem damage and loss of soil,
OM and nutrients) and off-site (siltation and pollution of receiving waters and carbon
dioxide (CO2) release from broken-down clay particles and SOM). Lal (1995) estimated
that global water erosion and related processes release 1.14 Pg C/yr (1 Pg=1015 g) to the
atmosphere. Furthermore, extensive erosion can reduce the ability of soils to withdraw or
sequester C as CO2 from the earth‟s atmosphere. However, results from other studies
indicate that erosion enhances CO2 uptake (Smith et al., 2001) and that erosion and
deposition reduce CO2 emissions from the soil into the atmosphere by exposing low Cbearing soil at eroding sites and by burying SOC at depositional sites (Liu et al., 2003).
For example, tillage erosion and deposition leads to the burial of c. 7 Tg C y–1 in the area
covered by the CO-oRdination of INformation on the Environment (CORINE) (CORINE,
2000) database (assuming a 2% topsoil C content) (Van Oost et al., 2009).
28
Chapter 2
Soil loss on arable agricultural land is typically an order of magnitude higher than under
undisturbed native vegetation (Van Oost et al., 2009), and two orders of magnitude
higher than rates of soil formation (Montgomery, 2007). There is much evidence to show
that soil erosion due to rainfall and overland flow is exacerbated by tillage operations.
However, of similar importance is the extent of tillage erosion resulting directly from
tillage operations. This generally results in a movement of soil from convex shaped to
concave shaped landscapes, and leads to a nutrient-rich soil in the latter. While water
erosion is strongly controlled by soil characteristics such as soil stone level, texture and
crusting potential (Van Oost et al., 2009), experimental studies have shown that tillage
speed, depth, direction and implement characteristics are the primary controlling factors
on tillage erosion (Van Oost et al., 2006). It is of major importance that eroded nutrients
and sediment are retained in-field so as not to impact on surface water quality.
Given that rates of soil redistribution in the medium-term are influenced by tillage
displacement as well as water erosion, it is necessary to separate these two components of
soil redistribution in order to obtain a reliable assessment of water erosion rates (Blake et
al., 1999). By using a tillage erosion diffusion-type model based on the one Lobb et al.
(1999) proposed and land use databases, Van Oost et al. (2009) estimated that the mean
gross tillage erosion rates for the part of Europe covered by the CORINE land use
database was 3.3 ton ha-1 yr-1. For the same land area, they estimated the average water
erosion rate was 3.9 ton ha -1 yr-1 by using water erosion estimates for arable land,
orchards and vineyards compiled in a study by Cerdan et al. (2006) of datasets from 81
experimental sites across 19 European countries. The model also used large-scale land
use (CORINE), soil (Soil Geographical Database of Europe), topography (Shuttle Radar
Topography Mission) (Ciat, 2004) and soil erodibility datasets for Europe. For the
cropland area of Ireland, the same models estimated the average tillage and water erosion
rates to be 2.9 and 4.4 ton ha -1 yr-1, respectively. These erosion rates are higher than
average rates of soil formation (consisting of mineral weathering, soil biomass growth
and dust deposition) which range from 0.3-1.4 ton ha-1 yr-1, with the lower limit being
indicative of European conditions (Creamer et al., 2010). Research on tillage soils in
Ireland is needed to validate erosion rates given in the model of Van Oost et al. (2009).
29
Chapter 2
Table 2.1 Key tillage operations/practices that may impact on soil and water quality and possible mitigation options
Operation/practice
Impact on soil
Potential impact on water quality
Possible solutions
Cultivation and seeding
Machinery traffic coupled with low soil
Sediment and P loss in runoff from
Reduced axle load / larger tyres / controlled traffic on fields
strength and high moisture content leading to
sloping land following heavy rainfall
prone to soil erosion / avoid wet conditions / shallow
soil compaction
cultivation to increase bearing strength
Topsoil disturbance increases soil aeration
Soils with less than approximately 3.4%
Non-plough systems such as minimum tillage (this can give
which in turn increases SOM decomposition
SOM can be considered erodible which
rise to problems with less reliable establishment, grass weeds
thus lowering soil structural stability
can lead to sedimentation in rivers
and compaction) / residue and organic matter incorporation /
Loose fluffy very fine seed bed following
Rill and gully development leading to
excessive cultivation
very high sediment losses in concentrated
organic manures / cover crops / mixed rotations
Avoid excessive cultivation particularly on light soils
overland flow
Down-slope movement of soil by mechanical
Increased potential for P transfer to
Contour tillage and cultivation / contour grass strips / proper
tillage on sloping land
aquatic systems by water erosion
land management reflecting site specific conditions
Application of
Machinery traffic on tramlines using narrow
Concentrated flow path for sediment and
Tramline management (disruption of the compacted tramline
pesticides/herbicides/fertiliser
(row crop) tyres can lead to intense
P loss in runoff throughout the growing
surface to a depth of 60 mm with a tine)
compaction.
season.
Harvesting of crop / application of
Sub-soil compaction resulting from very high
May contribute to erosion and P loss
Large tyres or tracks on combine harvester and large tyres on
slurry
axle loads
particularly with the latter
slurry spreader
Post-harvest cropping
Fallow/bare soil leading to net C loss due to
Bare soil with low SOM can have poor
Reduced winter fallow by using winter and cover crops –
an absence of C uptake
structure and is particularly susceptible to
volunteer growth also helps / crop residue incorporation
erosion
Long-term cultivation
Compaction and impaired structure
Reduced SOM/annual C uptake
Sediment and associated P loss in runoff
Residue and organic matter incorporation / Land-use change
following heavy rainfall
– rotation with grass
Increased erosion with greater potential
Residue and organic matter incorporation / Land-use change
for sediment delivery to waterways
– rotation with grass
30
Chapter 2
The potential total P (TP) losses associated with these estimates of erosion could have
serious implications for water quality in Ireland if the eroded sediment reaches surface
water bodies, given that the typical range for TP content of non-polluted agricultural
soils in Ireland is estimated at between 0.02 to 0.2% (McGrath et al., 2001) with the
median TP content of Irish soils (0-10 cm) being 0.11% (Fay et al., 2007). Applying
this range of TP in Irish agricultural soils to the estimates of water erosion reported by
Van Oost et al. (2009) for Ireland gives a range of 0.88 to 8.8 kg P ha -1 yr-1 lost from
arable land. This is a conservative estimate of P loss from arable fields due to erosion,
given that P has low solubility and is primarily bound to finer soil fractions like clay,
which runoff preferentially transports (Quinton et al., 2001). In a study of TP export
coefficients from different CORINE land cover classes in 50 experimental subcatchments of the rivers Colebrooke and Upper Bann in Northern Ireland, Smith et al.
(2005) determined the TP export coefficient from non-irrigated arable land to surface
waters to be 4.88±1.12 kg P ha -1 yr-1 with 95% confidence limits. This export
coefficient was almost twice as high as that measured during the study for any other
CORINE land cover class and almost five times as high as the export coefficient for
pasture.
With the exception of tillage erosion occurring adjacent to waterways, soil transported
in the field by mechanical tillage operations is unlikely to reach surface waters
without transportation by water erosion. Though tillage erosion does not have the
same direct detrimental effect on surface water quality as water erosion, it can
increase the risk of nutrient delivery to waterways by progressive accumulation of
nutrient-rich sediment in low-lying areas of fields which may be exposed to
concentrated overland flow and leaching, and, therefore, it must be accounted for in
any assessment of soil erosion. If field- and catchment-scale research is to identify
sediment sources and test mitigation options within arable areas, it must be designed
to explicitly attribute losses to tillage or water erosion processes. This will require
assessment of each type of erosion in isolation and while interacting with one another.
This is essential if we are to understand the role played by tillage erosion in delivering
sediment to surface waters.
31
Chapter 2
Rainfall variability in Ireland often results in tillage field operations being carried out
in less favourable conditions (soils at or near field capacity) with increased risk of soil
compaction from field machinery traffic. The trend towards larger machines with
increased axle loads further increase the risk of soil compaction. Compacted soils with
poor structure are more prone to surface capping and poor infiltration of water due to
reduced porosity and consequent reduction in hydraulic conductivity, leading to
earlier saturation and thus increased surface runoff and erosion in sloping areas. Soil
compaction can occur as surface compaction i.e. within the tilled layer or as subsoil
compaction which occurs beneath the plough layer. Surface compaction is normally
dealt with in the next tillage operation, while subsoil compaction is much more
persistent and difficult to remove. While sub-soiling has been the subject of much
research and can reduce bulk density and compacted layers, it is generally considered
better to avoid subsoil compaction than to rely on alleviating the compacted soil layer
afterwards (Alakukku et al., 2003; Spoor et al., 2003). Prevention of subsoil
compaction is essential for economically and environmentally sustainable agriculture
(Arvidsson et al., 2000). Compaction can be reduced by: (1) use of low ground
pressure wheel equipment on machinery (Chamen et al., 2003); (2) working in good
soil moisture conditions and minimising the weight of machinery (Van den Akker et
al., 2003); (3) minimising the number of passes of machinery (Marsili et al., 1996)
and (4) controlled traffic systems (Chamen et al., 2003).
2.2.5 Processes and mechanics of erosion and deposition
The processes governing the detachment, transport and deposition of sediment are
intricate and interactive in nature, and must be fully understood if rates of soil erosion
are to be accurately determined.
2.2.5.1 Raindrop erosion
Rainfall characteristics influence processes affecting infiltration, runoff, soil
detachment, and sediment and chemical transport (Truman, 2007). Rainfall intensity
(I) is generally considered to be one of the most important factors influencing soil
erosion by overland flow and rills because it affects the detachment of soil particles
32
Chapter 2
by raindrop impact and enhances their transport by runoff. It was reported by Guy et
al. (1987) that 85% of sediment transported from inter-rill areas could be attributed to
enhancement of transport capacity of runoff by raindrop impact, while only 15% was
attributed to undisturbed runoff. Meyer and Wischmeier (1969) showed that in natural
rainfalls, soil detachment by raindrop impact is proportional to the rainfall intensity
squared (I2). Soil detachment resulting from raindrop impact is a function of raindrop
size, mass, velocity, and kinetic energy, raindrop impact angle, soil strength, flow
depth, sediment load and surface sealing (Zhang et al., 2003). Erosion due to rainfall
impact can dominate erosion due to overland flow in shallow sheet flow of limited
length over gentle slopes (Proffitt and Rose, 1991), or in flow between rills (Marshall
et al., 1996). In inter-rill areas, particle transport by raindrop-impacted shallow flow is
the dominant transport process (Zhang et al., 2003).
Ground cover, in the form of vegetation and crop residue, acts as a protective layer
between the soil surface and falling raindrops by absorbing the kinetic energy that
would otherwise be imparted to the topsoil. Rain intercepted by the crop canopy either
evaporates or finds its way to the soil by dripping from leaves, or by running down
plant stems as stemflow. The action of direct throughfall and leaf drainage on the soil
produces rainsplash erosion, the most important detaching agent in the erosion
process (Figure 2.4). Rainsplash erosion can also be reduced by manure application,
which enhances soil aggregation. Arable land is particularly susceptible to rainsplash
erosion, as it can be found bare and unprotected during stages of the arable cycle. The
impact of raindrops on topsoil can contribute to erosion in two ways: first, by
breakdown of the structure of the soil; and, second, by adding directly to the
concentration of sediment in surface water (Rose, 2004).
33
Chapter 2
Figure 2.4 Raindrop falling on a bare unprotected soil surface and detaching soil
particles from the original soil matrix: a) raindrop velocity, b) raindrop impact
producing a disrupting force in the form of laterally flowing jets (Saavedra,
2005)
For detachment to occur, the raindrop must overcome the interstitial forces holding
the primary soil particles together in bundles or aggregates. The detachability of the
soil, therefore, depends not only on the soil texture, but also the shear strength of the
topsoil (Cruse and Larson, 1977). Raindrops that impact on the soil bed exert shear
stress which can be very high locally, compared to shear stresses commonly exerted
by overland flow, given relatively shallow water depths (Rouhipour et al., 2006). A
study by Poesen (1985) to determine the kinetic energy required to detach 1 kg of soil
by rainfall impact showed that particles between 0.063 and 0.25 mm in size are the
most detachable by raindrop impact. The coarser soils are resistant to detachment
because of the weight of their larger particles, and the finer soils are resistant because
the raindrop energy has to overcome the adhesive or chemical-bonding forces that
link the minerals comprising the clay particles (Morgan, 2005). For this reason, silt
loams, loams, fine sands and sandy loams are the most detachable soil types. The low
OM content of sandy soils further disposes them to rainsplash erosion, as they have
low shear strength. In some soils, the process of structural disaggregation, due to
raindrop impact, can result in a surface seal or crust forming. This phenomenon
decreases the infiltration rate, reduces the available water to the plant in the root zone,
diminishes the natural recharge of aquifers, increases runoff and soil erosion, affects
34
Chapter 2
seedlings and plant growth, and decreases crop yields (Assouline, 2004), and is
particularly common in sandy soils when subjected to rainfall.
When the rainfall rate exceeds the rate of infiltration of the soil, water starts to pond in
surface depressions or hollows, and runoff does not begin until the surface storage
capacity is exceeded. This shallow overland flow rarely exerts enough shear force on
the original soil matrix to detach particles from it, but it does carry sediment dislodged
by raindrop impact (Trout, 1993). In inter-rill areas, raindrops impacting on this
shallow overland flow can increase its transport capacity. Rainfall detachment
decreases with increased flow depth (Moss and Green, 1983) due to cushioning of the
raindrop impact by the water layer (Ferreira and Singer, 1985). It has been shown
experimentally by Palmer (1965) and Ghadiri and Payne (1981) that the maximum
raindrop impact occurs for water depths of less than one raindrop diameter.
Furthermore, Proffitt et al. (1991) found that raindrop impact becomes negligible for
flow depths greater than 3 drop diameters.
While the rainfall proceeds, runoff is continuously splashed by falling raindrops,
further breaking down the soil particles it is carrying and helping to keep them in
suspension. Therefore, these finer particles, which are typically more effective at
absorbing the plant nutrients (P and N), will travel further (Rose and Dalal, 1988), are
more likely to reach waterways, and may lead to eutrophication. Coarser particles that
do settle out of solution form a deposited layer on the soil surface. The sediment in the
deposited layer is much more easily eroded than the original soil matrix, as it does not
have time to form cohesive bonds with neighboring sediment. The deposited layer
also partially protects the original soil matrix from further raindrop impact. When
sediment is removed from the deposited layer by raindrop impact, it is termed „rainfall
re-detachment‟ (Figure 2.5). Under the same rainfall, re-detachment of the deposited
layer will occur at a much higher rate than detachment of the original soil if that soil is
cohesive (Rose, 1988).
35
Chapter 2
Figure 2.5 Schematic showing the distinction between rainfall detachment of the
original soil matrix, and re-detachment of previously eroded and deposited
sediment (from Rose, 1993).
2.2.5.2 Flow-driven erosion
For flow-driven erosion to occur, a threshold stream power value must be exceeded.
The stream power is a product of the shear stress exerted by the flowing water on the
soil surface and the velocity of that water, and is dependent on the nature of the soil
surface. Roughness elements at the soil surface (e.g. crop residues, rock fragments,
vegetation, geotextiles) strongly reduce the erosivity of overland flow (both inter-rill
and concentrated overland flow) and, hence, soil detachment rates (Knappen et al.,
2009). An increase in the depth or flow rate of the surface water will increase the
stream power. Once the force exerted by the flow exceeds the forces keeping the soil
particle or aggregate at rest, flow-driven erosion will occur.
Flow-driven erosion can be broken up into 3 processes: entrainment, deposition and
re-entrainment. Entrainment is defined as the removal by flowing water of the original
cohesive soil by the mutual shear stress between the soil surface and the water
flowing over it. The rate of entrainment is, therefore, strongly influenced by soil
physical properties, in particular, cohesion/strength, and also by hydraulic
characteristics of flow. The length of time eroded soil particles remain in suspension
before settling back to the soil surface in deposition is dependent on their settling
36
Chapter 2
velocities, which, in turn, depends on their size. Coarser particles having higher
settling velocities will settle out first while finer particles may be transported great
distances before falling out of suspension. Once returned to the soil surface, these
particles have insufficient time to form cohesive bonds with neighboring particles and
are, therefore, much more easily removed in a process termed „re-entrainment‟. It is
the ease of removal of this cohesionless, deposited soil that distinguishes the reentrainment process from that of the entrainment process (Rose, 1993). In effect, a
soil particle or aggregate transported by the processes just described will make a
consecutive series of hops while transitioning from a position of rest on the soil bed
up into to the flowing surface water layer and back to the soil bed. The size and
trajectory of these hops are dependent on the particles settling velocity in water. This
hopping motion is commonly referred to as „saltation‟ and is common-place in
shallow overland flow.
Flow-driven erosion is commonly differentiated into sheet erosion and rill erosion
(Mulqueen et al., 2006). Sheet erosion (or interrill erosion) removes a thin layer of
soil, whereas rill erosion excavates a discrete centimeters-deep channel with the
hydraulic power of concentrated overland flow (Whiting et al., 2001). Generally,
sheet flows are a result of high-intensity, short-duration rainfall events. Sheet flow
occurs as a shallow sheet of water flowing for short distances over gently sloping
land. The resulting erosion is termed „sheet erosion‟ and removes a uniform layer of
soil from the soil surface, which is often only recognizable when this soil is deposited
at the bottom of a slope or near a fenceline. Typically, sheet flow does not have
sufficient depth or velocity to detach soil particles from a bare soil surface, and acts
primarily as a transport mechanism for soil previously dislodged and disaggregated by
raindrop impact. As such, sheet flow normally results in the loss of the finer particles,
such as clay, silt and OM. Plant nutrients are generally more concentrated on these
finer particles such that even modest loss of them is of considerable practical concern
to both agriculture and the quality of receiving waters (Rose et al., 2006). Though
rarely seen, sheet erosion accounts for large volumes of soil loss from cultivated land
each year in Ireland.
37
Chapter 2
Tillage disturbance, natural microtopographic variation, and drainage patterns which
develop as a result of the soil erosion process itself can all produce concentrated flows
(Hairsine and Rose, 1992). Therefore, sheet flow occurring over relatively rough
surfaces begins to form small concentrated channels at a critical distance downslope.
Once the rills become sufficiently developed, they can capture much of the rainfallgenerated flow from the interrill area, and function as a sediment source and delivery
system for erosion on hillslopes. In rills, the dominant processes are entrainment and
re-entrainment by concentrated flow aided by mass movement of soil into the rill due
to sidewall sloughing and slips, undercutting of sidewalls, and head cutting of rills
(Mulqueen et al., 2006). Generally, the development of rills is accompanied by a
dramatic increase in erosion. This is due to the enhanced streampower in the rill
(Marshall et al., 1996). Streampower has been related to detachment rate in rills by
Nearing et al. (1997) and Hairsine and Rose (1992), and was shown to be the best
parameter to predict detachment capacity for rills by Elliot and Laflen (1993).
2.2.6 Measuring and quantifying soil erosion on arable land
There is a need for information on both gross and net erosion rates from agricultural
land, so that the sediment delivery ratio (SDR), or proportion of the sediment
mobilised by soil erosion that is transported towards local watercourses, rather than
being deposited close to the original source, can be determined (Blake et al., 1999). If
the level of erosion of Irish tillage soils is to be accurately determined, work must be
undertaken that quantifies rates of soil movement to surface waters at the catchment
scale.
Traditional monitoring techniques used to establish soil erosion rates have the
inherent flaw of failing to determine the fate of eroded sediment and, therefore, give
no indication of the impact of measured erosion rates on surface water quality. Blake
et al. (1999) note that it is particularly difficult to assemble information on the spatial
distribution of erosion and deposition rates within the landscape, and on the associated
SDRs using traditional monitoring techniques. Much of the information available on
erosion rates has been collected from flume and erosion plot studies; however, these
only provide information on the net rate of soil loss from the bounded area, as
38
Chapter 2
represented by the flux of sediment across its lower boundary. As such, plot studies
typically overestimate erosion rates by failing to encompass major catchment
sediment stores (Collins and Walling, 2007). These stores get larger as catchment area
increases because the fraction of less steep slopes, like valley bottoms where sediment
deposition occurs, also increase (Verstraeten and Poesen, 2001). It is for these reasons
that the representativeness of plot results in terms of the wider landscape is often
questioned. As the scale at which erosion is being studied increases from flume-toplot and up to field- and catchment-scale, the parameters influencing this erosion
change and, therefore, so must methods used to measure erosion. The use of sediment
fingerprinting and composite fingerprints to determine the provenance of eroded
sediment is one preferable method at larger scales which will be discussed in Section
2.6.1.
2.3 Phosphorus transfer from agricultural soils to surface water bodies
The movement of P from agricultural soils to water through the action of rainfall and
overland flow has serious implications for surface water quality in Ireland.
Agriculturally derived P is estimated to account for 70% of P loads in rivers and
estuaries in Ireland (EPA, 2004). Diffuse losses from agriculture were reported by
McGarrigle and Donnelley (2003) to account for 59% of TP exported from a rural
Irish catchment.
2.3.1 Sensitivity of surface waters to eutrophication
Phosphorus is a naturally occurring element in the environment essential for
agricultural crop and livestock production. Concentrations of P present in streams,
rivers and lakes exceeding critical values for algal growth can lead to eutrophication
(Carpenter et al., 1998; Pote et al., 1999; Sharpley, 2000; Haygarth et al., 2005). As
DRP is readily available for biological uptake, it poses an immediate threat for
accelerated algal growth, which may negatively affect water quality in rivers and
lakes (Sharpley and Smith, 1989).
39
Chapter 2
In 2002, the World Health Organisation (WHO, 2002) reported that when P is the
limiting factor (i.e. if N:P is greater than 7:1 in the water column), a phosphate
concentration of 0.01 mg L-1 is enough to support plankton, and concentrations from
0.03 to 0.1 mg L-1, or higher, will be likely to promote algal blooms. Empirical
comparison of in stream phosphate levels and biological quality has demonstrated that
once median phosphate concentrations exceed 0.03 mg L -1 P, significant deterioration
is seen in Irish river ecosystems (Clabby et al., 2008). The sensitivity of surface
waters to eutrophication from diffuse agricultural P losses is further highlighted by the
relatively low concentrations at which eutrophication can occur: 0.03 mg L−1 of MRP
in rivers and 0.02 mg TP L−1 in lakes (Lucey et al., 1999). These are an order of
magnitude lower than DRP concentrations in the soil solution necessary to support
plant growth (0.2 - 0.3 mg P L−1) (Heathwaite and Dils, 2000; Aase et al., 2001;
Sharpley et al., 2003a).
Prior to July 2009 water quality standards for P in Ireland were set out in the
Phosphorus Regulations (SI 258 of 1998). In these regulations, rivers having annual
median phosphate concentrations of < 0.03 mg L -1 were classed as unpolluted. New
environmental quality standards for chemical and physico-chemical elements in rivers
were brought into law in the European Communities Environmental Objectives
(Surface Waters) Regulations (SI 272 of 2009). These regulations address the
requirements of the WFD and also repeal SI 258 of 1998. Furthermore, they define
rivers with MRP ≤ 0.035 mg L-1 as being in good status and therefore not polluted.
Therefore, for good water quality in Irish water bodies, it is considered that P
additions from all sources should not give rise to a concentration in the water of
greater than 0.035 mg MRP L-1. A value of 0.03 mg MRP L-1 represents a more
conservative figure as it lies midway between good (≤ 0.035) and high (≤ 0.025)
status. The model of Donohue et al. (2006) which links catchment characteristics and
water chemistry with the ecological status of Irish rivers also supports the use of 0.03
mg MRP L-1 as the environmental standard for river water quality.
40
Chapter 2
2.3.2 Soil test phosphorus
It is generally accepted that there is a positive relationship between STP and P loss to
water in runoff events (Tunney et al., 2000; Vadas et al., 2005). In Ireland, Morgan‟s
extractant (Peech and English, 1944) is currently used to match P fertiliser
recommendations with crop requirements. Phosphorus advice for grassland and tillage
crops in Ireland is based on a four- category soil P-index system (Table 2.2). The
basis of this system is a set of soil indices based on the measured Morgan‟s P (P m) in
the soil and the crops response to fertiliser application as measured by field
experimentation. For tillage soils at P Index 4, the addition of P is prohibited with the
exception of soils planted with potatoes, beet, and turnips.
The current agronomic optimum P m value for Irish soils is 6 mg L-1 for grass
production (Daly et al., 2001). In Ireland, low soil Pm concentrations of 1 mg L-1 in
the 1950s severely limited crop production. Since then, continual fertiliser application
at high rates on agricultural land has resulted in excessive levels of plant available P
in soils (Tunney, 2000). Although national sales of P fertiliser have fallen from
62,410 ton yr-1 to 26,350 ton yr-1 during the period 1995-2008 (DAFF, 2009b),
primarily due to new farming practices, implementation of the Nitrates Directive and
rising fertiliser costs, the mean P m concentration in Irish soils is currently 8 mg L -1
(Daly et al., 2001). Maintenance of the P fertility of arable soils is important as cereal
crops perform better in soils of good P status (6.1-10 mg L-1 Pm) than on soils of low
P status that have been supplemented with higher levels of P fertilisers (Schulte et al.,
2010b). Fertiliser advice is modified for some tillage crops, according to crop yields,
soil texture, or expected summer rainfall amount (Coulter and Lalor, 2008).
As soil P increases, P loss in surface runoff and subsurface flow increases (Sharpley et
al., 2001b). Therefore, the higher the P m level in fields of a catchment, the greater the
risk of high concentrations of in-stream P during wet months (Lewis, 2003). Previous
grassland studies at plot (Pote et al., 1999) and field (Tunney et al., 2000) scale have
shown that there is a positive relationship between the Pm level in soils and DRP lost
in surface runoff. Schulte et al. (2010a) developed a model of STP decline on eight
principal soil series/associations representative of a range of STP concentrations for
41
Chapter 2
grassland in Ireland and found that where the P m was at 28 mg L-1 and with no further
P inputs (estimated to be equivalent to an annual field P-balance deficit of 30 kg ha -1
yr-1), it took from 7-15 years for a soil to move from Index 4 to Index 3 (Table 2.2).
Almost half the river monitoring sites sampled for phosphates in the South Eastern
River Basin District (SERBD) - where tillage is common - in 2008 would not achieve
good status based on this nutrient (Lucey et al., 2009). All lakes assessed from 2007
to 2009 in the SERBD were of moderate or poor ecological status largely due to TP
and chlorophyll, possibly related to intensive agriculture (McGarrigle et al., 2010).
Table 2.2 Phosphorus Index System (from SI 610 of 2010 and adapted from
Schulte et al., 2010a)
Soil P
Soil P ranges
Index
(mg L -1)
1
Interpretation
Grassland
Tillage
0.0-3.0
0.0-3.0
Soil is P deficient; build-up of
soil P required.
2
3.1-5.0
3.1-6.0
Low soil P status: build-up of
soil P is required for productive
agriculture
3
5.1-8.0
6.1-10.0
Target soil P status: only
maintenance rates of P required
4
> 8.0
> 10
Excess soil P status: no
agronomic response to P
applications.
2.3.3 Phosphorus use on tillage land
While tillage land accounts for a relatively small area (9.6% of agricultural area
utilised in Ireland (CSO, 2009)), it accounts for a lot of the high P status soils due to
higher fertilisation rates on tillage land, and may therefore make a disproportionate
contribution to the TP input to surface water systems from agricultural soils. Mean P
fertiliser use in Ireland for cereals and root crops (less than 10% of tillage area) in
42
Chapter 2
2008 was 20 and 46 kg ha -1, respectively, while P fertiliser use for grassland was only
5 kg ha-1 (Lalor et al., 2010). These figures highlight the potential for higher P losses
in surface runoff from tillage land than from grassland.
Excessive manure and fertiliser application is not only wasteful, but it can lead to a
build-up of P in excess of crop requirements in the soil. The excess P may then be
mobilised by surface runoff during periods of heavy rainfall. The United States
Department of Agriculture estimates that about half the fertiliser used each year in the
United States simply replaces soil nutrients lost by topsoil erosion (Montgomery,
2007). Soil test P, accumulated to very high concentrations, can take up to 20 years of
continual crop harvesting - with no addition of P from any source - to reduce to
concentrations normally recommended for agronomic production and to pose no
threat to surface water quality (Sharpley and Rekolainen, 1997).
Tillage land has higher P application rates than grassland due to the higher offtakes
and the need for new seeding each year. Sufficiently high available P levels are
needed for satisfactory seed germination. Advice given to farmers on P application for
cereal crops is based on maintaining the STP at the agronomic optimum level of Index
3 (Table 2.2). This is achieved by applying enough P to replace the anticipated crop
off-take (a grain yield of 1 ton ha-1 = an off-take of 3.8 kg P ha-1), based on the
expected yield of the crop to be fertilised (Coulter and Lalor, 2008). Where proof of
higher yield is available, an additional 3.8 kg P ha -1 can be applied on soils at P
Indices 1, 2 and 3 for each additional tonne above a threshold crop yield dependent on
crop variety (SI 610 of 2010). Where the Soil Index is below Index 3, build-up levels
are necessary in addition to anticipated crop off-take in order to raise the Soil Index to
Index 3. Regular soil testing should be carried out to ensure that soils are maintained
within the agronomic optimum Soil Index. Root crops, like potatoes and fodder beet,
are very responsive to P and it is necessary to apply P (when sowing) even at Index 4
to achieve the agronomic optimum.
The impact of land use (agriculture) and soil characteristics (parent material and
wetness) on plant available P distribution in soils is given credence by Zhang et al.
(2008) in a geochemical mapping study of Ireland in which Pm was measured in 1310
43
Chapter 2
surface (0-10 cm) soil samples collected from pre-determined positions - at a density
of 2 samples per 100 km2 - based on an unbiased grid sampling scheme. They
delineated the areas having high available P using the index bands for tillage soils in
the P index system (Table 2.2), which state that soils having > 10 mg L -1 Pm levels
are in excess of crop requirement. The authors attributed high levels of available P in
County Louth, east Dublin and southeast Wexford to a combination of light-textured
soils, and vegetable and tillage farming in these areas. Similarly, in northwest Kerry,
tillage farming on light-textured soils resulted in elevated P levels. Furthermore, they
attributed high levels in east and central Cork to a combination of intensive dairying
and tillage on highly fertile soils, while high levels in north Carlow and south Kildare
may be due to intensive tillage on limestone-derived soils. Reducing these soil P
levels may not be possible in the short term as Schulte et al. (2010a) showed that
elevated soil P concentrations, resulting from agricultural land use, may take many
years to be reduced to agronomically and environmentally optimum levels.
2.3.4 Critical source areas of phosphorus
The loss of P tends to be highly sporadic in nature and is often restricted to small
geographic areas (Edwards and Withers, 2008). In many regions, small portions of
saturated land, known as variable source areas, generate the majority of overland flow
(Pionke et al., 2000), the amount of which is largely independent of rainfall intensity
(Walter et al., 2000). These variable source areas commonly exist because the
groundwater table is close to the land surface near the stream, causing seep zones and
high soil moisture levels that limit available storage in the soil profile (Gburek and
Sharpley, 1998). They can contract and expand both seasonally and during storms as a
function of precipitation, topography, soil type, geology, soil moisture status, and
water table level (Hart et al., 2004). Runoff generated from variable source areas is
dominated by saturation excess overland flow and, to a lesser extent, rapidly
responding subsurface flow (Gburek and Sharpley, 1998). Outside of these areas,
infiltration and groundwater recharge are the dominant hydrological processes, and
runoff generation is normally low with the exception of high intensity storm events.
Hydrologically CSAs have been defined in terms of the coincidence of sources of P
with variable source areas (Doody et al., 2012). However, for variable source areas
44
Chapter 2
located some distance from a water course, continuous hydrological connectivity
along the flowpath is necessary to link fields to water courses. Connectivity will
ultimately determine whether P source areas become CSAs and create problems in
receiving waters. As such, it is more appropriate to define CSAs in terms of their
impact on the aquatic environment with particular emphasis on connectivity between
P source and aquatic receptor. Wall et al. (2011) recently described CSAs as „regions
where high nutrient loading or status coincides with a high propensity to be
hydrologically connected to water courses‟.
A large proportion (up to 90%) of P exported from catchments on an annual basis is
generated from a relatively small portion of the catchment and during only one or two
storm events (Sharpley and Rekolainen, 1997). Tunney et al. (2000) showed that 40%
of the total amount of DRP lost in runoff for 1997 from four grassland fields ranging
in size from 0.5 – 14.5 ha was lost when about 150 mm of rain fell in a 4-d period. In
contrast, a study of nutrient and sediment loss to water from agricultural grassland
catchments of the Dripsey River, Co. Cork in 2002, found that more than 80% of TP
loss was for the five months of October to February, with a large proportion coming
from about 10 storm events where high P concentrations occurred simultaneously
with high stream flows (Lewis, 2003). This evidence suggests that, while extreme
rainfall events with large return periods like that reported by Tunney et al. (2000) can
be responsible for a large proportion of DRP lost over an atypical year, more normally
one would expect P loss to be spread across a number of large storms throughout the
year. In addition, research at plot-scale on arable land in the UK by Quinton et al.
(2001) showed that more frequently occurring smaller events accounted for a greater
proportion of the P lost over a 6-yr period than infrequent large events. It is important
to note that losses in the study of Quinton et al. (2001) were measured at the end of an
erosion plot and that even though a smaller proportion of P was lost in larger events,
these events have greater transport potential and are more likely to deliver eroded
sediment and P to surface waters.
The identification of CSAs, where the potential for pollution is higher, has significant
implications for RBMP, because the blanket application of a specific mitigation
measure across an entire catchment will not be as cost-effective as its deployment
45
Chapter 2
solely in those areas where it is most appropriate. Pionke et al. (1997) suggested that
effective mitigation of P losses from agriculture must focus on defining, targeting, and
remediating CSAs of P loss. Focusing on CSAs of P could significantly improve the
environmental efficiency and cost effectiveness of the POM adopted for the WFD
(Schulte et al., 2009; White et al., 2009). Schulte et al. (2009) and Doody et al. (2009)
developed a suite of cost-effective catchment specific P mitigation measures by
identifying and characterising CSAs on farms in the Lough Melvin catchment using a
modified P ranking scheme. Similarly, Hughes et al. (2005) used field and catchmentscale P ranking schemes to identify CSAs for P loss in Ireland. Phosphorus ranking
schemes are discussed further in Section 2.4.1. Quantifying the influence of CSAs of
surface and near surface runoff is a next phase in the Agricultural Catchments
Programme (ACP) experimental design, targeting diffuse events of P and sediment
transfer in dissimilar parts of the catchment and comparing with the continuous data at
the outlets (Wall et al., 2011).
2.3.5 Phosphorus mobilisation
Mobilisation is the first key step in the separation of P molecules from their source
and includes chemical, biological and physical processes (Figure 2.6). These
processes group into either solubilisation or detachment mechanisms, defined by the
physical size of the P compounds that are mobilised (Haygarth et al., 2005).
Solubilisation potential of P from soil surfaces and soil biota into soil water increases
with increasing concentrations of STP, which result from long-term historical addition
of fertiliser and manure in excess of crop requirements. The detachment and transfer
of non-dissolved P in association with soil particles is more pronounced where
farming practices generate erosion (Chambers et al., 2000), and provides a physical
mechanism for mobilising P from soil into surface waters (Sharpley and Smith, 1990;
Toy et al., 2002). The size threshold most commonly used to operationally define
detachment is > 0.45 µm and has been used for the threshold between dissolved and
PP (Haygarth and Jarvis, 1997).
46
Chapter 2
Figure 2.6 Processes in the transfer of P from terrestrial to aquatic ecosystems
(from Sharpley et al., 1995).
Haygarth and Jarvis (1999) have argued for the inclusion of a third mode by which P
can be mobilised for transport to water - incidental transfer of dissolved P and PP
occurring when fertiliser or manure applications, which are not incorporated into the
soil, are coincident with onset of rainfall. They conclude that even though incidental
transfer will include mobilisation and detachment, it should be kept separate from
these mechanisms due to the unique circumstances leading up to its occurrence and
control. The relative proportions of PP and dissolved P in surface run-off, therefore,
depend on the complex interaction between climate, topography, soil type, soil P
content, type of farming system, and farm management (Withers, 1999).
Particulate P encompasses all primary and secondary mineral P forms, plus organic P,
P sorbed by minerals, and organic particles eroded during runoff. It constitutes the
major proportion of P transported from cultivated land (75-90%) (Sharpley et al.,
1995). Fang et al. (2002) reported that PP contributed from 59 to 98% of total runoff
47
Chapter 2
P for unvegetated, packed runoff boxes. Unlike most DRP, which is readily available
for plant uptake, PP acts as a long-term source of P for submerged aquatic vegetation
and algal growth (Sharpley, 1993; Søndergaard et al., 2001), particularly in lakes
where inflowing rivers deposit nutrient-enriched sediment on the lake floor.
Phosphorus release at the sediment-water interface may occur in the following
conditions: (1) during periods of anoxia or hypoxia (Theis and McCabe, 1978;
Steinman and Ogdahl, 2008); (2) by wind-induced resuspension and bioturbation
(Steinman and Oghahl, 2008); or (3) when there is an increase in pH of the interstitial
water (Sharpley and Rekolainen, 1997; Daly, 1999).
Soil cultivation is a major factor contributing to an increased risk of PP transfer to
water, but when reduced cultivation such as non-plough tillage is practised to decrease
losses of PP, there can be a build up of P near the soil surface, which increases the
risk of DRP loss in surface runoff. Disturbance of soil structure by tillage operations
also increases aggregate dispersion and the degree of interaction between soil and
runoff water, thereby enabling more dissolved P to be mobilised from soils with high
P (Sharpley et al., 2001a). Diffuse P loss from arable land can be as high as 1-2 kg P
ha-1 yr-1 in the northern temperate zone, especially in areas with widespread soil
erosion (Ulén et al., 1991).
2.3.6 Hydrological pathways of phosphorus transport
The hydrological pathways of P movement from fields include surface runoff
comprising overland flow, and subsurface flow comprising preferential flow,
interflow, groundwater discharge and drainflow (Brogan et al., 2001). All of these
pathways have the potential to contribute to P export depending on the connectivity
with adjacent water bodies (Doody et al., 2012). It is the landscape position of a P
source, both in terms of its upslope contributing area and its downslope flow path, that
determine the likelihood of a connection being made to receiving waters. Surface and
near surface pathways can be considered as the main link between source and delivery
of P (Wall et al., 2011). Both particulate and dissolved P can be lost via these
pathways. Fingerprinting studies show that the majority of SS in rivers is derived
from the surface soil (Walling, 2005). In the surface pathway, overland flow is the
48
Chapter 2
main pathway of diffuse P loss from agricultural soils in Ireland (Kurz et al., 2005;
Tunney et al., 2007) following hydrological (storm) events. Surface runoff has a
strong affinity for P transport because the surface soil has the greatest effective depth
of interaction and the highest concentrations of P (Heathwaite et al., 2005). However,
research has shown that P export from catchments can also occur via subsurface
pathways, with the most significant instances of subsurface movement of P being
associated with excessive application of P in manure and fertiliser (Sims et al., 1998).
In general, the transport of PP in subsurface flow is not large. Particulate phosphorus
is the dominant P fraction exported from arable land during overland flow events due
to soil erosion (Doody et al., 2012).
When attempting to identify the primary flow paths of nutrient transport in fields
where surface pathways dominate, attention should be paid to surface runoff, flow in
tramlines and tyre tracks, and flow along roads or other impermeable features
(Heathwaite et al., 2005). In cases where subsurface pathways dominate, attention
should be paid to land drains, near surface interflow, deeper subsurface storm flow
and groundwater flow (Heathwaite et al., 2005). These pathways may or may not be
activated depending on criteria such as antecedent moisture, topography, rainfall
intensity and duration (Heathwaite and Dill, 2000). The identification of primary flow
paths of nutrient transport is essential for water quality protection because these routes
form the critical link between P sources and P outputs measured in streamflow.
Furthermore, the introduction of pipe drainage systems and tramlines, the presence of
compacted headlands and gateways and/or the proliferation of ditches, tracks and
roads, has greatly lengthened the distance over which sediment and P can be
transported before reaching a water body (Withers et al., 2007). In situations where
reducing soil P to environmentally acceptable levels will take many years of restricted
fertiliser use (based on nutrient management planning and soil P testing), methods of
flow path manipulation (such as buffer zones or sediment traps) that reduce
connectivity between P source and receiving waters should be used, as these can
immediately reduce the amount of P reaching surface waters. In a study by Schulte et
al. (2009) to identify the dominant P pressure and pathway risks governing P loss in
the catchment, and to evaluate and select potential mitigation measures based on an
assessment of cost-effectiveness and farmer preference, installation of sediment traps
49
Chapter 2
in drainage ditches was identified as the most cost-effective and popular measure
aimed at reducing P transport vectors in the short term.
2.3.7 Rainfall simulation as phosphorus research tool
The expensive nature of field experiments and inherent variability in natural rainfall
has made rainfall simulators and laboratory microcosms a widely used tool in P
transport research (Hart et al., 2004). Due to the complexity of soil erosion by water,
field experimentation can be complimented by hypothesis testing in controlled
reductionist laboratory experiments. Conclusions drawn from these reductionism
experiments help to „reduce the uncertainty in explanation of complex patterns‟
occurring at field- and catchment-scale (Haygarth, et al., 2005). While there are still
some reservations regarding the use of simulated rainfall in place of natural rainfall
(Potter et al., 2006), there is widespread support for the use of rainfall simulation
experiments to obtain some estimate of the magnitude of potential losses from
different land management systems, soil types, and landscapes (Pote et al., 1999;
Sharpley et al., 2001b; Bundy et al., 2001; Schroeder et al., 2004; Tarkalson and
Mikkelsen, 2004; Little et al., 2005). Numerous studies – outside of Ireland – have
utilised rainfall simulation to evaluate nutrient losses in runoff from tillage systems
(Zhao et al., 2001; Daverede et al., 2003; Franklin et al., 2007). Studies have also
been conducted using laboratory rainfall simulation on runoff boxes packed with
tillage soil to predict runoff of SS and PP using simple soil tests (Udeigwe and Wang,
2007), and to examine variability in mobilisation and transport of nutrients and
sediment by overland flow across a range of soils (Miller et al., 2009). In addition,
flume studies using concentrated overland flow as opposed to simulated rainfall have
been used by Knappen et al. (2008) to show that the effect of conservation tillage on
soil detachment rates is a result of soil property modifications affecting soil
erodibility, rather than a result of the surface residue decreasing flow erosivity.
Laboratory-scale work such as this is essential in understanding erosion processes and
in selecting suitable erosion prevention measures for further testing at larger scales.
50
Chapter 2
2.4 Phosphorus loss risk assessment tools
2.4.1 Phosphorus risk index
The original P risk index of Lemunyon and Gilbert (1993) was designed as a
screening tool for use by field staff, catchment planners, and farmers to rank the
vulnerability of sites to P loss in surface runoff. It is a simple, field-scale analysis,
which integrates soil test data, soil erosion and runoff potentials; and P fertiliser or
organic waste application rate, method, and timing (Sims et al., 1998). It was
developed with a caveat that it would require modification to account for regional
variations in agricultural management practices, climate, topography, hydrology and
surface water characteristics (Hughes et al., 2005). The majority of P indices currently
in use are modified versions of the Lemunyon and Gilbert (1993) index that have been
made suitable for local conditions where they are being applied. The widespread
adoption of the indexing concept in America (at least 49 states) shows the consensus
among scientists, the fertiliser industry and policymakers with regard to the validity of
the P risk index approach (Sharpley et al., 2003b). Phosphorus risk indices have also
been developed in Sweden, Ireland, Norway and Denmark.
Despite its increased popularity, surprisingly few studies have been carried out to
show that the ranking of a P risk index actually reflects the ranking of P transfer from
fields (Bechmann et al., 2007). In Ireland, Magette (1998) developed a P ranking
scheme for Irish conditions which was modified by Magette et al. (2007) and is based
on the P risk index approach. Hughes et al. (2005) tested the P ranking scheme on 3
fields in Ireland and found that it correctly predicted the rank order of P losses from
the fields. At the plot scale, Sharpley et al. (2001b) accurately predicted the potential
for dissolved P loss (R2=0.79) and total P loss (R2=0.83) from manured plots (2 m2)
using the Pennsylvania P index. As this was a plot scale study, factors associated with
the off-site transport of P could not be assessed. At the subcatchment/field scale,
Bechmann et al. (2007) correctly ranked catchments in terms of potential P transfers
(R2=0.66) using the Norwegian P index which is based on the Pennsylvania P index.
The Norwegian P index was also tested using monitoring data for six catchments
51
Chapter 2
(ranging in size from 168 – 8700 ha) in Norway, which showed that 79% of the
variation in TP losses was explained by the P index rating.
The modified P ranking scheme of Magette et al. (2007) was designed specifically for
use in catchments in Ireland (the majority of which are dominated by grassland use),
and, as such, the soil erosion factor used in it, only provides a coarse estimate of the
risk of P loss due to erosion. The erosion risk factor is based on a field being classed
as either well managed pasture (low risk), poorly managed pasture (medium risk), notill crop systems (medium risk), or row crops under tillage (high risk). Most versions
of the P risk index used in America utilise the Revised Universal Soil Loss Equation
(RUSLE) (Renard et al., 1994) to estimate annual soil loss in fields where a P risk
index is to be determined. In presenting a „conceptual framework for a CSA approach
to the development of supplementary measures for mitigating diffuse P export in Irish
catchments under the EU WFD‟, Doody et al. (2012), proposed that data collected
using the modified P ranking scheme of Magette et al. (2007) be used to characterise
CSAs in Irish catchments so as to identify the key factors controlling P loss to water.
If the CSA approach, proposed by Doody et al. (2012), is to distinguish between key
factors controlling P loss to water, from grassland and arable land, then the risk of P
loss associated with eroded soil needs to be better accounted for by using RUSLE to
estimate soil loss.
2.4.2 SCIMAP (Sensitive Catchment Integrated Modelling and Analysis
Platform)
Given an observed downstream water quality degradation, and provided this can be
attributed to diffuse sources, the primary challenge is to determine which parts of the
landscape (subcatchments, farms, or fields) are most likely to be contributing to that
degradation (Reaney et al., 2011a; Lane et al., 2006), and are therefore CSAs. The
Sensitive Catchment Integrated Modelling Analysis Platform (SCIMAP) provides a
framework for understanding the probable spatial origins of diffuse pollution
problems within agricultural catchments. It is based upon a conception of catchments
as organising entities: catchments can be conceptualised as a set of flow paths that
accumulate distributed sources of possible contaminants from across the landscape
52
Chapter 2
into receiving waters where, for surface waters, diffuse pollution may become
„visible‟, either through detection of temporal changes in water quality via routine
monitoring (e.g. elevated nitrate concentrations) or through the more limited evidence
from physical water quality deterioration (e.g. algal blooms; or long-term changes in
ecological quality) (Lane et al., 2009; Reaney et al., 2011a; Milledge et al., 2012). As
such, it focuses on the question of „where in the catchment is the pollution load
coming from‟ rather than aiming, as many models do, to produce estimates of the
actual pollution load.
SCIMAP uses a simple approach for determining the probable relative risk of a point
in the landscape producing pollution (Lane et al., 2006). The SCIMAP risk mapping
framework comprises two dimensions of analysis: the delimitation of hydrologically
connected source areas or CSAs; and the accumulation of these CSAs through to
locations of concern (Lane et al., 2006). Surface hydrological connectivity is assessed
through analysis of the potential pattern of soil moisture and saturation within the
landscape. Assessment of the ability of each point in the landscape to generate
saturated overland flow is done using the prediction of the spatial pattern of soil
moisture and allows the probability of continuous flow to the river channel network to
be assessed. The export of risk (e.g. sediment) in surface flow from a point on the
landscape is dependent on each downslope point also being saturated, otherwise the
risk will be captured and not reach the river channel. The total risk that a point on the
landscape represents is a function of the risk of connectivity to the river channel and
the point scale risk (e.g. sediment). These risks are accumulated through the
catchment from field through to river so as to identify and map points on the
landscape where there is a high risk of diffuse pollution impacting on the aquatic
receptor. It should be noted that SCIMAP‟s hydrological treatment is most suited to a
surface and shallow subsurface flow regime, where residence times are short and
flows are predominantly lateral rather than vertical (Milledge et al., 2012). A full
description of the SCIMAP model is provided in Reaney et al. (2011a), who show
how it can be used to understand the relationships between land use, hydrological
connectivity and salmonid fry abundance.
53
Chapter 2
A range of mitigation measures are available to reduce sediment and P loss from
arable land (Kronvang et al., 2005; Deasy et al., 2009 and 2010; Stevens et al., 2009)
and decision support tools are now required to help target these measure most
effectively (Withers et al., 2007). The SCIMAP approach can be rapidly applied to
locate mitigation measures within the landscape in a systematic, targeted and cost
effective way (Reaney et al., 2011b). It is not surprising, therefore, that a high
resolution digital elevation model (5 m × 5 m) and the SCIMAP approach were
recently used by Wall et al. (2011) to distinguish areas within catchments of the ACP
with the propensity for low to high hydrological connectivity. The authors combined
soil P data with hydrological connectivity outputs from the SCIMAP assessment to
get an estimate of CSAs within the catchment at a scale of less than 2 ha (Figure 2.7).
Figure 2.7 Field-by-field soil P status (left) collated by index and showing below
optimum to the north and above optimum to the south. The SCIMAP
connectivity map (centre) on a 5 m pixel scale is also collated as a mean score per
field (right) and indicates CSA potential in the south where the propensity for
runoff (score closer to 1) is higher (from Wall et al. (2011)).
54
Chapter 2
2.5 Mitigation measures to prevent sediment and phosphorus loss from tillage
soils
Research to evaluate the effectiveness of well-established mitigation options for
prevention of soil erosion and reduction of P loss from arable land was carried out in
studies by Chambers et al. (2000), Koskiaho (2002), Quinton and Catt (2004), Ulén
and Jakobsson (2005), Kronvang et al. (2005), Knappen et al. (2008 and 2009), Deasy
et al. (2009), Stevens et al. (2009) and Silgram et al. (2010).
2.5.1 Soil and land management to prevent erosion
Various land management practices have been shown to minimise erosion risk on
susceptible soils: low erosion risk crops and cover crops, tillage timing and intensity,
and the use of buffer strips (Creamer et al., 2010). For example, intensively cultivated
soils amended with spent mushroom compost, a bi-product of the mushroom growing
industry in Ireland, exhibited improved structural stability as measured by an
aggregate stability (AgSt) test (Curtin and Mullen, 2007). The UK Department for
Environment, Food and Rural Affairs (Defra) highlights potatoes, winter cereals,
sugar beet, maize and grazed fodder crops as having the highest erosion risk based on
crop cover (Defra, 2005). To minimise erosion risk on susceptible soils, low risk
crops like oilseed rape (OSR), which establish a crop cover earlier, should be sown
(Chambers and Garwood, 2000). Furthermore, winter barley may be more beneficial
than spring barley, as it provides winter cover. However, wet weather trafficking may
offset benefits.
Minimum (or minimal) tillage, which involves shallow cultivation to a maximum
depth of 10 cm using a tine cultivator, helps conserve SOM, promotes AgSt and thus
reduces erosion (Quinton and Catt, 2004). In Ireland, minimum tillage normally
involves: (1) shallow cultivation using a tine cultivator or disc harrow to a depth of
75-100 mm immediately followed by rolling; (2) spraying with herbicide a few days
prior to sowing, following a stale seedbed period of a number of weeks (where
possible) to eliminate volunteers and established weeds; and (3) sowing with a
cultivator drill to a target depth of 40 mm (Forristal and Murphy, 2009). To date, the
55
Chapter 2
effectiveness of minimum tillage to reduce erosion has not been investigated in
Ireland. Research in the UK by Deasy et al. (2009) found that for 5 site-years, trialled
losses of SS and TP decreased by an average of 151 kg SS ha-1 and 0.3 kg TP ha-1
under minimum tillage, compared to traditional plough cultivation. Contour grass
strips have received some research attention and have been shown to reduce sediment
losses (Stevens et al., 2009) by reducing slope length and by acting as a barrier to
slow down overland flow. Deasy et al. (2010) found that although minimum tillage,
crop residue incorporation, contour cultivation and beetle banks (raised vegetative
barriers placed on the contour) all have potential to be cost effective mitigation
options for SS and TP losses, tramline management (disruption of the compacted
tramline surface to a depth of 60 mm with a tine) is one of the most promising
treatments for mitigating diffuse pollution losses as it was able to reduce sediment and
TP losses by 72-99% in four out of five site years trialled. As a management practice
to reduce P loss from tillage soils in Ireland, Carton et al. (2002) advised that attention
be paid to tramline compaction and that if soils become severely compacted,
corrective action, such as subsoiling, should be taken where appropriate.
The Nitrates Directive (91/676/EEC), as implemented in Ireland, sets out crop cover
requirements where arable land is ploughed between 1 st July and 30th November. The
regulations require that the owner/occupier take appropriate measures to provide for
emergence of green cover from a sown crop within 6 weeks of ploughing. In the UK,
as part of the cross compliance regime (Defra, 2006a), farmers are further required to
carry out a field erosion risk assessment as a means of reducing risk to acceptable
levels. The validity of this approach to erosion risk identification was verified by
Boardman et al. (2009). Conservation tillage in autumn may reduce losses of soil and
PP by improving soil structure. In Norway, ploughing and shallow cultivation of
sloping fields in spring, instead of ploughing in autumn, have been shown to reduce
particle transport by up to 89% on highly erodible soils (Ulén et al., 2010). Rational
land use policies such as the promotion of „set-aside‟ on erodible soils, use of grass
strips on erodible arable slopes, and buffer strips in riparian zones were identified as
mitigation options to reduce soil erosion by Fullen et al. (2003).
56
Chapter 2
There are some preventative measures in place to prevent land degradation processes
from arable agriculture (Table 2.1). In Ireland, farmers protect vulnerable tillage soils
by complying with „good agricultural and environmental condition‟ guidelines as a
condition for receipt of the area-based single farm payment under the EU crosscompliance regime (DAFF, 2005). Land that‟s been in continuous tillage for six years
or more must be tested for OM content as a requirement for the single payment
scheme. Soils having less than 3.4% SOM may require remedial action depending on
soil type. As the process of building up SOM is very slow, the remedial action to be
taken is set out over a 10-yr period. The remedial action will continue until such time
as the OM levels are shown to have recovered to greater than 3.4%, or a level deemed
acceptable for that soil type. Hackett et al. (2010) provide information on how various
management practices affect SOC dynamics in arable soils. Land application of
fertiliser and manures is now subject to „closed periods‟ that coincide with the most
frequent average occurrence of transport vectors. Farmers are also prohibited from
applying fertilisers in close proximity to a watercourse. „Buffer strips‟ of 1.5m and 5m
for mineral fertiliser and organic fertilisers, respectively, must be observed. The
effectiveness of these aspects of the regulations is currently being monitored in the
ACP (Schulte et al., 2010b).
Soil data currently available in Ireland exists in variable forms and is not fully
mapped at the target European scale (1:250,000). Digital soil mapping, combined with
conservative ground-truthing, is currently underway in the form of the Teagasc (Irish
Agriculture Food and Development Authority) and EPA funded Irish Soil Information
System (ISIS). This aims to complete the soil map of Ireland at a 1:250,000 scale
(Daly and Fealy, 2007), by the year 2014, by generating knowledge-based predictive
soil maps using digital terrain data, subsoil maps and other geo-spatial layers in an
advanced GIS technology platform. The models generated will be calibrated and
verified through an intensive two-year traditional field sampling campaign which will
provide hard soils data on 300 new reference profiles and over 3,000 auger points
across the country. In addition to the 1:250,000 soil map of Ireland will be an
associated digital soil information system which will be fully open and accessible to
all. The project is ground-breaking, as no other country has adopted such a
57
Chapter 2
complimentary approach of combining novel digital techniques with ground-truthing
using traditional soil survey methodologies at a National scale (Creamer, 2010).
In a review of strategies to improve soil conservation in Europe, Fullen et al. (2006)
identified several best management practices including: initiation of national soil
conservation services; and full mapping, monitoring and costing of erosion risk by
national soil survey organisations. If the SFD is eventually ratified, Ireland will be
required to identify areas where erosion has occurred in the past or is likely to occur
in the future. At that time, the soil information provided by the ISIS will be essential
in identifying these areas.
2.5.2 Developing
phosphorus
management
guidelines
for water quality
protection
The relationship between STP in tillage soils and DRP concentration in runoff water
needs to be adequately understood and quantified for local soils (Wright et al., 2006).
To date, in Ireland, no study has investigated the link between STP and P loss to water
from tillage soils. Guidelines presently used in Ireland are based on international
findings and agronomic nutrient advice. Determination of upper critical limits for P in
soil should consider both the STP necessary for economic crop production and the
STP necessary to avoid excessive P loss due to erosion, surface runoff and leaching.
This is essential for the development of P management guidelines for water quality
that will satisfy the requirements of the WFD. Relationships developed between
runoff P and STP have been used in Europe and the USA to establish threshold STP
levels above which the potential threat of eutrophication in surface waters is
unacceptable (Sibbesen and Sharpley, 1997; Sims et al., 2002).
In a study to evaluate Mehlich-3 P (M3-P) as an agri-environmental soil P test for the
Mid-Atlantic USA, Sims et al. (2002) concluded that agronomic soil tests, such as
M3-P, can be used to guide environmentally-based P recommendations, and that
higher risks are clearly associated with M3-P values that are in excess of
concentrations needed for economically optimum crop yields. As a result of the WFD,
there is increasing pressure in Europe and Ireland to develop P-based management
58
Chapter 2
practices that will reduce the risk of diffuse losses from agricultural land to surface
waters. Modelling of P for grassland undertaken by Schulte (2006b) showed that it
was possible to change the range of the target P index from 6 - 10 to 5.1 - 8 mg L-1 Pm
(Table 2.2), while still facilitating optimum productivity and herbage quality and
minimising the risk of diffuse P losses to water. Index 3 (5.1 - 8 mg L-1 for grassland)
in the new P-index system (Table 2.2) represents a target index that is both
agronomically and environmentally sustainable for all soils (Schulte, 2006b) in
Ireland. The target index for tillage crops (6 - 10 mg L-1) has not changed and it is
uncertain if similar work on tillage soils is necessary, as the risk of diffuse P loss from
them has not been quantified in Ireland.
The adoption of management measures in river basins requires the ability of river
basin managers to quantify the importance of different P pathways, identify and map P
risk areas with a certain spatial resolution, and estimate the effect of various
management measures for changes in P losses (Kronvang et al., 2005). Limited
resources and time will likely hinder the carrying out of a full P loss assessment
(incorporating site characteristics and nutrient management practices) on all
agricultural fields in a catchment. Therefore, in the interim, there is a need to identify
a STP level, sometimes referred to as an environmental threshold, above which the
improvement of P management practices should be a high priority.
2.5.3 Catchment-scale research
Research that will quantify the P and sediment losses associated with arable land
compared to agricultural grassland in Ireland is underway in the form of the ACP.
This will provide a scientific evaluation of the effectiveness of the Nitrates Directive
National Action Programme measures over time for the major farming and
environmental stakeholders in Ireland. The programme is designed to assess
effectiveness of measures well before improvements are expected to translate into
improved water quality of the final aquatic receptors, which in some cases may take
up to 20 yr (Schulte et al., 2010b). In the first stage, four catchments (2 arable and 2
grassland) were selected for studying from 1500 possible candidates using spatial
multi-criteria decision analysis (Fealy et al., 2010). Combined, the four catchments
59
Chapter 2
represent the range of intensive grassland and arable agriculture interests in Ireland
across a soil and physiographic gradient that defines potential risk of P and/or N
transfers (Fealy et al., 2010). A fifth catchment in a karst limestone region in the west
of Ireland is also being studied. The arable catchments, having between 30 and 50%
arable land use in each, are located in County Louth/Cavan on intermediately drained
soils and in County Wexford on well-drained soils, enabling measurement of storminduced diffuse transfers of P and losses of N to groundwater through leaching.
The ACP will focus on source, pathways and delivery of nutrients to waterways over
time. At the outlet of each catchment, the following parameters are being monitored:
TP, total dissolved P (TDP), total reactive P (TRP), DRP, total N, nitrate (NO3-N),
turbidity, electrical conductivity, temperature, and flow rate. Particular attention is
being paid to P hotspots (fields at soil P index 4) and linking these to P loads in
streams. This will facilitate the identification of areas that are vulnerable to P loss and
which will require measures to reduce losses. On-site bank-side nutrient analysers
(Jordan et al., 2007) will enable immediate analysis of nutrients susceptible to
transformation if left in sample bottles for long periods of time. Novel methodologies
will be used to quantify the amount of sediment leaving a catchment and relate this to
the source of the sediment and to specific areas and land uses. Measurements of
turbidity and electrical conductivity are also monitored to provide ancillary
information of sediment associated nutrient flux, pollution spikes and water flow
pathways (surface vs. sub-surface) (Wall et al., 2011). More detailed information on
the methodological design of the ACP and preliminary results can be found in Wall et
al. (2011).
Information on soil erosion and P loss across different land uses (e.g. tillage and
grassland) and its effect on water quality at catchment-scale will help Ireland meet the
requirements of the WFD. Detailed analysis of catchment characteristics, assessment
of risk to water bodies, further analysis of existing information and collection of new
data are all needed to support the implementation of the WFD (Irvine et al., 2005).
Given that there is still much to understand about the complex relationship between
the catchment and the movement of sediment and P, and the response of the aquatic
ecosystem to anthropogenic impacts, modelling that can elucidate key variables and
60
Chapter 2
predict responses is a valuable tool (Irvine et al., 2005). A review of available models
for modelling soil erosion and sediment and phosphorus delivery to surface waters at
the catchment scale is provided in Appendix B. This includes a comparison of model
predictions with measured export of P and SS from a number of Irish and
international catchments.
2.6 Future research direction in the quantification of phosphorus and sediment
loss from Irish tillage soils
2.6.1 Sediment provenance
Traditional techniques, aimed at identifying the source and the pathway of the
sediment, have included methods such as risk assessments, field observation and
mapping (Lao and Coote, 1993), landowner questionnaires (Krause et al., 2008),
remote sensing (Vrieling, 2006), use of erosion pins (Lawler et al., 1997), and
terrestrial photogrammetry (Barker et al., 1997).
Given the time and cost involved in establishing and operating plot experiments, and
that data available from them is limited, attention has been directed to the use of
environmental radionuclides for documenting erosion rates (Sepulveda et al., 2008).
By comparing the fallout radionuclide Caesium-137 (137Cs) inventory at a particular
sampling point with the reference inventory (the total
137
Cs activity per unit surface
area for a level, stable undisturbed site), the rates of soil erosion and deposition at that
point can be estimated. Measurements of 137Cs and unsupported 210Pb afford a means
of obtaining retrospective, medium-term (i.e. ca. 45 years for
years for unsupported
210
137
Cs and up to 100
Pb) estimates of both the magnitude and spatial distribution
of soil redistribution rates generated by sheet and rill erosion, by means of a single
site visit (Blake et al., 1999). Due to its long retention time on soil particles once
absorbed,
137
Cs (t1/2 = 30.1 yr) has the disadvantage of not being suitable for the
investigation of erosion resulting from individual events occurring over short periods,
and is unable to distinguish between tillage and water erosion. It can, however, be
used to estimate changes in soil erosion rates associated with changes in soil
management practices on cultivated land (Schuller et al., 2004). In contrast to
61
137
Cs,
Chapter 2
radioactive Beryllium-7 (7Be) is short-lived with a half-life of only 53 days and, as
such, is ideal for estimating short-term rates and patterns of soil redistribution relating
to individual events (tillage or water erosion) or short periods.
Because the radionuclides
137
Cs, 7Be, and
210
Pb have different distributions in the soil
profile, their measurement in eroded sediment, referred to as „sediment
fingerprinting‟, will determine what depth in the profile the soil was eroded from and,
hence, the depth and areal extent of sheet and rill erosion can be quantified as was
done in a study by Whiting et al. (2001). Sediment fingerprinting is a method to
allocate sediment nonpoint source pollutants in a watershed through the use of natural
tracer technology with a combination of field data collection, laboratory analyses of
sediments, and statistical modelling techniques (Davis and Fox, 2009). When
estimating sediment erosion rates, sediment fingerprinting has the added advantage
over plot studies of identifying both the source and fate of eroded sediment, which has
significant implications for the development of best management practices to address
soil erosion and sediment delivery to waterways.
In a study of one of Northern Ireland‟s prime salmon rivers (the River Bush) aimed at
quantifying fine sediment loads and tracing in-stream fine sediment sources using
sediment fingerprinting, Evans et al. (2006) were able to rank the four main agents
generating those sources, which were (in order of importance): drainage maintenance
work, bank erosion (caused by increasing flow and livestock poaching), ploughed
arable land, and forestry clearfell. Ploughed arable land was found to be responsible
for 36.6% of the suspended load and 7.5% of the bed load measured in the River Bush
over a 1-year period. Evans et al. (2006) commented that the most likely mechanisms
for transfer of topsoil to the river channel were after ploughing prior to planting and
harvesting of the crop. The best management practices recommended for the Bush
catchment to reduce sediment delivery from arable land by reducing bare ground
were: (1) critical area planting on land prone to long-term soil erosion; (2) planting at
appropriate times as assessed on the basis of storm forecasting; and (3) vehicle
movement limited across fields prone to soil erosion. Unfortunately, as Evans et al.
(2006) recognised, the 1-year period of monitoring in this project was too short to
provide a reliable picture of sediment dynamics in the Bush catchment. An EPA
62
Chapter 2
Strive funded project in conjunction with the Agri-food and Biosciences Institute
Northern Ireland (AFBINI) is currently underway that will use sediment
fingerprinting techniques to determine CSAs of sediment in two catchments in Co.
Down and Co. Louth. A similar project is underway as part of the ACP.
2.7 Summary
This chapter reviewed the current state of research and regulations on diffuse P and
sediment losses from tillage soils, and provided a review of the processes controlling
the mobilisation, transport and fate of P and sediment. An examination of the key
threats to soil quality associated with tillage soils, and the methods used to model and
quantify P loss and soil erosion was also detailed.
Modelling of water and tillage erosion rates in Ireland suggests that soil is being lost
at a rate greater than it can be replenished by natural soil formation. This has
significant implications for the sustainability of crop production. Furthermore, the
occurrence of erosion adjacent to waterways may result in the transfer of P and
sediment to them. Therefore, there is a need for laboratory- and field-scale research on
tillage soils in Ireland to determine the extent of erosion and associated P loss. Given
that a large proportion of P exported from agricultural catchments on an annual basis
is generated from a relatively small portion of the catchment and during only one or
two storm events, research to quantify P and sediment loss from Irish tillage soils
should utilise high intensity rainfall typical of summer storm events.
As P is often the limiting nutrient for eutrophication in surface waters (Jarvie et al.,
1998), river basin managers need to reduce P losses from agricultural land by adopting
plans for mitigation strategies. The ability to identify CSAs of P loss is essential if
mitigation measures are to be cost effective. The identification of an environmental
soil P threshold, above which surface runoff from tillage soils may have a negative
impact on water quality, will help Ireland meet the requirements of the WFD.
63
Chapter 3
Chapter 3 Determining phosphorus and sediment release
rates from five Irish tillage soils when subjected to simulated
rainfall and increasing overland flow rates
Overview
A controlled laboratory flume study, used to provide experimental data on the release
of P and sediment from 5 Irish tillage soils when subject to a simulated rainfall
intensity of 30 mm hr -1 and overland flow rates of 225 and 450 ml min-1, is presented
in this chapter. The impact of slope, time between storm events and overland flow rate
on P and sediment release from the study soils is also examined. The contents of this
chapter are, in part, published in the Journal of Environmental Quality (39:185-192,
2010).
3.1 Introduction
Phosphorus loss in surface runoff from soils is an important pathway in many agroenvironments (Sims et al., 2002; Wright et al., 2006). A survey of 1151 rivers in
Ireland from 2004 to 2006 (Clabby et al., 2008) estimated that the amount of pollution
attributed to agriculture was approximately one-third. The loss of fertile topsoil due to
soil erosion on agricultural land is a growing problem in Western Europe, and has
been identified as a threat to soil quality and the ability of soils to provide
environmental services (Boardman et al., 2009). Boardman and Poesen (2006)
estimated that arable agriculture accounts for approximately 70% of soil erosion in
Europe. It has numerous effects on soil, including thinning by removal of topsoil,
textural coarsening, decline of SOM and loss of nutrients (Guerra, 1994).
Soil erosion is also associated with P transfer by overland flow, especially from arable
land, where PP is the dominant P fraction exported (Doody et al., 2012). The
susceptibility of arable land to P losses by erosion and overland flow is largely a result
of land being left bare for periods of the year. An increase in winter cereal cropping
64
Chapter 3
has exacerbated this problem, because it combines the period of maximum rainfall
with long periods of bare soil (Leinweber et al., 2002). Furthermore, the disturbance
of soil structure by tillage operations, increases aggregate dispersion and the degree of
interaction between soil and runoff water, thereby enabling more dissolved P to be
mobilised from soils with high P status (Sharpley et al., 2001a).
In Ireland, the main pathway of P loss from soils is via overland flow (Kurz et al.,
2005; Tunney et al., 2007), which is greatest during storm events and is largely
inactive at other times (EPA, 2008). Saturation excess overland flow (characterised by
saturation of the soil over which it is moving) is the dominant type of overland flow
generated under Irish conditions (Daly et al., 2000; Diamond and Sills, 2001),
although research has shown that infiltration excess overland flow also occurs in
Ireland (Schulte et al., 2006a; Doody et al., 2010). Infiltration excess overland flow
occurs where the infiltration rate for a given soil profile is exceeded. The infiltration
and saturation excess-generating mechanisms are not mutually exclusive on a
watershed, nor even mutually exclusive at a point on a watershed (Smith and
Goodrich, 2005). For many soil profiles, saturation excess overland flow is a special
case of infiltration excess overland flow whereby infiltration is occurring, albeit at a
negligible rate, because of the low hydraulic conductivity of the underlying strata
(Nash et al., 2002). Where saturation excess and infiltration excess conditions
combine, the result is a complex pattern of P loss (McDowell et al., 2012).
Both saturation and infiltration excess overland flow can occur on tillage soils
provided the conditions necessary for it to occur are present. Tillage increases the
initial infiltration rate, loosens the topsoil, disrupts soil aggregates and compacts the
subsurface soil (Coles and Moore, 1998). This can result in a subsurface soil with
much lower hydraulic conductivity than the surface soil, and may lead to saturation of
the topsoil. The loose topsoil is then susceptible to erosion by saturation excess
overland flow. Infiltration excess overland flow is common with cultivated soils, or
where surface soil structure has degraded or consolidated to form a „seal‟, but can also
occur on unsealed soil surfaces, especially with high rates or amounts of rainfall
(Rose, 2004). Factors that increase the volume, velocity and turbulence of overland
flow, such as impaired infiltration, high intensity storms, run-on, reduced soil cover,
65
Chapter 3
cultivation and high slopes, increase detachment compared with dissolution (Nash et
al., 2002). Furthermore, steeper slopes increase the potential for runoff-dominated
erosion due to faster flow threads and lower surface area connectivity (Armstrong et
al., 2011).
The objective of this chapter was: (1) to quantify the amount of DRP, PP, TP and SS
released into overland flow from 5 tillage soils of varying Pm when subject to a
rainfall intensity of 30 mm hr -1 and overland flow rates of 225 and 450 ml min-1
applied in 3 successive events; and (2) to determine the impact, if any, of slope, time
between storm events and overland flow rate on P and sediment release from the
study soils.
3.2 The tillage soils, laboratory flume set-up, and analysis methods used in this
study
The USEPA National Phosphorus Research Project (NPRP, 2001) protocol uses soilpacked runoff boxes subjected to simulated rainfall to investigate the relationship
between STP and DRP in surface runoff. Runoff boxes containing homogenised soil
minimise significant variability in physical and chemical characteristics that may
occur in field plots. They also facilitate large numbers of replications not possible at
field-scale. Kleinman et al. (2004) found that regression coefficients of DRP in runoff
and M3-P were consistent between grassed field plots and soil-packed boxes, but
noted that runoff boxes may not be fully representative of field conditions. However,
they concluded that despite large differences in rainfall, hydrology, and erosion
between field plots and packed boxes, both can be used to produce comparable P
extraction coefficients for process-based models and P site assessment indices.
3.2.1 Soil collection and preparation
Fourteen tillage field sites, spread across Ireland, were investigated to find suitable
soils with wide ranging physical and chemical properties. After preliminary
characterisation, 6 soils were then selected based on soil type, STP, particle size
distribution (PSD), tillage history, and evidence of prior erosion problems. The soils
66
Chapter 3
selected were from: (1) Tullow, Co.Carlow; (2) Clonmel, Co. Tipperary; (3)
Letterkenny, Co. Donegal; (4) Bunclody, Co. Wexford; and (5) Fermoy Co. Cork; (6)
Duleek, Co. Meath (Figure 3.1). The Duleek soil was not used in the simulated
rainfall/overland flow study and is only considered in Chapter 5. The sites selected
were in tillage for a minimum of 15 years. The soils had a Pm index of 1 to 4 and
ranged from 2.8 to 17.5 mg L-1 (Table 3.1). The Pm values broadly reflected the P
fertiliser history of the 5 sites. The soils (Fermoy and Letterkenny) with low Pm (2.8
and 4.8 mg L-1, respectively) received P fertiliser below the recommended agronomic
levels over their rotation history, while those (Bunclody, Clonmel and Tullow) with
medium and very high Pm (7.1, 15.8 and 17.5 mg L-1, respectively) received P
fertiliser above agronomic levels. The Clonmel and Tullow soils had a history of
receiving above 40 kg P ha-1 yr-1 in excess of crop requirement.
A sample of the plough layer of each of the 5 tilled soils was collected, air-dried,
sieved (< 5 mm), and thoroughly mixed for use in a rainfall simulation/overland flow
study. This strategy was similar to that adopted by Miller et al. (2009). Other studies
by Sharpley (1980) and Fang et al. (2002) used soils sieved to less than 4 mm in
flume studies to determine the effect of storm interval on DRP in runoff and to
estimate runoff P losses, respectively. Subsamples of each soil were further sieved (<
2 mm) for physical and chemical characterisation (Section 3.2.3).
67
Chapter 3
Figure 3.1 Arable land use in Ireland (CORINE, 2006) and selected/rejected soil
sampling sites.
68
Chapter 3
Table 3.1 Chemical and physical properties of selected Irish tillage soils
Location
Soil Type pH
Pm1
CEC2
mg L-1 cmol kg-1
AgSt3 CaCO3
%
OM4
Sand
_______________________
Silt
Clay
g kg-1_____________________
M3-P5 Pcacl26 WEP7
____________________________
Pox8
Alox9
Feox10
mg kg-1_____________________
Psatox11
%
Tullow, Co. Carlow
GBP12
6.9
17.5
13.4
96.4
5
49
579
267
154
96.3
3.0
11.5
566
1033
3482
36.2
Clonmel, Co. Tipperary
GBP
6.7
15.8
11.2
90.0
5
42
528
306
167
89.4
2.1
6.6
457
903
3867
28.8
Bunclody, Co. Wexford
BP13
7.7
7.1
13.9
92.9
26
71
410
387
203
58.7
1.0
3.5
414
2560
4755
14.8
Letterkenny, Co. Donegal
BP
6.5
4.8
13.7
98.5
17
55
491
395
114
52.1
1.3
2.8
592
1700
5468
23.7
Fermoy, Co. Cork
ABE14
6.4
2.8
13.1
97.8
4
51
569
286
145
29.1
1.4
2.3
273
1227
3886
15.4
1
Pm, P determined by Morgan‟s extraction; 2CEC, cation exchange capacity; 3AgSt, aggregate stability; 4OM, organic matter by loss on ignition; 5M3-P, Mehlich-3
extractable P; 6Pcacl2, calcium chloride extractable P; 7WEP, water extractable P; 8Pox, acid ammonium oxalate extractable P; 9Alox, acid ammonium oxalate extractable Al;
10
Feox, acid ammonium oxalate extractable Fe;
11
Psatox, soil P saturation as determined by acid ammonium oxalate extraction;
14
podzolic; ABE, acid brown earth.
69
12
GBP, grey brown podzolic;
13
BP, brown
Chapter 3
3.2.2 Simulated rainfall experiment
A laboratory scale runoff experiment was designed to compare the nutrient and
sediment releases from the 5 study soils inclined at slopes of 10 and 15 degrees, when
subjected to high intensity (30 mm hr -1) simulated rainfall. In order to minimise the
effects of soil variables on P release to runoff and give better control over
hydrological and soil surface conditions, laboratory rainfall simulations were chosen
to compare the nutrient and sediment releases from the 5 soils. As the study was
focused on understanding process rather than soil management, this was considered to
be a reasonable approach. The layout of the flume set-up is illustrated in Figure 3.2
(the overland flow reservoir shown is only used in the experiments described in
Section 3.2.3). This experiment used two laboratory runoff boxes („flumes‟), 200-cmlong by 22.5-cm-wide by 5-cm-deep with side walls 2.5 cm higher than the soil
surface, and 5-mm diameter drainage holes, drilled in triplicate and located at 300mm-centres, in the base. Cheese cloth was placed at the base of each flume before
packing to prevent soil loss through drainage holes. Runoff water was collected at the
lower end of the sloped flume by a U-shaped aluminium trough equipped with a
canopy to prevent rainfall water entering the runoff collection containers.
Figure 3.2 Laboratory flume set-up for rainfall simulator/overland flow
experiment.
70
Chapter 3
A rotating disc, variable-intensity rainfall simulator (after Williams et al., 1997), was
constructed and calibrated for use in the rainfall simulation study. The rainfall
simulator (Figure 3.3) consisted of a motorised rotating disk module for regulating the
rainfall intensity and a single 1/4HH-SS14SQW nozzle (Spraying Systems Co.,
Wheaton, IL), which, at a pressure of 100 kPa, creates a distribution of drop sizes
approximating natural rainfall (Bubenzer et al., 1985), with similar raindrop impact
energy; i.e. 260 kJ mm-1 ha-1 (nozzle) and 240 kJ mm-1 ha-1 (natural). The rainfall
simulator was attached to a 2.5 m by 2.5 m by 4.5-m-high metal frame, and calibrated
prior to each experiment, to ensure that the rainfall intensity had not changed since the
last experimental run. During calibration, the two flumes were placed under the
rainfall simulator equidistant on either side of the nozzle so that each received
approximately 30 mm hr -1 of rainfall. The total mass of water in each flume was used
to determine how much rain had fallen on each flume (area receiving rainfall = 0.45
m2). The rainfall distribution in each flume was calculated by collecting rainfall for
three 30-min periods in 18 identical cylindrical containers spread across the area of
the flume. The mass of water in each container was determined and converted to a
depth value (mm h-1) to obtain the intensity and distribution of the rainfall. The
Christianson coefficient (Cu) of application uniformity (Christianson, 1942): Cu = (1average deviation from mean / mean depth of applied water) × 100, was used to
evaluate depth distribution. A uniform depth distribution generates a Cu = 100. A Cu
> 85% was achieved in all calibrations.
Figure 3.3 Rainfall Simulator (isometric drawing and photo of underside)
Soils were packed to achieve an approximate bulk density of 1.3 - 1.5 g cm-3 (Figure
3.4). The packed soil was then saturated using the simulator, and left to drain for 24 hr
71
Chapter 3
before the experiment commenced. A furnace filter was placed on the soil surface to
protect the soil from raindrop impact during saturation. The furnace filter was
removed before the start of the first rainfall event. All soils were approximately at
field capacity before the first rainfall event (field capacity was determined to have
been achieved once drainage from the base of the runoff box had ceased).
Figure 3.4 Soil in laboratory flume before and after rainfall simulation.
The return period for a 30 mm hr -1 rainfall event ranged from 30-100 years across the
selected soil locations. This was based on the depth-duration-frequency model of
Fitzgerald (2007). However, these return periods may be overestimated, given the
ongoing changes in precipitation and storm frequency due to climate change. The 10yr moving average for Ireland shows that rainfall amounts increased from 800 mm in
the 1890s to 1100 mm in the 1990s (McElwain and Sweeney, 2006). Furthermore,
Sweeney et al. (2008) modelled the effect of increased global emissions on the
hydrology of nine Irish river catchments and concluded that the magnitude and
frequency of flood events will increase due to climate change, with the greatest
increases associated with floods of a higher return period. By the 2020s, three of the
catchments modelled, in which there is a significant area under tillage (Blackwater,
Suir and Barrow), showed an increase in the frequency of the 50-yr flood, with the
same flood expected every 8.4 - 12.6 yr under a medium-to-low emission scenario and
every 3.8 - 7.4 yr under a high-to-medium emission scenario. In effect, this means that
high intensity rainfalls, such as the 30 mm hr -1 intensity investigated in this study,
could potentially occur every 5 - 10 yr in the period 2020 - 2030. These projected
decreases in the time period between extreme floods are likely to result in greater
levels of erosion in tillage areas in Ireland. Large rainfall return periods are not
72
Chapter 3
uncommon when investigating the effect of high-intensity storm events. A study by
Vadas et al. (2005) used simulated rainfall intensities that represented storm return
periods ranging from 5 - 50 yr.
The source water for the rainfall simulations was potable tap water with DRP, NO3-N,
and ammonium-N (NH4-N) concentrations of < 0.005, 0.036, and 0.038 mg L-1,
respectively. The tap water had an electrical conductivity = 0.421 dS m-1, measured
using a conductivity meter, and a calcium cation (Ca2+), magnesium cation (Mg2+),
and sodium cation (Na2+) concentration of 3.11, 2.24, and 22.55 mg L-1, respectively,
measured by atomic absorption spectrophotometry. The Sodium Adsorption Ratio
(SAR = Na/[(Ca + Mg)/2]1/2, where all concentrations are expressed in meq/liter) of
the tap water was 2.38. Annual mean concentrations (in mg L -1) of Ca2+, Mg2+, and
Na2+ in rainwater were 0.85, 0.93, and 6.78, respectively, between 1992 and 1994 for
the island of Ireland (Jordan, 1997). Over the same period, the SAR was 1.21 and the
electrical conductivity ranged from 0.029 - 0.176 dS m-1. More divalent cations
present in tap water than in natural rainfall may encourage larger aggregates in surface
runoff (Aase et al, 2001). If this occurs, P losses in runoff using tap water could
potentially be lower than for natural rainfall, since finer soil particles and aggregates
have higher P concentrations than larger soil particles. The higher electrical
conductivity of the tap water used here, compared to natural rainfall in Ireland, may
result in lower P desorption from the study soils. However, Aase et al. (2001) found
that average DRP concentrations in runoff from a calcareous soil using two different
water sources, with electrical conductivities of 0.02 and 0.4 ds m-1, were equivalent to
eachother. Tap water has been used previously by Penn et al. (2006) when estimating
dissolved P concentrations in runoff from three physiographic regions of Virginia and
by McDowell and Sharpley (2002) when investigating P transport in overland flow.
Most recently, tap water (0.005 mg P L -1) was used by Wang et al. (2010) when
estimating DRP concentration in surface runoff from major Ontario soils.
Each rainfall simulation comprised 3 successive 1-hr rainfall events at time zero
(Rainfall 1), 1 hr (Rainfall 2) and 24 hr (Rainfall 3) to determine the effect of storm
interval on surface runoff. Previously, Sharpley (1980) used storm intervals of 5 and
30 min and 1-day when comparing the effects of short intervals and 1-day intervals on
73
Chapter 3
the concentration of soluble P in runoff. During the rainfall simulation, 6 drainage
holes remained open to better replicate field conditions. This limited drainage
scenario is designed to replicate a tillage field where subsurface compaction has
impeded drainage and resulted in the topsoil becoming saturated. It is also
representative of areas at the base of slopes or along rivers where the water table is
near the surface. As the risk of surface runoff increases with slope, each soil was
examined at 2 slopes, 10 and 15 degrees, to investigate the effect of slope on nutrient
and sediment losses in the runoff. Surface runoff samples were collected when runoff
began: once every 2.5 min for the first 20 min and in each subsequent 5-min interval
to evaluate changes in runoff volume, and nutrient and sediment concentration over
time.
3.2.3 Overland flow experiment
In this laboratory-scale overland flow experiment, soils were prepared and tested in
exactly the same way as in the simulated rainfall experiment described in Section
3.2.2, with the exception of the introduction of two distinct overland flow rates via an
overflow reservoir (Figure 3.2) located at the top of the runoff box. Two separate
experiments were conducted, in which an overland flow of either 225 or 450 ml min-1
was added at the top of the runoff box (when inclined at a 10 degree slope) in the
presence of rainfall (30 mm hr
-1
), in order to investigate the effect of increasing
overland flow rates on nutrient and sediment release from the study soils. The
overland flow was generated by pumping water into a reservoir at the top of each
flume using a Cole-Parmer Masterflex® L/STM peristaltic pump, which was calibrated
prior to each experimental run. The water was allowed to flow over a metal plate that
was level with the soil surface. It was envisaged that this approach would best
replicate sheet flow arriving at the top of the flume. The two overland flow rates
applied represent possible worst case scenarios in fields, where the soil has become
saturated due to high intensity rainfall. Increases in high intensity storm events due to
climate change are likely to result in tillage soils being subject to higher volumes of
overland flow. The approximate surface runoff rates at the end of the runoff boxes for
the 3 conditions of rainfall only (simulated rainfall experiment), rainfall and overland
74
Chapter 3
flow at 225 ml min-1, and rainfall and overland flow at 450 ml min -1, were 200, 425
and 650 ml min-1, respectively.
Each simulation comprised 3 successive 1-hr rainfall/overland flow events at time
zero, 1 hr (this event began 1 hr after the first event finished) and 24 hr (this event
began 24 hr after the 2nd event finished) to determine the effect of storm interval on
surface runoff. At each time interval, each soil was subjected to either rainfall and
overland flow at 225 ml min-1, or rainfall and overland flow at 450 ml min-1, to
investigate the effect of increasing overland flow rate on nutrient and sediment losses
in the runoff. Similarly, Hairsine (1988) introduced clear water at the top of the flume
in the presence of rainfall when investigating erosion of a cohesive soil in a flume
testing facility. Shallow depths of overland flow permit raindrop impact to have a
significant influence on the removal of sediment from a soil bed and its subsequent
displacement downslope (Hairsine, 1988). Following the commencement of surface
runoff, water samples were collected as described in Section 3.2.2.
3.2.4 Soil analysis
The soil characteristics measured in each of the 5 test soils were: (1) pH (1:1
soil/solution ratio); (2) PSD by sieve and pipette analysis; (3) P m was determined by
adding 8 ml of dried and sieved (< 2mm) soil to 40 ml of Morgan‟s Reagent (Morgan,
1941) (1480 ml of 40% sodium hydroxide and 1444 ml of glacial acetic acid to 20 L
distilled water at pH 4.8) and shaking for 30 min on a Brunswick Gyratory shaker.
The filtered extracts were analysed colorimetrically for P; (4) SOM by loss on ignition
at 550°C (Byrne, 1979); (5) ammonium oxalate-oxalic acid extractable P (Pox),
aluminium (Alox), iron (Feox) measured by inductively coupled plasma-atomic
emission spectroscopy. Soil P saturation (Psat ox) was calculated as Pox (mmol kg-1),
divided by α[Alox + Feox] (α = 0.5 for non calcareous sandy soils), and multiplied by
100 (Schoumans, 2009); (6) WEP was measured by shaking 0.5 g of soil in 40 ml of
distilled water for 1 hr, filtering (0.45 µm) the supernatant water and determining P
colorimetrically; (7) M3-P (Mehlich, 1984) (8) cation exchange capacity (CEC;
Bascomb (1964)); (9) calcium carbonate (CaCO3) was determined by the volumetric
method (ISO, 1995; ISO 10693) using a Scheibler apparatus; (10) calcium chloride
75
Chapter 3
extractable P ( Pcacl2) by extraction with 0.01 M CaCl2; (11) AgSt was determined
using the wet sieving apparatus (Eijkelkamp Agrisearch Equipment, The
Netherlands). Selected soil chemical and physical properties are shown in Table 3.1.
To ensure homogeneity of the individual soils, 3 subsamples of each soil were tested
for Pm to determine the coefficient of variation (standard deviation divided by mean
Pm concentration) for each soil. The coefficient of variation was < 0.05 for the 5 soils.
3.2.5 Runoff analysis
Immediately after collection, runoff water samples were filtered (0.45 µm) and
analysed colorimetrically for DRP using a nutrient analyser (Konelab 20, Thermo
Clinical Labsystems, Finland). Total reactive phosphorus was determined on
unfiltered samples, which were immediately frozen after collection. During
defrosting, SS settled out facilitating the extraction of 2 ml of clear water by syringe
from just below the surface of the water sample which was then analysed
colorimetrically for TRP using the nutrient analyser (data not shown). Runoff water
samples were frozen at -20 °C until TP was conducted. Total phosphorus was
determined for every second runoff sample after acid persulphate digestion. Total
phosphorus comprises both PP and TDP. As in other studies by Randall et al. (2005)
and Udeigwe and Wang (2007), PP was calculated by subtracting DRP from TP. As
tests indicated that TDP was similar to DRP in the runoff water, and two orders of
magnitude smaller than TP, this simplification was deemed appropriate. All digested
samples were analysed colorimetrically for P using the nutrient analyser. Suspended
sediment concentrations were determined for all samples by vacuum filtration of 50
ml of well-mixed runoff water through Whatman GF/C (pore size 1.2 µm) filter
paper. All samples were tested in accordance with the Standard Methods (APHA,
2005) by the candidate at the Department of Civil Engineering, NUI, Galway.
3.2.6 Data analysis
Flow-weighted mean concentration (FWMC) values for nutrients and sediment in
runoff from the flumes were determined by dividing the total mass load for the 1-hr
runoff event by the total flow volume for the same period.
76
Chapter 3
3.2.7 Statistical methods
In the case of the rainfall only experiment, a generalized linear mixed model (GLMM)
was fitted to each surface runoff response to test whether the effect of soil type
depends on slope and rainfall event. A random effect with a first-order autoregressive
variance-covariance structure was fitted to account for non-independence of
successive rainfall events (The GLIMMIX Procedure, SAS V9.1). A log link function
was required for all surface runoff responses to satisfy the assumption of normality of
residuals. When investigating P transport in surface runoff from packed soil boxes,
Kleinman et al. (2004) also logarithmically transformed P concentrations as did Little
et al. (2005), when investigating P losses from a plot experiment.
The inclusion of soil type as a factor in the GLMM allows quantification of the
variation due to soil, so that we can then determine how much of this variation is
accounted for by the different soil characteristic parameters. For DRP and TRP, the
effect of soil type differed depending on slope and rainfall event. A stepwise
regression selection procedure was subsequently conducted for each slope-rainfall
combination to determine which characteristics were important in explaining variation
in DRP and TRP. For SS, TP and PP, the stepwise procedure was performed
separately for each slope, as the effects of soil type differed depending only on slope.
In the case of the simulated rainfall/overland flow experiments, a linear mixed model
(LMM) was fitted to each surface runoff response to test whether the impact of soil
type on DRP, TP, PP and SS concentrations in surface runoff was affected by flow
rate and rainfall event. A random effect with either a compound symmetry or a firstorder autoregressive variance-covariance structure was fitted to account for nonindependence of successive rainfall events (The GLIMMIX and MIXED Procedures,
SAS, 2004). A log transformation was required for all surface runoff responses to
satisfy the assumption of normality of residuals. The analysis was conducted as a
factorial combination of overland flow rate and rainfall event, with soil type as a
blocking factor. The general classification by soil type allowed testing of the effects
of the overland flow rate and event, but a number of covariates were recorded as a
characterisation of the soil type. A series of models were fitted by removing the soil
77
Chapter 3
type category and substituting mostly continuous variables in an attempt to improve
understanding of the processes involved. Covariates were fitted initially in a
hypothesis-based set of tests and subsequently best-fit models were obtained using a
combination of hypotheses and stepwise selection of regressor variables. All covariate
testing was carried out on an analysis model incorporating the experimental factors so
that the full data set could be used without any bias due to the structure of the
treatments.
3.3 Experimental results from simulated rainfall/overland flow experiments
The results from the simulated rainfall/overland flow events on flumes packed with
Irish tillage soils are presented here. These data provided information on amounts of P
and sediment lost in surface runoff from the study soils.
3.3.1 Characteristic properties of the study soils
The selected study soils covered a Pm index of 1 - 4 and ranged from 2.8 to 17.5 mg
Pm L-1. Mehlich-3 P ranged from 29.1 to 96.28 mg kg -1 and was well correlated with
Pm (r= 0.98), WEP (r = 0.89), Psatox (r = 0.86), and Pcacl2 (r = 0.80). Water extractable
P ranged from 2.3 to 11.5 mg kg -1 and was well correlated with Pm (r = 0.92), Psatox (r
= 0.9) and Pcacl2 (r = 0.95). Calcium chloride extractable P ranged from 1.0 to 3.0 mg
kg-1 and was well correlated with Pm (r = 0.87) and Psatox (r = 0.93). Soil P saturation
ranged from 14.8 to 36.2% and was well correlated with Pm (r = 0.87). If an
agronomic soil P test is to be used for environmental purposes, it is important that it
be well correlated with the forms of soil P most susceptible to losses in surface runoff
and with Psatox (Sims et al., 2002). Soil pH, OM, and AgSt ranged from 6.4 to 7.7,
41.7 to 70.5 g kg-1, and 90 to 98.5%, respectively.
Particle size analysis showed that sand was the dominant size fraction across the 5
soils, and ranged from 410 to 579 g kg-1. The silt fraction ranged from 267 to 395 g
kg-1, and the clay fraction ranged from 114 to 203 g kg -1. The basic soil textural class
ranged from loam to sandy loam, with sandy loam dominating. This was
representative of tillage which predominates in the east and south of Ireland, where
78
Chapter 3
soils are highly suited to tillage, and generally have a light-to-medium texture, friable
consistence and free drainage (Gardiner and Radford, 1980). Heavier textured soils
are less suitable for tillage in Ireland as they generally occur at higher elevations
leading to slope problems, have a weak structure, and are imperfectly drained
(Gardiner and Radford, 1980). Soil CEC ranged from 11.2 to 13.9 cmol kg-1.
3.3.2 Suspended sediment and phosphorus concentrations in runoff from
simulated rainfall events
Generally, the highest SS and P concentrations occurred within 15 min of the
commencement of surface runoff from the flumes and had reached steady-state 30
min after the commencement of the first rainfall event (Figure 3.5 and Figure 3.6).
The high DRP concentration in runoff at the start of a storm event may partly be
explained by dilution as a function of runoff rates, which did not reach equilibrium
until 5 min into a rainfall event. In addition, given that soils were pre-wet for up to 24
hr before rainfall commenced, a larger portion of the readily soluble and more slowly
soluble P forms may have reached solution, thus elevating P levels in the soil water.
As the rainfall event took place, the older pre-wet water became diluted by clean
simulation water, resulting in a reduction in runoff DRP concentration. During the
remainder of the rainfall event, the DRP measured in runoff was controlled by the
pool of P that was freely available for desorption, and was rapidly desorbed and
transferred into overland flow. Soils with high concentrations of extractable soil P had
the highest concentration of DRP in surface runoff (Figure 3.5). This is a result of
there being more freely available P in the soil solution at higher extractable soil P
levels.
The FWMC of DRP in surface runoff was highest from the Tullow and Clonmel soils,
which peaked at 0.09 and 0.042 mg L-1, respectively, during the first rainfall event
and may have negatively affected water quality. These FWMC equated to DRP loads
in surface runoff from the Tullow and Clonmel soils of 0.893 mg (or 19.85 g ha -1) and
0.338 mg (or 7.52 g ha-1), respectively, during the 1 hr rainfall events. In contrast, the
79
Chapter 3
Rainfall 1 - 1st Rainfall
Rainfall 2 - 1 hr after Rainfall 1
Rainfall 3 - 24 hr after Rainfall 2
Dissolved Reactive P (mg L-1)
0.16
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0
5000
SS (mg L-1)
4000
3000
2000
1000
0
Total P (mg L-1)
8
6
4
2
0
Particulate P (mg L-1)
8
6
4
2
0
0
10
20
30
40
50
60 0
10
20
30
40
50
60 0
10
20
30
40
Time from start of rainfall (min)
Figure 3.5 Phosphorus and sediment losses over time from tillage soils inclined at
a 10 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody (○),
Fermoy ( ).
80
50
60
Chapter 3
Dissolved Reactive P (mg L-1)
Rainfall 1 - 1st Rainfall
Rainfall 2 - 1 hr after Rainfall 1
Rainfall 3 - 24 hr after Rainfall 2
0.16
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0
Clonmel
Letterkenny
Fermoy
Bunclody Clonmel
Tullow
Letterkenny
Fermoy
Bunclody Clonmel
Tullow
Letterkenny
Fermoy
Bunclody
Tullow
5000
SS (mg L-1)
4000
3000
2000
1000
0
Clonmel
Total P (mg L-1)
8Fermoy
Letterkenny
Bunclody
Clonmel
Tullow
axis 2 Fermoy
Letterkenny
Bunclody
Tullow
axis 2
Clonmel
Fermoy
Letterkenny
Bunclody
Tullow
axis 2
6
4
2
0
Clonmel
Letterkenny
Fermoy
Bunclody Clonmel
Tullow
Letterkenny
Fermoy
Clonmel
Bunclody
Tullow
50
60
0
Letterkenny
Fermoy
10
20
30
40
Bunclody Clonmel
Tullow
Letterkenny
Bunclody
Tullow
Clonmel
50
60
0
Letterkenny
Fermoy
10
20
30
40
Fermoy
Bunclody
Tullow
Bunclody
50
60Tullow
Particulate P (mg L-1)
8
6
4
2
0
Clonmel
0
Letterkenny
Fermoy
10
20
30
40
Time from start of rainfall (min)
Figure 3.6 Phosphorus and sediment losses from selected tillage soils inclined at a
15 degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody (○),
Fermoy ( ).
81
Chapter 3
threat to surface water quality posed by the Fermoy soil, as a result of its potential to
release DRP, was much lower, as evidenced by a peak FWMC of DRP of only 0.014
mg L-1, which occurred at a 15 degree slope, during the 1st rainfall event. This equates
to a DRP load of 0.127 mg (or 2.82 g ha-1). As such, the Tullow soil (Soil P Index 4),
when subjected to simulated rainfall, has the potential to transfer 7 times as much
DRP into surface runoff as the Fermoy soil (Soil P Index 1) and therefore poses a
much greater risk to surface water quality. Whether this risk ultimately translates into
impairment of a surface water body depends primarily on connectivity between the
location where surface runoff occurred and a water body that is sensitive to pollution
from P inputs. If connectivity between the P source and the pollution sensitive water
body can be shown, then the area is termed a CSA. As DRP is readily available for
uptake by aquatic plants, the likelihood of it resulting in eutrophication upon reaching
a water body is far greater than that of an equivalent mass of PP, the availability of
which can be as low as 10%. In general, the peak value of DRP observed during the
first rainfall event on each soil reduced in subsequent rainfall events. Similar trends
were noticed in the SS, TP and PP concentrations of runoff from the flumes.
The potential for particulate losses in surface runoff was found to be high for the 5
soils, with the highest coming from the Clonmel soil when inclined at a slope of 15
degrees, where peak SS and PP concentrations of 4263 mg L -1 and 5.99 mg L-1,
respectively, were measured during the first rainfall event (Figure 3.6). The SS and PP
loads measured in surface runoff from the same 1 hr event were 8.67 g (or 192.6 kg
ha-1) and 12.8 mg (or 284 g ha-1), respectively. The environmental risk posed by the
Fermoy soil as a result of its low susceptibility to particulate losses, was lower than
that of the other soils as evidenced by PP loads measured in surface runoff at 10 and
15 degree slopes of 1.58 g (or 35.1 kg ha -1) and 3.92g (or 87.1 kg ha-1), respectively,
during the first rainfall event. Particulate P contributed 84 to 99% of total runoff P.
This is in close agreement with Fang et al. (2002), who reported that PP contributed
from 59 to 98% of total runoff P for unvegetated packed boxes. The greater
contribution of PP to total runoff P in this study is probably a result of the steeper
slopes investigated. Similarly, Sharpley et al. (1994) reported that PP contributed 75
to 95% of total runoff P from conventionally tilled land. As PP is generally bound to
the minerals (particularly Fe, Al, and Ca) and organic compounds contained in soil, it
82
Chapter 3
constitutes a long-term P reserve of low bioavailability. The availability of PP to
plants and algae is variable, ranging from 10 to 90% of the TP, but it can represent a
long-term source of P for algae and plant uptake from surface water bodies, in
particular lakes. Reducing dissolved P loss is far more difficult than reducing P loss
associated with erosion (McDowell and Sharpley, 2001) and control measures are
mainly limited to preventing soil P accumulation to environmentally sensitive levels
(Sibbesen and Sharpley, 1997). Albeit nutrients lost in surface runoff from packed
boxes are broadly consistent with those lost from field plots, the exposed bare soils of
packed boxes are vulnerable to erosion, resulting in greater PP concentrations in
runoff (Kleinman et al., 2004). The PP and SS losses measured in surface runoff
during this study may represent a worst case scenario because of the steep slope
(typical of sites where a P loss risk assessment is necessary), high rainfall intensity
(typical of storm events), and bare soil with reduced AgSt compared to in situ soil. It
is unlikely that SS and PP edge of field losses from these soils in situ would be so
high given the variable surface slope and proximity to surface waters. Recognised
management practices, namely, riparian buffer strips, conservation tillage, and
contour ploughing, where used, are also effective in controlling PP loss from
agricultural fields.
3.3.3 Suspended sediment and phosphorus concentrations in runoff from
simulated rainfall/overland flow events
Generally, the highest SS and P concentrations across the 5 soils, for an overland flow
rate of 225 ml min-1, occurred within 15 min of the commencement of the zero hr, 1
hr and 24 hr events, and reached steady-state no later than 30 min after
commencement of runoff (Figure 3.7) with the exception of the Tullow soil, which
did not achieve steady-state for particulate losses. Introducing overland flow rates of
225 and 450 ml min-1 at the top of the flume had the effect of increasing the runoff
rate at the end of the flume from approximately 200 ml min -1 (for rainfall only) to 425
and 650 ml min-1, respectively.
For the higher flow rate of 450 ml min -1, nutrient and sediment concentrations only
achieved steady-state for some soils (Figure 3.8). An increase in overland flow rate (0
83
Chapter 3
up to 225 ml min-1 and up to 450 ml min-1) resulted in an increase in concentrations of
SS, PP and TP in surface runoff across all soils (p < 0.05). These increases in
concentrations were generally not proportional to the increase in runoff rate measured
at the end of the flume. This is to be expected given the vulnerable nature of the soils,
after being sieved and then packed into flumes. As might be expected, there was a
strong relationship (r = 0.92, p = 0.0001) between SS and PP concentrations measured
in surface runoff across the 5 soils. In general, soils that experienced higher SS losses
at 450 ml min-1 had higher levels of variability in nutrient and SS concentrations
between replicate samples. An increase in extractable soil P resulted in an increase in
concentrations of DRP in surface runoff (p < 0.05) across all soils. The FWMCs of
DRP in surface runoff (overland flow rate = 225 ml min-1) from the Tullow and
Clonmel soils peaked at 0.07 and 0.028 mg L-1, respectively, during the time zero
events. Therefore, runoff from these soils has the potential to cause eutrophication on
reaching a sensitive water body. These FWMCs equated to DRP loads in surface
runoff from the Tullow and Clonmel soils of 1.79 mg (or 39.8 g ha-1) and 0.732 mg
(or 16.3 g ha-1), respectively, during the 1-hr rainfall events.
The potential for particulate losses in surface runoff was very high for the 5 soils, with
the highest losses coming from the Tullow soil (overland flow rate = 450 ml min-1),
where FWMCs for SS and PP of 3.86 g L-1 and 4.25 mg L-1, respectively, were
measured (Figure 4.3) for the time zero event. These high concentrations were to be
expected given the worst case scenario being investigated. The impeded drainage of
the flume, high rainfall intensity, overland flow run-on, soil similar to that of a finely
harrowed field, and 10 degree slope investigated are all conducive to increased rates
of soil detachment. While FWMCs of PP measured in surface runoff across the study
soils were far greater than FWMCs of DRP, it must be borne in mind that PP is
attached to sediment which may settle out of suspension if the runoff transporting it,
encounters less steep slopes than those in which it initially entrained the sediment.
Similarly, if runoff encounters vegetation capable of reducing its velocity, the larger
sediment will settle out, thereby reducing the concentration of PP being transported
towards the aquatic receptor. There is greater potential for PP concentrations
measured in surface runoff from the study soils to be reduced, as runoff makes its way
toward the aquatic receptor; this must be considered when comparing the risk posed
84
Chapter 3
Dissolved Reactive P (mg L-1)
Time zero
1-hr interval
24-hr interval
0.16
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0
15
SS (g L-1)
12
9
6
3
0
Clonmel
Letterkenny
Tullow
Fermoy
Bunclody
Total P (mg L-1)
16
12
8
4
0
Particulate P (mg L-1)
16
12
8
4
0
0
10
20
30
40
50
60 0
10
20
30
40
50
60 0
10
20
30
40
Time from start of overland flow (min)
Figure 3.7 Phosphorus and sediment concentrations in runoff water from tillage
soils subjected to rainfall and overland flow (225 ml min-1) when inclined at a 10
degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody (○), Fermoy (
).
85
50
60
axis 2
Chapter 3
Dissolved Reactive P (mg L-1)
Time zero
1-hr interval
24-hr interval
0.16
0.14
0.12
0.10
0.08
0.06
0.04
0.02
0
20
SS (g L-1)
16
12
8
4
0
Total P (mg L-1)
20
15
10
5
0
Clonmel
Letterkenny
Fermoy
Bunclody
Tullow
Particulate P (mg L-1)
20
15
10
5
0
0
10
20
30
40
50
60 0
10
20
30
40
50
60 0
10
20
30
40
Time from start of overland flow (min)
Figure 3.8 Phosphorus and sediment concentrations in runoff water from tillage
soils subjected to rainfall and overland flow (450 ml min-1) when inclined at a 10
degree slope. Clonmel (▲), Tullow (□), Letterkenny (♦), Bunclody (○), Fermoy (
).
86
50
60
Chapter 3
by PP and DRP. Furthermore, PP is of limited availability for uptake by aquatic
plants, whereas DRP is generally considered to be 100% available.
While the effect of overland flow rate on DRP concentrations measured in overland
flow across the 5 soils tested was variable, its effect on mass losses was quite clear
(Figure 3.9a). As the overland flow rate increased, there was an almost proportional
increase in DRP lost in surface runoff, indicating that any dilution effect of higher
flow rates was minimal. This is discussed further in the Section 4.3.3. Mass losses of
DRP were highest for the Tullow soil, where 2.9 mg (or 64.4 g ha-1) was released for
time zero at an overland flow rate of 450 ml min-1 compared to 0.893 mg (or 19.85 g
ha-1) when subjected to 30 mm hr -1 rainfall only.
Increases in mass losses of SS, PP and TP (data not shown because it is
indistinguishable from PP) due to increased overland flow rates were more
pronounced (Figure 3.9b-c) than those for DRP, with the exception of the Letterkenny
soil, which was more resistant to degradation. For example, mass losses of SS from
the Tullow soil at time zero were 2.67 (59.33 kg ha-1), 99.3 (2206 kg ha-1) and 375.8
(8352 kg ha-1) g when subjected to rainfall only, 225 ml min-1, and 450 ml min-1,
respectively. While FWMCs of SS from the Tullow soil at time zero were 0.269, 3.86
and 9.63 g L-1 when subjected to rainfall only, 225 ml min-1, and 450 ml min-1,
respectively. In contrast, mass losses of SS from the Letterkenny soil at time zero
were significantly lower (Figure 3.9).
87
Chapter 3
Dissolved Reactive P (mg)
3.0
(a)
2.5
2.0
1.5
1.0
0.5
0
700
(b)
Suspended Sediment (g)
600
500
400
300
200
100
0
600
(c)
Particulate P (mg)
500
400
300
200
100
0
Figure 3.9 Mass loss of phosphorus and sediment in runoff from selected tillage
soils subjected to rainfall and two overland flow rates (225 and 450 ml min-1)
while inclined at a 10 degree slope.
88
Chapter 3
3.3.4 Effect of slope, time between events and overland flow rate on
concentrations in surface runoff
The GLMM (Table 3.2) used for the rainfall only experiment, indicated that the effect
of soil type on the FWMC of DRP (p = 0.013) and TRP (p = 0.007) depended on both
slope and time between rainfall events. The effect of soil type depended only on
surface slope for the FWMCs of SS (p = 0.044), TP (p = 0.014) and PP (p = 0.022) in
surface runoff.
The LMM analyses (Table 3.3) used for simulated rainfall/overland flow experiments,
indicated that the effect of soil type on the FWMC of DRP interacted with/depended
on both overland flow rate and time between overland flow events (p = 0.0351). The
effect of soil type depended only on overland flow rate for the FWMCs of SS (p <
0.0001), TP (p = 0.0015) and PP (p = 0.0013) in surface runoff. These results were as
expected given the small storm intervals being investigated. Larger storm intervals
that allow the soil time to dry might be expected to impact on the levels of SS, TP and
PP lost in runoff. There was significant interaction of soil type with at least one of the
experimental factors for each of the variables examined, indicating the importance of
soil type in assessing the potential to release DRP, SS, TP and PP into overland flow
(Table 3.3).
89
Chapter 3
Table 3.2 Overall Anova for responses from GLMM analyses (rainfall only)
Dissolved Reactive Phosphorus
Source
DF
F Value
Pr >F
Soil type
4
768.58
<.0001
Slope
1
67.38
<.0001
Rainfall interval
2
12.94
<.0001
Soil*Slope
4
240.97
<.0001
Soil*Rainfall interval
8
3.54
0.005
Slope*Rainfall interval
2
1.68
0.203
Soil*Slope*Rainfall interval
8
3.01
0.013
Soil type
4
904.11
<.0001
Slope
1
86.33
<.0001
Rainfall interval
2
15.19
<.0001
Soil*Slope
4
43.99
<.0001
Soil*Rainfall interval
8
5.14
0.001
Slope*Rainfall interval
2
2.53
0.097
Soil*Slope*Rainfall interval
8
3.4
0.007
Soil type
4
33.82
<.0001
Slope
1
10.62
0.002
Rainfall interval
2
32.7
<.0001
Soil*Slope
4
3.49
0.014
Soil type
4
27.51
<.0001
Slope
1
9.62
0.003
Rainfall interval
2
30.33
<.0001
Soil*Slope
4
3.13
0.022
Soil type
4
9.56
<.0001
Slope
1
14.87
0.001
Rainfall interval
2
9.65
0.001
Soil*Slope
4
2.66
0.044
Total Reactive Phosphorus
Total Phosphorus
Particulate Phosphorus
Suspended Sediment
90
Chapter 3
Table 3.3 Overall ANOVA for responses from LMM analyses (rainfall and
overland flow)
Dissolved Reactive Phosphorus
Source
DF
F value
Pr >F
Soil type
4
639.33
<.0001
Overland flow
2
17.71
0.0001
Overland flow*soil type
8
29.95
<.0001
Event
2
32.69
<.0001
Event*soil type
8
5.24
0.004
Overland flow*event
4
3.89
0.0116
Soil type*overland flow*event
16
2.14
0.0351
Soil type
4
23.17
<.0001
Overland flow
2
83.87
<.0001
Overland flow*soil type
8
6.00
0.0015
Event
2
65.05
<.0001
Event*soil type
8
4.04
0.0011
Overland flow*event
4
1.24
0.3053
Soil type
4
21.27
<.0001
Overland flow
2
83.3
<.0001
Overland flow*soil type
8
6.16
0.0013
Event
2
80.73
<.0001
Event*soil type
8
5.23
0.0004
Overland flow*event
4
1.49
0.2310
Soil type*overland flow*event
16
1.75
0.0898
Soil type
4
5.29
0.0073
Overland flow
2
35.13
<.0001
Overland flow*soil type
8
9.89
<.0001
Event
2
36.46
<.0001
Event*soil type
8
1.66
0.1348
Overland flow*event
4
1.91
0.1240
Total Phosphorus
Particulate Phosphorus
Suspended Sediment
91
Chapter 3
3.4 Summary
This study quantified the amount of DRP, PP, TP and SS released into overland flow
for 5 tillage soils when subject to a rainfall intensity of 30 mm hr -1 and overland flow
rates of 225 and 450 ml min-1 applied in 3 successive events. The main conclusions
from this study were:
1. Increasing the overland flow rate over the soil surface in the presence of
rainfall had the effect of increasing the concentrations of SS, PP and TP (but
not DRP) in surface runoff across all soils. This increase in concentration
varied in magnitude across soils and was highest for the Tullow soil.
2. Overall, there was no evidence of a relationship between overland flow rate
and DRP concentration measured in surface runoff. This implies that rainfall
can be used in isolation when developing relationships between soil P level
and potential DRP lost in runoff.
Chapter 4 uses the data compiled in this chapter to determine threshold STP values
above which surface runoff may cause eutrophication and to identify potential risk
indicators for estimating P and sediment release from tillage soils.
92
Chapter 4
Chapter 4 Threshold values and potential risk indicators
for estimating phosphorus and sediment release from Irish
tillage soils
Overview
Using the relationship between STP level in the 5 soils and the DRP measured in
surface runoff, a runoff dissolved phosphorus risk indicator (RDPRI) was developed
to quantify the Morgan‟s P (Pm), Mehlich 3-P (M3-P), water extractable P (WEP),
calcium chloride extractable P (Pcacl2) and soil P saturation (Psatox) levels for 5 Irish
tillage soils, above which there may be a potential threat to surface water quality. A
statistical analysis of the experimental data collected in Chapter 3 is used to rank soil
extractable P methods with respect to their potential to be used as P loss risk
indicators. Finally, this chapter identifies important parameters for which to test when
attempting to predict SS, TP and PP loss from tillage soils. The contents of this
chapter are, in part, published in the Journal of Environmental Quality (39:185-192,
2010).
4.1 Introduction
In tillage soils excessive organic and inorganic fertiliser application can lead to a build
up of P in excess of crop requirements. This may result in DRP loss in runoff, which
is readily available for biological uptake, and poses an immediate threat for
accelerated algal growth, which may negatively affect water quality in rivers and
lakes. Summer storm events, coupled with impervious agricultural soils, can lead to P
addition to waterways during the aquatic growing season. While PP loss can be
minimised through the use of buffer zones and minimum tillage, reducing dissolved P
loss is far more difficult (McDowell and Sharpley, 2001) and control measures are
mainly limited to preventing soil P accumulation to environmentally sensitive levels
(Sibbesen and Sharpley, 1997). The relationship between STP and DRP loss to water
in runoff events needs to be adequately understood and quantified for local soils
93
Chapter 4
(Wright et al., 2006) in order to determine upper critical limits for P in soil that will
reduce the risk of diffuse losses from tillage land to surface waters. A laboratory
flume study was chosen over a field study, as soils in flume studies can be
homogenised minimising variability in soil physical and chemical characteristics. It is
also less expensive and facilitates testing under standardised conditions including
surface slope, soil conditions, rainfall intensity and overland flow rate.
Soil P saturation has been suggested as a method to characterise the potential for P
loss from agricultural soils (Maguire et al., 2001). It is thought to be a more
meaningful indicator of potential losses to water than STP, since it describes soil in
terms of P sorption sites already saturated and is therefore independent of soil type
(Daly et al., 2001). Furthermore, its use in soil studies and for environmental purposes
is becoming more frequent (Beauchemin and Simard, 1999; Kleinman et al., 2000;
Schroeder et al., 2004; Randall et al., 2005; Penn et al., 2006; Little et al., 2007; Wang
et al., 2010). It has been shown to be well correlated with runoff DRP (Pote et al.,
1996; Pote et al., 1999). Soils with higher Psat ox pose a greater risk of P loss because
eroded soil particles will be enriched in potentially desorbable P (Pautler and Sims,
2000). Dilute calcium chloride has also been proposed for use as an environmental P
test because it represents more readily desorbable forms of P in soil (Daly and Casey,
2005) and has been shown to relate well to the concentration of DRP in surface runoff
from soil using rainfall simulators (McDowell and Sharpley, 2001). These proposed
environmental P loss risk indicators need to be assessed with respect to their potential
to predict DRP loss in surface runoff from Irish tillage soils.
The aims of this chapter were: (1) to investigate the relationships between soil
extractable P measured in the study soils and DRP measured in surface runoff; (2) to
rank a number of soil extractable P methods with respect to their potential to be used
as P loss risk indicators for a selection of Irish tillage soils; (3) to rank the importance
of soil physical and chemical parameters, in the prediction of P and sediment release
from soil to overland flow; and (4) to develop a RDPRI for tilled soils in Ireland.
94
Chapter 4
4.2 Data analysis methods
A RDPRI was first developed using the relationship between the FWMC of DRP lost
in surface runoff (generated by using simulated rainfall events on soils inclined at 10
and 15 degree slopes – Section 3.2.2) and each of Pm, M3-P and WEP measured in the
study soils. This was achieved by constructing 95% confidence limits around the DRP
relationships using the upper and lower confidence bands for the linear predictor. The
resulting confidence limits were then back-transformed from the log-linear model to
the original scale to identify the level of each P measure above which there might be a
potential threat to surface water quality. An improved RDPRI was developed using
the relationship between FWMC of DRP lost in surface runoff (generated by using
simulated rainfall/overland flow events on soils inclined at 10 degree slopes – Section
3.2.3) and each of Pm, M3-P, WEP, Pcacl2, and Psatox.
In order to rank the relative importance of the soil parameters, each parameter was
assessed individually in its effect on the surface runoff response, averaged across the
overland flow rates and flow events. The Akaike Information Criterion (AIC) is a
statistic that gives a measure of the goodness of fit of a model. As the models were
not nested, the extractable soil tests were ranked according to the AICs of their
individual models. Each P indicator was added, in turn, to a model with the
experimental factors and then assessed to determine which of them produced the best
increase in the goodness-of-fit of the model used for determining the critical value of
DRP.
4.3 Analysis of experimental results from simulated rainfall/overland flow
experiments
4.3.1 Soil properties affecting phosphorus release from soil to water
Selected soil chemical and physical properties are presented in Chapter 3, Table 3.1
and briefly discussed in Section 3.1.1. A more detailed discussion of study soil
properties affecting the release of P from soil to runoff is provided here. In particular,
attention is paid to Al-oxides, Fe-oxides and CaCO3, as these properties play an
95
Chapter 4
important role in the relationship between each of Pcacl2 and Psatox and DRP in
surface runoff.
Soil pH ranged from 6.4 - 7.7 with most of the soils in the neutral or slightly acid
range, as is common for tillage soils in Ireland where the median pH value is 6.4 (Fay
et al., 2007). Only the Bunclody soil had a pH > 7. The CaCO 3 content of the soils
ranged from 0.4 - 2.6%, with the Letterkenny and Bunclody soils being classed as
slightly calcareous (CaCO3 > 1%) in nature according to Defra (2006b). While CaCO3
is the soil parameter most notably associated with P sorption in calcareous soils
(Lindsay, 1979), some studies have found that P sorption in these soils is more closely
related to Fe-oxide, Al-oxide and clay contents (Leytem and Westermann, 2003).
Generally, soils with higher pH would normally contain greater amounts of
extractable Ca, as well as lower extractable Al and Fe (Brady and Weil, 2007). The
Bunclody soil was an exception in the studied soils, for while it had high CaCO3, it
contained relatively large amounts of Fe ox and Alox, even though its pH was 7.7. This
high pH may be a result of the use of beet factory sludge lime as a fertiliser in the
past. Reactions that reduce P availability occur in all ranges of soil pH, but can be
very pronounced in alkaline soils (pH > 7.3) and in acidic soils (pH < 5.5) (Busman et
al., 1998). In acidic soils, P is largely sorbed to Al-oxides and Fe-oxides, whereas in
neutral to alkaline soils, P occurs primarily as Ca-phosphates and Mg-phosphates
often precipitated, or sorbed, onto Ca and Mg carbonates (McDowell et al., 2012).
The presence of free CaCO3 in calcareous soils reduces the amount of soluble P
present in soil and prevents it from being released into runoff (Torbert et al., 2002). In
this study, the CaCO3 levels present in the soils were sufficiently low that the more
likely controllers of P retention were Fe and Al.
Distilled water and dilute calcium chloride solutions are not considered to be soil test
indices like P m and M3-P, but are thought to be good predictors of dissolved P in
runoff and subsurface drainage, respectively (Pote et al., 1996, 1999; McDowell and
Sharpley, 2001; Maguire and Sims, 2002) and their use as environmental indicators of
P loss has been proposed by Irish researchers (Daly and Casey, 2005). Calcium
chloride extractable P, WEP and Pm extracted the least amount of P from the soil,
while M3-P extracted larger amounts because it is more acidic than the soil solution
96
Chapter 4
and able to dissolve calcium phosphates, a fraction of P which is normally of low
availability. Both distilled water and 0.01 M CaCl2 possess extraction matrices that
closely mimic soil solution and their limited propensity to extract P, measuring
primarily the soluble and easily desorbable P forms results in lower P extraction
(Schindler et al., 2009). Distilled water extracts more P than 0.01 M CaCl2 because
Ca2+ enhances P sorption in soils. Acid ammonium oxalate extracted much larger
amounts of soil P than other extractants, suggesting that most of the P in the study
soils was sorbed or precipitated on amorphous oxides of Fe and Al (Pote et al., 1996).
There is general agreement in the literature that these forms of Fe and Al are the most
important in terms of P retention. Levels of Feox measured in all the soils of this study
were higher than Alox, as was the case for Irish soils studied by Doody et al. (2006),
Maguire et al. (2001) and Evans and Smiley (1976). While reports differ on their
relative importance in terms of P retention, Evans and Smiley (1976) showed that Alox
was two and a half times as effective as Feox in retaining P in Irish soils. Schroeder et
al. (2004) demonstrated that areas with larger Feox to Alox ratio may produce
proportionally greater P loss as the STP increases than areas with lower Feox to Alox
ratio.
Soil P saturation has been identified as a potential P loss risk indicator for agricultural
soils because it has a strong relationship with runoff P concentrations (Maguire et al.,
2001; Sims et al., 2002). It is different from other soil P tests because it not only
considers the quantity of P present in a soil, but also includes the capacity of the soil
to retain additional P. Soils with higher Psat ox maintain higher solution P
concentrations, and any eroded soil particles will be enriched in potentially desorbable
P (Pautler and Sims, 2000). Based on Psatox levels measured across the study soils,
the Tullow and Clonmel soils had the highest risk of P desorption, with Psatox levels
of 36.2 and 28.8%, respectively. Maguire et al. (2001) found that for a selection of
Irish and American soils with high P levels, a single oxalate extraction for Al, Fe, and
P proved to be most useful for predicting long-term P desorption, through calculation
of the Psatox and for predicting the ability of the soils to sorb more P by calculating
free [Feox and Alox]. In Section 4.3.5, the effect of Feox and Alox on DRP in runoff is
investigated.
97
Chapter 4
4.3.2 A method for identifying the critical soil test phosphorus threshold above
which simulated rainfall-induced surface runoff may pose a threat to surface
water quality
Researchers have shown that there is a relationship between STP and runoff DRP
concentrations from both grassland (Pote et al., 1996; Torbert et al., 2002) and
cultivated soils (Cox and Hendricks, 2000; Vadas et al., 2005). The retention of P in
rivers is dominated by physical processes such as flow velocity, discharge, and water
depth. The long-term storage of P in the water column through chemical processes,
such as assimilation by river sediments, is inhibited by the rapid mobilisation and
transport that occurs during hydrological storm events. Furthermore, biological uptake
can account for the majority of dissolved P transformations in streams (Reddy et al.,
1999), thereby increasing the population of algae and aquatic plants, affecting the
quality of the water and disturbing the balance of organisms present within it.
Eutrophic symptoms in rivers are commonly linked to transient increases in DRP
concentrations during times of ecological sensitivity (Jarvie et al., 2006). A critical
STP threshold must exist above which runoff water will negatively affect surface
water quality. In order to estimate defensible, upper critical soil P limits that are
environmentally sensitive, it is necessary to first develop analytical methods that
measure soil P availability relevant to the release of soil P to runoff; quantify the
relationship between soil and runoff P; and identify transport potential for a site
(Sibbesen and Sharpley, 1997). Given that eutrophication has been shown to occur in
rivers with > 0.03 mg MRP L-1 and lakes with > 0.02 mg TP L-1, runoff water with >
0.03 mg P L-1 entering rivers and lakes is likely to contribute to further deterioration
in surface water quality. A RDPRI was developed to identify a STP threshold
(measured in terms of Pm, M3-P and WEP) above which runoff water across the 5
soils tested will have a DRP concentration greater than 0.03 mg P L-1. Present nutrient
advice for tilled crops in Ireland prohibits the application of fertiliser or manure at P m
concentrations above 10 mg L-1 (SI 610 of 2010) with the exception of potatoes, beet,
and turnips. In Delaware, a M3-P concentration > 100 mg kg
-1
is considered above
optimum and therefore no further P addition is recommended. This limit is based on
crop yield response to fertiliser P, but needs to be assessed from an environmental
standpoint. The RDPRI enables this assessment to be performed.
98
Chapter 4
The logarithm of the FWMC of DRP in surface runoff from each soil was linearly
related to the Pm of each soil for all rainfall events and slopes examined (Figure 4.1a f). A RDPRI was developed by constructing 95% confidence limits around the Pm DRP relationships using the upper and lower confidence bands for the linear
predictor. The resulting confidence limits were then back-transformed from the loglinear model to the original scale to identify the Pm level above which there might be a
potential threat to surface water quality (Figure 4.1g - h).
5
4
3
Log FWMDRP
2
1
0
5
4
3
2
1
0
0 2 4
FWMDRP (mg L-1)
0.10
6 8 10 12 14 16 18
0.08
0.06
0.04
0.02
0
0
2
4
6 8 10 12 14 16 18
0 2 4
6
8 10 12 14 16 18
Morgan‟s P (mg L-1)
Figure 4.1 Runoff Dissolved Phosphorus Risk Indicator using Morgan’s P for
selected tillage soils at 10 and 15 degree slopes under a 30 mm hr-1 rainfall.
Rainfall 1 - 1st rainfall event; Rainfall 2 - 1hr after Rainfall 1; Rainfall 3 - 24 hr
after Rainfall 2.
99
Chapter 4
The logarithm of the FWMC of DRP in surface runoff from each soil was also
linearly related to WEP and M3-P levels in each soil for all rainfall events (Figure 4.2
and Figure 4.3).
The DRP lost to runoff water increases log-linearly as Pm, WEP and M3-P increase.
For both 10 and 15-degree slopes, the 1st rainfall event had the highest FWMC of
DRP in runoff from the 5 soils. The FWMC of DRP in surface runoff was lower
during the 2nd rainfall event, but was higher in the 3rd. This increase may be attributed
to more soluble P becoming available to runoff during a 24-hr rainfall cessation than
during a 1-hr cessation. Sharpley (1980) found that when the time interval between
rainfall events exceeds 1 d, the initial soluble P concentration in a runoff event
increases.
Research by Foy et al. (2002) and Pote et al. (1999) found a positive relationship
between Pm and DRP lost in surface runoff. A positive relationship between surface
runoff DRP and soil WEP was also reported by Penn et al. (2006), McDowell and
Sharpley (2001), Wang et al. (2010) and Pote et al. (1996) using simulated rainfall on
packed soil boxes. The WEP method closely mimics the interaction between
rainwater and soil particles, and provides a good indication of DRP in runoff.
Under the test conditions, the 95% confidence bands indicated that provided the P m
does not exceed 9.5 mg L-1, WEP does not exceed 4.4 mg kg -1 and M3-P does not
exceed 67.2 mg kg-1, for tillage soils, the concentration of DRP in runoff will be
within the currently acceptable P range for surface water quality of < 0.03 mg L-1. The
finding for Pm in this study reinforces the statutory requirements of SI 610 of 2010,
which prohibit fertiliser application to tillage soils with a P m > 10 mg L-1 (Figure 4.1
h). In contrast, agri-environmental interpretation of M3-P in Delaware indicated that
improved P management was necessary to reduce potential for nonpoint P pollution
when M3-P > 150 mg kg-1 (Sims et al., 2002). However, some states are adopting a
critical M3-P of 65 mg kg
-1
(Sharpley, 1995) as a cutoff point in their P indexing
systems for rating the potential for P loss in runoff. A M3-P of 65 mg kg -1 represents
the level at which no yield response to fertiliser P addition is expected (Sharpley,
1995) and is therefore comparable to the Pm of 10 mg L-1 used in Ireland which is also
100
Chapter 4
based on there being no response to fertiliser P addition above this value. Fang et al.
(2002) deemed a M3-P of 65 mg kg-1 to be more consistent with environmental levels
of P that produce eutrophication than are higher values used by many other states such
as Delaware. The threshold M3-P of 67.2 mg kg -1 determined in this study is in close
agreement with the critical M3-P of 65 mg kg-1 now being adopted in some states.
5
4
3
2
Log FWMDRP
1
0
5
4
3
2
1
0
FWMDRP (mg L-1)
0.12
0
2
4
6
8
10
12
0.10
0.08
0.06
0.04
0.02
0
0
2
4
6
8
10
12
0
2
4
6
8
10
12
Water Extractable Phosphorus (mg kg-1)
Figure 4.2 Runoff Dissolved Phosphorus Risk Indicator using water extractable
P for selected tillage soils at 10 and 15 degree slopes under a 30 mm hr-1 rainfall.
Rainfall 1 - 1st rainfall event; Rainfall 2 - 1 hr after Rainfall 1; Rainfall 3 - 24 hr
after Rainfall 2.
101
Chapter 4
5
4
3
2
Log FWMDRP
1
0
5
4
3
2
1
0
0
FWMDRP (mg L-1)
0.10
20
40
60
80
100
0.08
0.06
0.04
0.02
0
0
20
40
60
80
100
0
20
40
60
80
100
Mehlich 3 Extractable Phosphorus (mg kg-1)
Figure 4.3 Runoff Dissolved Phosphorus Risk Indicator using Mehlich-3
extractable P for selected tillage soils at 10 and 15 degree slopes under a 30 mm
hr-1 rainfall. Rainfall 1 - 1st rainfall event; Rainfall 2 - 1hr after Rainfall 1;
Rainfall 3 - 24 hr after Rainfall 2.
Under the test conditions, the 95% confidence bands indicated that provided the P m
does not exceed 9.5 mg L-1, WEP does not exceed 4.4 mg kg -1 and M3-P does not
exceed 67.2 mg kg-1, for tillage soils, the concentration of DRP in runoff will be
within the currently acceptable P range for surface water quality of < 0.03 mg L-1. The
finding for Pm in this study reinforces the statutory requirements of SI 610 of 2010,
which prohibit fertiliser application to tillage soils with a P m > 10 mg L-1 (Figure 4.1
h). In contrast, agri-environmental interpretation of M3-P in Delaware indicated that
improved P management was necessary to reduce potential for nonpoint P pollution
when M3-P > 150 mg kg-1 (Sims et al., 2002). However, some states are adopting a
102
Chapter 4
critical M3-P of 65 mg kg
-1
(Sharpley, 1995) as a cutoff point in their P indexing
systems for rating the potential for P loss in runoff. A M3-P of 65 mg kg -1 represents
the level at which no yield response to fertiliser P addition is expected (Sharpley,
1995) and is therefore comparable to the Pm of 10 mg L-1 used in Ireland which is also
based on there being no response to fertiliser P addition above this value. Fang et al.
(2002) deemed a M3-P of 65 mg kg-1 to be more consistent with environmental levels
of P that produce eutrophication than are higher values used by many other states such
as Delaware. The threshold M3-P of 67.2 mg kg -1 determined in this study is in close
agreement with the critical M3-P of 65 mg kg-1 now being adopted in some states.
The predictions of the RDPRI may be affected by soil management decisions.
Changes in soil management resulting in the deterioration or improvement of soil
drainage may have the effect of reducing the predictive power of the RDPRI by
increasing or decreasing, respectively, the overland flow volume generated in upslope
soil areas.
The next step is to examine the effect of increasing overland flow rates, which may
result from intense summer rainfalls (more frequent intense rainfall events during the
summer have been predicted for Ireland by Sweeney et al. (2008)), on the critical soil
test P thresholds developed using the RDPRI.
4.3.3 Investigating the effect of increasing overland flow rates on the critical soil
test phosphorus threshold above which runoff may pose a threat to surface water
quality
The logarithm of the FWMC of DRP in surface runoff from each soil was linearly
related to the Pm of each soil for all overland flow rates (Figure 4.4). The logarithm of
the FWMC of DRP in surface runoff from each soil was also linearly related to WEP,
M3-P, Pcacl2 and Psatox concentrations in each soil for all overland flow/simulated
rainfall events (Appendix C).
As was the case for simulated rainfall only experiments, the DRP lost in surface
runoff when both simulated rainfall and overland flows were applied to the soil,
103
Chapter 4
increased log-linearly as Pm, WEP, M3-P, Pcacl2 and Psatox increased. For both
overland flow rates on all 5 soils, the time zero event had the highest FWMC of DRP
in runoff, while the FWMCs of DRP in surface runoff for the 1-hr and 24-hr events
were lower than the time zero event, and were generally not significantly different in
magnitude from each other.
Time zero
1-hr interval
33
3
3
22
2
2
11
Ln(DRP) =1.4265 + 0.1651*P m
00
1
Ln(DRP) = 1.4682 +0.1388*P m
0
0
2
4
6
8
10
12
14
16
(d)
44
33
18
5
0
2
4
6
0
8 10 12 14 16 18 0
5
(e)
4
4
3
3
Ln(DRP) = 1.3564 + 0.1544*P m
2
4
6
8 10 12 14 16 18
(f)
1
``
22
2
2
Ln(DRP) = 2.1289 + 0.089*Pm
Ln(DRP) = 2.2665 + 0.0936*P m
11
00
55
0
2
4
6
1
0
8 10 12 14 16 18 0
5
(g)
Ln(DRP) = 2.3167 + 0.0779*P m
1
2
4
6
0
8 10 12 14 16 18 0
5
(h)
44
4
4
33
3
3
22
Ln(DRP) = 2.2814 + 0.0842*Pm
0
2
4
0 2 4
6
0
8 10 12 14 16 18 0
6
8 10 12 14 16 18
(i)
Ln(DRP) = 2.1751 + 0.0834*P m
1
1
00
4
2
2
Ln(DRP) = 2.2499 + 0.0956*Pm
11
2
1
Runoff rate = 425 ml min-
55
Log FWMDRP
1
2
4
6
6 8 10 12 14 16 18 0 2 4 6
0
8 10 12 14 16 18 0
2
4
6
Rainfall/overland (450)
4
Rainfall/overland (225)
4
Simulated rainfall only
44
(c)
Runoff rate = 650 ml min-1
(b)
(a)
24-hr interval
5
Runoff rate = 425 ml min-
5
Runoff rate = 200 ml min-1
55
8 10 12 14 16 18
8 10 12 14 16 18 0 2 4 6 8 10 12 14 16 18
Morgans P (mg L-1)
Figure 4.4 Log FWMDRP against Morgan’s P for selected tillage soils at a 10
degree slope under a 30 mm hr-1 rainfall and subjected to two distinct overland
flow rates (225 and 450 ml min-1).
A critical value was derived for each of the soil extractable P methods above which
surface water quality may exceed 0.03 mg L-1 DRP, the concentration above which
water quality may deteriorate (Figure 4.4). The 95% confidence limits/intervals
104
Chapter 4
(Figure 4.5 and Figure 4.6) indicate that provided the P m does not exceed 7.83 mg L-1,
WEP does not exceed 4.15 mg kg -1, M3-P does not exceed 61 mg kg -1, Pcacl2 does
not exceed 1.2 mg kg-1 and Psatox does not exceed 17.1% for tillage soils, the
concentration of DRP in surface runoff will be below 0.03 mg L-1. While these new
values for Pm, WEP, and M3-P are lower than those determined using the RDPRI in
the previous section, the confidence intervals of the two sets of values overlap by
more than 25% (Table 4.1), indicating no significant difference. Furthermore, in an
attempt to improve on the model used in the previous section, an all-in approach was
adopted which allowed testing of any change from one combination of experimental
factors to the next. The estimates of noise/error, on which confidence intervals are
based, are better when all available data are used (as in the case of an all-in approach)
and this may, in part account for the new lower values and narrower confidence
intervals (Table 4.1) of the new model. The larger data set used in this model also
improved its precision. Overall, there was no evidence of a linear relationship
between overland flow rate and DRP concentration measured in surface runoff (p =
0.125). As such, the new lower values for Pm, WEP, and M3-P are primarily a result
of improvements in the model used to predict DRP in runoff. Table 4.1 shows the
95% confidence limits for each P extraction method at water quality limits of 0.03 and
0.035 mg L-1 (the new limit as set by SI 272 of 2009).
The finding for Pm in this study is in close agreement with the agronomic optimum
(Pm = 6.1-10 mg L-1) for plant growth and crop yields. It is also in close agreement
with the statutory requirements of SI 610 of 2010, which prohibits fertiliser
application to tillage soils with a Pm > 10 mg L-1 (Figure 4.5a) and suggests that given
the worst case storm scenarios tested, a change in the statutory P m limit is
unwarranted at this time. However, the results also suggest that the limit may have to
be reviewed in future should the predicted climate change storms materialise. The
threshold M3-P of 61.2 mg kg-1 determined in this study is in close agreement with
the critical M3-P of 65 mg kg-1 now being adopted in some states in America (Figure
4.5c). A M3-P value of 45-50 mg kg-1 in soil is generally considered to be optimum
for plant growth and crop yields (Sims, 2000). The findings of this study suggest that
keeping the M3-P level in the study soils close to this agronomic optimum will ensure
that the concentration of DRP in overland flow will be below 0.03 mg L-1. Caution
105
Chapter 4
must be exercised when interpreting STP results in an environmental context as they
comprise only a small percentage of the total soil P reservoir and do not account for
potential detachment (Haygarth and Condron, 2004) or dissolution from eroded
sediments.
Table 4.1 Comparing RDPRI outputs (Section 4.3.4) with new model outputs
(this section) to identify thresholds for each of Morgan’s P, P m; water extractable
phosphorus, WEP; Mehlich-3, M3-P; calcium chloride extractable phosphorus,
Pcacl2; and Soil P saturation, Psatox, above which DRP in surface runoff may
exceed 0.03 mg L-1 (old limit - SI 258 of 1998) and 0.035 mg L-1 (new limit - SI
272 of 2009)
Upper 95% CL
Lower 95% CL
% Overlap of
intervals
RDPRI Pm (mg L-1)
Phosphorus < 0.03 mg L-1
(Old limit - SI 258 of 1998)
9.5
16
New Pm (mg L-1)
7.83
11.31
RDPRI WEP (mg kg-1)
4.4
6.83
New WEP (mg kg-1)
4.15
6.57
RDPRI M3-P (mg kg-1)
67.2
89.3
New M3-P (mg kg-1)
61.2
76
Pcacl2 (mg kg-1)
1.2
1.79
Psatox (%)
17.1
24.1
-
Pm (mg L-1)
Phosphorus < 0.035 mg L-1
(New limit - SI 272 of 2009)
8.78
12.42
-
-1
WEP (mg kg )
4.87
7.35
-
-1
M3-P (mg kg )
65.2
80.5
-
-1
Pcacl2 (mg kg )
1.36
2
-
Psatox (%)
18.9
26.44
-
52
89
59
CL, Confidence limit
Soil P tests developed for environmental purposes such as WEP and Pcacl2 are less
affected by soil type than agronomic soil tests like M3-P and Pm (Self-Davis et al.,
2000), and can be valuable for estimating labile forms of P (Simard et al., 1995).
106
Chapter 4
Their ability to either better mimic the interaction between soil and runoff or better
represent the likelihood of P release from soil to runoff (Vadas et al., 2005), makes
them ideal P loss risk indicators. The relationships developed in this study, between
runoff DRP and each of WEP and Pcacl2 are shown in Figure 4.5b and Figure 4.6a,
respectively. In Section 4.3.5, soil P tests are compared to determine the best
environmental P loss indicator.
A Psatox value of 25%, or more, has been established on the basis of laboratory data
for non-calcareous, sandy soils from the Netherlands, as a critical value above which
the potential for P losses through runoff and leaching become unacceptable. The water
quality standard used in the Netherlands in determining this critical value is 0.1 mg
ortho-P L-1 ( Breeuwsma et al., 1995). The findings of the present study suggest that a
lower limit of 17.1% Psatox would better protect water quality in Ireland by ensuring
that DRP losses from Irish tillage soils remain below 0.03 mg P L-1. Soils with Psat ox
values that exceed this threshold, such as the Tullow and Clonmel soils, are more
heavily saturated with P and are vulnerable to losses to overland flow by P desorption.
This lower Psatox threshold, as determined for Irish tillage soils subjected to overland
flow and rainfall in this study, may be a result of the P sorption capacity (PSC) used
in the determination of Psatox for the study soils. The PSC of soils varies widely
depending on clay content, clay mineralogy, OM content, exchangeable Al, Fe, and
Ca concentrations, and soil pH (Tisdale et al., 1993). Furthermore, results from
Beauchemin and Simard (1999) indicate that the relationship between PSC and [Fe ox
+ Alox] contents may vary among soil groups. Therefore, the extension of the 25%
Psatox threshold to other soil types and other water quality standards may not be
appropriate (Beauchemin and Simard, 1999).
107
Chapter 4
FWMDRP (mg L-1)
Morgan’s P (mg L-1)
Water Extractable P (mg kg-1)
Mehlich-3 P (mg kg-1)
Figure 4.5 Runoff Dissolved Phosphorus Risk Indicator using Morgan’s P, water
extractable P, and Mehlich-3 P for selected tillage soils at a 10 degree slope,
under a 30 mm hr-1 rainfall and subjected to two distinct overland flow rates
(225 and 450 ml min-1).
108
FWMDRP (mg L-1)
Chapter 4
CaCl2 Extractable P (mg L-1)
Soil P Saturation (%)
Figure 4.6 Runoff Dissolved Phosphorus Risk Indicator using calcium chloride
extractable P and soil P saturation for selected tillage soils at a 10 degree slope,
under a 30 mm hr-1 rainfall and subjected to two distinct overland flow rates
(225 and 450 ml min-1).
4.3.4 Phosphorus and sediment loss risk indicators for Irish tillage soils subjected
to simulated rainfall
Soil extractable P can be measured in numerous ways in an attempt to predict DRP
available to runoff. A stepwise regression selection procedure was used to identify the
soil extractable P method best suited to predicting DRP loss from Irish tillage soils
when subjected to simulated rainfall. Measurement of soil WEP was selected as
important when predicting DRP in runoff across all soils, slopes and rainfall events.
109
Chapter 4
This is in agreement with other studies where WEP provided the strongest correlation
with DRP concentrations in runoff when compared to other STP methods such as M3P and Pm (Pote et al., 1996), and was able to simulate actual runoff DRP
concentrations (Yli-Halla et al., 1995). Soil parameters also selected as important by
stepwise regression to test for when predicting DRP in runoff were: STP, AgSt, pH,
Pcacl2, and clay. However, these were only selected for certain combinations of slope
and rainfall event, and are less reliable indicators.
The effect of soil type on SS, PP and TP loss to water differs depending on slope and,
consequently, the stepwise procedure was performed separately for each slope. The
soil parameters selected as important (by stepwise regression) in predicting SS loss to
water were: OM, AgSt and clay. Soil P saturation was selected as important when
predicting PP and TP lost in runoff.
4.3.5 Phosphorus loss risk indicators for Irish tillage soils subjected to simulated
rainfall/overland flow
This study‟s findings show that, for soils subjected to both simulated rainfall and
overland flow, WEP (AIC = -12.3) and Pcacl2 (AIC = -6.9) performed better than the
agronomic soil P tests, Pm (AIC = 6.4) and M3-P (AIC = 7.1) in predicting DRP in
overland flow. Only large differences (here, at least 4) between AIC scores for soil
extractable P tests are taken as indicating a difference in the goodness of fit of the
model used for determining the critical value of DRP. The addition of WEP to the
model produced the best increase in goodness of fit (as is evidenced by WEP
receiving the lowest AIC score of -12.3 and the difference between it and the next
lowest AIC being > 4) and therefore performed better than the other measures of soil
P when predicting DRP in runoff. These results are in agreement with Pote et al.
(1999), Penn et al. (2006) and Wang et al. (2010), who reported that WEP had
consistently stronger relationships with DRP concentrations in surface runoff than
other measures of soil P. The authors attributed this to the fact that the extracting
solution for WEP (distilled water) is more similar to the simulated rainfall water (tap
water) than other extracting solutions.
110
Chapter 4
Soil P saturation (AIC = 20.6) ranked lowest, which is probably due to some of the
study soils being slightly calcareous in nature and the fact that soil texture ranged
from sandy loam to loam. In other studies, the calculation of Psat ox using Pox/0.5(Alox
+ Feox) has produced strong correlations with soluble P when the range of soils used
was homogeneous, but the relationship weakens if a wider range of soil types is
considered (Beauchemin and Simard, 1999). In a study of published data from 17
studies, Vadas et al. (2005) concluded that for noncalcareous soils, a test for soil P
saturation (determined by acid ammonium oxalate extraction) may provide a more
universal prediction of dissolved P in runoff than Mehlich-3, Bray-1, or water
extractions. Soil P saturation measures the degree to which soil P sorption sites have
been filled and has been found to be a good indicator of P availability to runoff
(Kleinman and Sharpley, 2002).
Due to the limited range in some of the soil parameters measured across the 5 soils
(this was largely due to tillage in Ireland being conducted primarily on sandy loam or
loam soils), it was difficult to ascertain which parameters had a significant effect on
DRP in overland flow. Although higher levels of Feox and [Feox + Alox] measured in
the study soils were found to have a significant lowering effect (p < 0.05) on the
concentration DRP measured in overland flow. This further emphasises the important
role these parameters play in determining the PSC of a soil.
4.4 Summary
The RDPRI, developed in this chapter, identified the levels of extractable soil P above
which surface runoff, resulting from both high intensity simulated rainfall and
overland flow on tillage soils, may, under the conditions tested, pose a threat to
surface water quality. The main conclusions from this study were:
1. Tilled soils, subjected to simulated rainfall only, may produce surface runoff P
concentrations in excess of 0.03 mg L -1 (the value above which eutrophication
of rivers is likely to occur) if their P m, WEP, and M3-P concentrations exceed
9.5 mg L-1, 4.4 mg kg-1, and 67.2 mg kg-1, respectively.
111
Chapter 4
2. The RDPRI developed here, using both simulated rainfall and overland flow,
showed that provided the P m does not exceed 7.83 mg L-1, WEP does not
exceed 4.15 mg kg-1, M3-P does not exceed 61 mg kg-1, Pcacl2 does not
exceed 1.2 mg kg-1 and Psatox does not exceed 17.1% for tillage soils, the
concentration of DRP in surface runoff will be below 0.03 mg P L-1.
3. The finding for Pm in this study is in close agreement with the agronomic
optimum (Pm = 6.1-10 mg L-1) used in Ireland for plant growth and crop
yields. It is also in close agreement with the statutory requirements of SI 610
of 2010, which prohibits fertiliser application to tillage soils with a P m > 10
mg L-1.
4. Of the five soil extractable P methods investigated, WEP was identified, using
stepwise linear regression, as having the greatest potential to be used as an
indicator of the risk of P movement from soil into runoff water via dissolution.
Despite its apparent advantage over Pm in determining environmental risk, it
would appear to be impractical and costly to run two soil P tests side by side
given that Pm gives a good approximation for both agronomic and
environmental purposes.
Chapters 3 and 4 have shown that if current guidelines for P application to tillage soils
are adhered to, the risk of P loss via dissolution should be minimal. They have also
demonstrated that P loss from tillage soils, when subjected to high intensity rainfall
and overland flow under controlled conditions, is primarily associated with eroded
sediment and is therefore a potential long-term source of P for algae and plant uptake
from surface water bodies, in particular lakes. The next step, therefore, is to determine
the erosion risk posed by the study soils in their natural field conditions. The
identification of tillage fields posing a high erosion risk will facilitate the deployment
of mitigation measures solely in those fields, thereby greatly reducing the number of
sites susceptible to PP loss in Ireland. Chapter 5 details the development and
preliminary testing of a novel screening toolkit with which, the farmer/specialist
advisor can identify fields where the erosion risk is high and soil quality is an issue.
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Chapter 5 Physical, chemical and visual evaluation of six
Irish tillage soils to assess soil quality and susceptibility to
erosion
Overview
In this chapter, a novel screening toolkit is developed with which farmers/specialist
advisors can screen tillage fields for likelihood of erosion and reduced soil quality. At
each study site, detailed soil classification results and simple soil quality assessments
are used in conjunction with observed erosion levels to select the most appropriate
indicators for assessing erosion risk and soil quality status. A preliminary validation
of the screening toolkit showed it to be effective at identifying fields where the
erosion risk is high.
5.1 Introduction
Soil is a vital, non-renewable resource which requires sustainable management to
ensure the viability of food and fibre production, nutrient retention and cycling, and
filtration of water (Creamer et al., 2010). Emergence of policies, such as the proposed
SFD which deal with concern over soil degradation and anthropogenic impacts to soil
is likely to increase the requirement for assessment of soil quality and identification of
soils at risk from degradation (Bone et al., 2012). The SFD aims to establish a
common framework to protect, preserve and prevent further degradation to soil and its
associated functions. Under the SFD, seven main threats to soil quality are
recognised: erosion (water, wind and tillage), decline of SOC, compaction,
contamination, salinisation, landslides and desertification (Soil Strategy in 2006
(COM (2006) 231). The European project „Sustainable Agriculture and Soil
Conservation (SoCo) - Case Studies‟ identified the main concerns for sustainable
agriculture on soils in Northern and Western Europe to be: soil erosion by water,
decline in SOM; diffuse soil contamination, in particular contamination associated
with nitrates and agrochemicals; and compaction. The case studies, which comprised
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the Northern and Western European area of the SoCo project, were concerned
primarily with arable farming systems under intensive management conditions. It is in
these systems primarily, that efforts must be focused in order to reduce soil erosion
losses closer to tolerable levels.
Despite having a significantly higher P export coefficient (Doody et al., 2012) and
greater susceptibility to erosion (Boardman and Poesen, 2006), arable land, which
comprises 10% of agricultural area utilised in Ireland, has not received the same level
of attention as the more dominant grassland and therefore its current degradation
status and contribution to surface water impairment in Ireland is as yet unknown.
Research in England and Wales by Chambers et al. (1992 and 2000) and Chambers
and Garwood (2000) has shown that soil erosion by water is most prevalent in the
southwest and southeast of the country where tillage cultivation on light sandy soils
predominates. Tillage land is similarly distributed in Ireland, with 48% of crop
production concentrated in the south of the country (Schulte et al., 2010a), where the
soils are highly suitable for tillage, having a light-to-medium texture. Unlike in
England and Wales, where extensive research of soil erosion in tillage areas has been
conducted at multiple scales over the past 40 years, research incorporating tillage in
Ireland is only recently underway in the form of the ACP. The ACP includes two
catchments (9.4 and 11.2 km2) with high proportions of winter wheat or spring barley
cropping (Wall et al., 2011), and will serve as the first assessment of nutrient and
sediment loss from agricultural catchments with high proportions of tillage land in
Ireland.
Ireland has a valuable resource in terms of its land and soil quality, and promoting
sustainable soil management is one of the areas of action included in Food Harvest
2020 (DAFF, 2010), the national strategy for the development of the agri-food sector.
Agricultural activities that can negatively impact on soil quality must be tackled if
Ireland is to meet the ambitious growth targets set out in this vision. However, the
main focus of current agricultural and environmental policy, such as the crosscompliance regulations that accompany the Common Agricultural Policy and agrienvironment schemes, tends to be on control of diffuse water pollution rather than
protecting or conserving soil in situ (Posthumus et al., 2011). As such, focus in
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Ireland has been on maintaining good water quality with limited attention being given
to soil quality. Just as soils are regularly tested for nutrients, their physical condition
must receive similar and appropriate attention (Batey, 2009) to ensure that
management practices are sustainable. Soil structure is a crucial soil property that
affects several processes important to soils productive capacity, environmental quality
and agricultural sustainability (Lal, 1991). Visual and tactile assessments of soil and
its structure carried out in the field can identify areas where soil threats like
compaction (vehicle or cultivation induced), decline of SOC (due to intensive
cultivation and removal of crop residues particularly on coarse textured soils), and
increased risk of surface runoff (due to reduced soil porosity) leading to erosion, may
be occurring. These threats to soil quality - if not prevented or their effects at least
mitigated - can put the soil‟s productivity and surface water quality at risk, thereby
hindering the achievement of the growth targets set out in Food Harvest 2020 and
delaying Ireland‟s progress towards achievement of at least “good status” for all
waters by December 2015 as required by the WFD (2000/60/EC: Council of the
European Union, 2000).
Given time and capital constraints, it would be prohibitive for countries to use
detailed quantitative approaches to assess all soils, because there is a possibility that
the highest priority soils would not be reached if such methods are employed (Bone et
al., 2012). Moreover, as government alone cannot take all the steps necessary to
safeguard our soil resource for future generations, farmers and other land managers
have an essential role to play in managing agricultural soils sustainably (Defra,
2009a). Methods of erosion risk assessment and soil structure and quality assessments
that are reliable, quick, inexpensive and conductible by specialist advisors and farmers
have been developed in other countries in recent years. In England, the system of riskassessment that farmers must follow as part of the „cross-compliance‟ regime is set
out in “Controlling Soil Erosion: a Manual for the Assessment and Management of
Agricultural Land at Risk of Water Erosion in Lowland England” Defra (2005). The
efficacy of the Defra (2005) risk assessment scheme was tested by Boardman et al.
(2009) and shown to be 90% successful (18/20 cases) at identifying high risk sites
given the land-use at the time. Visual Soil Assessment (VSA; Shepherd, 2009) and
Visual Evaluation of Soil Structure (VESS; Guimarães et al., 2011; Ball et al., 2011)
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based on the Peerlkamp test (Peerlkamp, 1959)) were developed in New Zealand and
the UK, respectively, to enable farmers to assess soil structure and quality. Mueller et
al. (2009) showed that soil structure scored by VSA and VESS was significantly
correlated with certain soil physical parameters (dry bulk density, soil strength, and
infiltration rate) affecting runoff, erosion and crop yields. Using these methods, soil
structure can be assessed quickly on site and the land user may take a positive part in
the evaluation (Batey and McKenzie, 2006), and by doing so, gain a better
understanding of the cultivation practices or management decisions that led to the
degradation. This is in keeping with the Department of Agriculture, Fisheries and
Food‟s Harvest 2020 vision which states that “primary producers have a valuable role
to play as guardians of the rural environment”.
While detailed quantitative assessment methods such as those employed to classify
soils in the Soil Survey of Ireland (Gardiner and Radford, 1980) are inappropriate for
assessing erosion risk and soil quality on a field by field basis, they do provide
valuable information on the erodibility (as determined using the K-factor (Wischmeier
and Smith, 1978)) of a mapped soil series. The K-factor is related to four crucial soil
properties triggering erosion: SOM, soil texture (particle size analysis), soil structure,
and permeability (Panagos et al., 2012). These properties were identified through
nationwide studies performed by the United States Department of Agriculture Natural Resources Conservation Service using rainfall simulation tests (USDANRCS, 2005). Data collected during the Soil Survey of Ireland and data currently
being collected as part of the new ISIS (Creamer et al., 2010) represents the only soil
physical data available in Ireland, and can be used to predict the soils potential for
erosion by water. If the SFD is eventually ratified, Ireland will be required to identify
areas where erosion has occurred in the past or is likely to occur in the future. At that
time, the soil information provided by ISIS will be essential in identifying these areas.
There is currently no standard for the assessment of erosion risk or soil quality in
Ireland. The aim of this study was to develop and conduct a preliminary validation of
a novel, easy-to-use screening toolkit for use by farmers/specialist advisors in
assessment of tillage fields for likelihood of erosion and reduced soil quality. Using
different methods of erosion risk, and soil structure and quality assessment applied to
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Chapter 5
a representative selection of Irish tillage soils in which different levels of erosion were
observed, I investigated the erodibility of these soils and developed and conducted a
preliminary validation of a screening toolkit containing a set of erosion and soil
quality indicators, which can be used by Irish tillage farmers to identify,
expeditiously, fields with high erosion risk and poor soil quality. The next step would
be to make the screening toolkit part of a larger-scale erosion study in which
measured edge-of-field sediment losses can be used to validate the erosion risk
rankings assigned using the screening toolkit.
5.2 Site information and soil assessment methods
5.2.1 Site selection
As previously mentioned in Section 3.2, the soils used in this study were broadly
representative of the major tillage areas in Ireland (Figure 3.1). Initially, fourteen
tillage field sites, spread across the major tillage areas of Ireland, were visited and the
most suitable sites (6 based on soil type, tillage history, slope and evidence of prior
erosion problems), which exhibited a wide range of physical and chemical properties,
were chosen. Tillage areas are predominantly in the midlands, south and east of
Ireland (Figure 3.1), where there is a drier climate (average annual rainfall (AAR) of
between ca. 657 and ca. 1400 mm) than other areas of the country and soil types are
free draining, making them more suitable for spring cultivations and less likely to be
damaged by harvesting machinery. This leads to better opportunities for seedbed
preparation and harvesting, as well as potentially higher grain yields (Collins and
Cummins, 1996). The physical characterisation of the soils, presented in Section 5.4,
was carried out between May and November 2007. The topography of the sites ranged
from undulating lowland at the Bunclody and Tullow sites to hilly at the Fermoy site.
The remaining sites had rolling lowland topography.
Exact location, tillage history, and observed erosion features are given in Table 5.1.
There was no identifiable reduction in grain yield over time at any of the sites
according to the land users, while yields were deemed to be improving at the Clonmel
site as a result of better husbandry. The Fermoy and Duleek sites relied solely on
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chemical fertiliser to supply crop nutrients because of a lack of easily accessible
organic fertilisers. On sites where organic fertilisers such as turkey litter, cheese
sludge and beet factory sludge lime were applied, the P m levels in the soil were higher
than those sites that relied solely on chemical fertiliser. At some sites (Table 5.1),
subsoiling to a depth of 36 cm was carried out along headlands of fields where tillage
compaction had occurred. On sites where OSR was planted, subsoiling was carried
out over the whole field area to shatter soil below plough depth, thereby eliminating
compact layers which inhibit the deep rooting crop.
5.2.2 Management history of the sites
Tramlines at the study sites were normally used 4-7 times per year for crops like OSR
and winter oats and up to 10 times per year for winter wheat. The greater number of
spraying operations associated with winter wheat and winter oat crops increases the
risk of compaction of tramlines and headlands. The planting of OSR facilitated the
use of minimum tillage in the following year. Minimum tillage was practiced in some
years on some sites (Table 5.1). In Ireland, minimum tillage normally involves: (1)
shallow cultivation using a tine cultivator or disc harrow to a depth of 75-100 mm
immediately followed by rolling; (2) spraying with herbicide a few days prior to
sowing, following a stale seedbed period of a number of weeks (where possible) to
eliminate volunteers and established weeds; and (3) sowing with a cultivator drill to a
target depth of 40 mm (Forristal and Murphy, 2009). The only dramatic erosion event
reported across all sites occurred at the Clonmel site, where there was a large wash
out of soil in November 2010, leaving a gully behind; gully dimensions were
approximately 35 m in length, 0.3 m in breath and 0.35 m in depth. The farmer used
earth moving equipment to return the soil after the crop was harvested.
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Table 5.1 Information on selected tillage sites.
Lat/long1
Tillage history
to present day
Bunclody,
Co. Wexford
52°36'59'' N,
6°33'46''W
Tullow, Co.
Carlow
52°48'00'' N,
6°50'07''W
Fermoy, Co.
Cork
52°08'00'' N,
8°12'08''W
1980- : rotation of beet
(replaced by beans/OSR
after 2005) every 3 yr and
cereals in between
1981- : rotation of beet
(replaced by WO/OSR
after 2005) every 3 yr and
cereals in between
1985- :rotation of beet
(beans/OSR after 2005)
every 4 yr and oats, barley
and wheat in between
1979- : rotation of beet
(replaced by WO/OSR
after 2005) every 3 yr and
cereals in between
1995- : rotation of
potatoes every 4 yr and
cereals in other years
Study site
Clonmel, Co. 52°16'51'' N,
7°47'07''W
Tipperary
Cultivation
Method
Erosion Features
Subsoiling
Organic
Fertiliser
Farmers Comments
Typical yields
(at 20% MC)
t/ha
Moderate compaction on
SW = 9.9;
headlands/tramlines prior
WB = 8.9;
to subsoiling
SB = 7.4
Ploughed to 18
Moderate erosion on
cm; some min
longer slopes during
till
prolonged heavy rainfall
2010 (36
cm)
BFSL (2001
and 2003)
Ploughed to 23
Moderate erosion on
cm; some min
longer slopes during
till
prolonged heavy rainfall
No
Turkey litter,
FYM, slurry,
and BFSL
Moderate compaction on
headlands/tramlines
WO = 8.6;
WW = 11.112.4
Ploughed to 20 Accumulation of soil at
- 25 cm; one- bottom of slope over time
pass system;
some min till
Ploughed to 20 Erosion on slopes > 4;
cm; one-pass gully eroded in 2011 due
system
to mismanagement
No
Some straw
incorporation
P deficiency; cultivation
and rolling carried out
across the contour
WW = 9.9
2010 (36
cm)
Cheese sludge
Ploughed at 35 angle to WW = 11.112.4;
slope direction; rolled
WO = 8.0;
across slope; compaction
OSR = 5.2
on headlands/tramlines
Letterkenny, 54°59'55'' N,
Ploughed to 18
Erosion during heavy
2007 (36
FYM (2007 Prone to compaction and
WB = 8.9
7°32'57''W
Co. Donegal
cm; one-pass
rainfall in particular on
cm)
and 2008)
cracking; poor growth at
system
tramlines; flooding
bottom of slope due to
deposition of sand/clay
Duleek, Co. 53°38'57'' N, 1990- : planted with beans Ploughed to 18
No
No
Difficult to
Prone to clodding
SB = 4.9 – 6.2
6°23'14''W
Meath
or left fallow every 4 yr
cm; seed
access
(irregular blocks created
with SB in between
spreader
by artificial disturbance)
1
Lat/long, latitude and longitude; MC, moisture content; OSR, oilseed rape; BFSL, beet factory sludge lime; SW, spring wheat; WB, winter barley; SB, spring barley; WO,
winter oats; FYM, farmyard manure; WW, winter wheat;
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Chapter 5
5.2.3 Methods used to assess erosion risk and soil quality at the study sites
No erosion risk or soil quality assessment standard currently exists in Ireland.
Therefore, the methods of erosion risk (Defra, 2005), soil structure (VESS) and soil
quality (VSA) assessment developed in the UK and New Zealand in recent years were
applied to the study soils in order to develop a risk assessment for erosion in Ireland
by improving on the Defra (2005) erosion risk assessment. While the primary
objective was to develop an improved erosion risk assessment by incorporating the
elements of VESS and VSA that effect surface runoff generation and soil erosion into
Defra (2005), it was also envisaged that these newly added elements would give an
indication of soil structural quality, as that was their original function in the VESS and
VSA methods. The methods of soil structure and soil quality assessment are compared
with the Soil survey of Ireland soil classification method in Table 5.2 and advantages
and limitations of each method are outlined in Table 5.3.
Table 5.2 Methods used in visual and tactile assessment
Technique
Soil textural class
Size distribution
(fragmentation)
Soil unit
morphology
Biological activity
Soil Survey of Ireland
profile description
(depth = up to 155cm)
Visual soil assessment
(depth = top 20 cm)
Visual evaluation of
soil structure (depth =
top 25 cm)
Hand assessment
-
Hand assessment
Drop shatter test (photo
comparison)
Aggregates are assessed
for size, shape, porosity
and strength
-
Aggregates are assessed for
size, grade, shape,
consistence, porosity and
roots
Record size, quantity, and
orientation of roots
External porosity
Record quantity, size, shape,
orientation and continuity of
pores using a hand lens
Internal porosity
Check quantity, size shape,
orientation and continuity of
pores within aggregates
Assess soil colour and
mottling according to
Munsell soil colour charts
Colour
Water content
Assess drainage condition
Rooting
Potential rooting depth –
record quantity, size and
location of roots in each
horizon and look for
clustering, thickening and
deflection of roots
Record earthworm
number and species type,
soil smell, and roots
Examine an exposed soil
face for spaces, holes,
cracks and fissures
between aggregates and
clods (photos)
Check for pores within
clods and aggregates
Comparison with soil
colour under fenceline and
quantification of mottling
(photos)
Assessment of degree of
surface ponding (photos)
Potential rooting depth –
record overthickening or
forced horizontal growth
of roots, firmness and
tightness of soil and hard
pans (in top 80cm)
120
Aggregates are assessed
for size, shape, porosity,
roots and strength
(photos)
Check for anaerobic
zones and presence of
large worm holes
Check for abundance and
clustering of roots,
macropores, and cracks
between aggregates
Check for pores and roots
within aggregates
Grey-blue colour
recorded if present
Look for clustering,
thickening and deflection
of roots
Chapter 5
Table 5.3 Advantages and limitations of respective methods
Method
Soil Survey of
Ireland profile
description
Visual soil
assessment
Visual evaluation
of soil structure
Advantages
Limitations
 Soil assessed to full depth of profile
 Parent material and profile drainage are
determined
 Sensitive enough to detect slight changes in
soil structure
 Time consuming and labour intensive
 Replication difficult for reasons above
 Agronomists and farmers must be
trained in its use
 Not suitable for identifying effects of
cropping systems
 Only deals with top 20cm of profile
 No assessment of profile drainage or
parent material
 Special training or technical skills are not
required
 Objective method of assessment
 Can identify effects of cropping systems on
soil structure and rooting
 Links soil condition to plant performance
 Relatively quick
 Takes into account surface cover, erosion
and nutrient loss
 Special training or technical skills are not
required
 Objective method of assessment
 Can identify effects of cropping systems on
soil structure and rooting
 Low cost, rapid and flexible test
 Large numbers of replicates are possible
for statistical analysis
 Only deals with top 25cm of profile
(can be adapted to assess deeper layers)
 Potential rooting depth and profile
drainage not determined
 No assessment of erosion features
 Limited measurement of biological
activity
 No assessment of soil texture which is
important for erosion risk
 No size distribution of aggregates
5.2.3.1 Field-Scale Defra (2005) Assessment
The Defra (2005) method is divided into two stages: first, each field is assessed by the
farmer on the basis of soil texture and slope (Table 5.4) and a map of erosion risk is
produced for the farm; secondly, the cropping regime or land use is classified by
degree of erosion susceptibility. The method followed in determining the soil texture
in each of the study fields is outlined in Appendix D. The slope of each of the study
fields in the present study was determined using an inclinometer. The observation of
rills, gullies, or sediment transport at the study sites during assessment overrode the
classifications given in Table 5.4.
Additional factors used in the assessment to upgrade or downgrade a study site
included: soil structure; SOM; valley features which tend to concentrate runoff water;
long, unbroken slopes and very steep slopes (i.e. greater than 11 degrees). If a study
site was reported to flood, on average, once in every 3 years or more frequently, it was
deemed to be highly vulnerable. The land uses deemed by Defra (2005) to leave the
soil in the most erosion susceptible condition include: late sown winter cereals;
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potatoes; sugar beet; field vegetables; and grazed fodder crops. These land uses
should be avoided on very high or high erosion risk fields unless erosion control
measures are in place. Spring cereals are generally at lower risk from erosion than
winter cereals, as seedbeds are not exposed to winter rainfall.
Table 5.4 Water erosion risk-assessment (from Defra, 2005)
Soils
Steep slopes
Moderate slopes
Gentle slopes
Level ground
> 7
3 - 7
2 - 3
< 2
Sandy and light silty
Very high
High
Moderate
Lower
Medium and
High
Moderate
Lower
Lower
Lower
Lower
Lower
Lower
calcareous
Heavy
Very high = rills are likely to form in most years and gullies may develop in very wet periods; High = rills are likely to develop
in most seasons during wet periods; Moderate = sediment may be seen running to roads, ditches or watercourses and rills may
develop in some seasons during very wet periods; Lower = sediment rarely seen to move but polluting runoff may enter ditches
or watercourses.
In 2009, Defra published a new soil protection review (Defra, 2009b), the purpose of
which is to tackle degradation threats to soil. It explains about risks, and guides
farmers through the process of identifying the risks on their soils and how to address
them. It requires the farmer to produce a map of his land identifying the fields at risk
of degradation (erosion, compaction and SOM decline) based on measured soil
texture, observation of runoff, erosion, compaction, waterlogging and SOM decline,
and knowledge of the site conditions and locations of fields. Measurement of slope is
not required in the new soil protection review approach to soil risk mapping, which
attempts to classify the field in terms of its susceptibility to three degradation threats
as opposed to focusing on erosion as the Defra (2005) approach does. While there are
certainly merits to this new approach, it was deemed more appropriate to use the
Defra (2005) approach in the present study, as the focus was on identifying sites with
a high risk of erosion which is difficult if slope is omitted. Furthermore, Boardman et
al. (2009) showed that the Defra (2005) approach incorporating slope was 90%
successful at identifying high risk fields, whereas the effectiveness of the new
approach in which slope is omitted, is unknown. The suitability of the Defra (2005)
erosion risk assessment for use in Irish conditions is discussed below.
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Any proposed approach to erosion risk assessment on tillage land in Ireland should be
one that requires limited investment of resources and, if possible, results in higher
productivity through improved soil structure and quality. Ireland is similar to England
in that erosion is not widespread but occurs in many localised areas and therefore
requires a field-scale reconnaissance approach to identify risk areas where mitigation
measures are warranted. The application of the Defra (2005) scheme to tillage land in
Ireland would be a positive first step towards quantifying the actual erosion risk
associated with tillage soils here. Given that (1) more frequent, intense rainfall events
in summer are projected and (2) winter rainfalls are expected to increase by
approximately 10% in Ireland by 2050 as a result of climate change (Sweeney et al.,
2008), there is an increasing need for erosion risk assessments that can identify areas
where erosion is occurring or is likely to occur under future climatic conditions, with
a view to ultimately developing measures to mitigate the effects of erosion. Boardman
et al. (2009) note that in the long-term, risk assessment procedures will have to take
into account these predicted climate changes. The application of the Defra (2005)
scheme to Irish conditions should not pose great difficulties because of its simple
design - requiring only minimal input by farmers. Furthermore, the soil types,
topography, rainfall and land uses do not differ greatly between the two countries so
adjustments to suit Irish conditions should also be minimal. Its adoption in Ireland
would present an opportunity to improve the scheme by including a simple
assessment of soil structure and quality, conductible by the farmer. The need to assess
soil structure when determining erosion risk was highlighted in the Defra (2005)
scheme; however, no procedure for assessing it was provided. If soil structure and
quality assessments are included in the scheme, the farmers will be made more aware
of the links between poor soil structure and erosion, and good soil quality and
productivity. This could increase uptake by farmers by creating a „win-win‟ situation
in which the farmer is rewarded for better soil management through improved crop
yields while the local environment benefits from lowered erosion levels. This may
also reduce the reliance on monetary incentives to ensure high uptake of agrienvironmental schemes by farmers.
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5.2.3.2 Visual Soil Assessment
All VSAs were carried out in the present study when the soil was moist or slightly
moist. Consultation with the farmer and a visual inspection at each site identified
specific locations where a VSA should be conducted. At each location, the VSA was
carried out in triplicate using the equipment in Figure 5.1. Areas of the field thought
to be susceptible to erosion, such as tramlines and areas assumed to be less susceptible
because they were under crop and less compacted, were studied. The scorecard, tables
and visual aids presented in the VSA field guide of Shepherd (2009) and used during
this study to score indicators of soil quality, are summarised in Appendix E. The VSA
indicators of soil quality are texture, structure, porosity, number and colour of soil
mottles, colour, earthworms, smell, potential rooting depth (PRD), surface ponding,
surface cover and surface crusting and erosion.
Figure 5.1 Equipment used to carry out visual soil assessments.
Soil fragmentation and friability (the ability of a solid substance to be reduced to
smaller pieces with little effort) were examined following a drop shatter test
(Shepherd, 2009). In this test, a 200 mm cube of topsoil was dropped a maximum of 3
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times from a height of 1m (for sandy loam soils the sample was dropped once from a
height of 0.5 m) onto a firm board. The aggregates were then ordered from largest to
smallest on a wooden board for comparison with sample photographs before scoring
the soil structure in the following manner: 2 = good condition, 1 = moderate condition
and 0 = poor condition (Appendix E - soil structure). For the porosity score, a spade
slice of soil was removed from the hole and broken in half so that the fresh horizontal
face could be compared with sample photographs (Appendix E - soil porosity). The
other VSA indicators of soil quality were scored according to the tables and visual
aids provided in Appendix E. On completion, each visual indicator of soil quality was
multiplied by its designated weighting and the products were added to give the soil
quality index (Appendix E - scorecard).
5.2.3.3 Visual Evaluation of Soil Structure
The locations identified for VSA were also assessed using VESS. A block of soil with
a depth of 25 cm, a length of 20 cm and a breath of 20 cm, was first extracted using a
spade and placed on a tray for examination. The block was examined for horizontal
layering (layers of differing structure) and, if present, each layer was scored
separately. The block was then broken up by hand to reveal any cohesive layers or
clumps of aggregates, and to separate the soil into natural aggregates and man-made
clods. The major aggregates were then broken apart until a piece of aggregate, 1.5-2.0
cm in size, remained. These aggregates were then assessed with regard to their
structural quality as outlined in Appendix F. Soil structure quality (Sq) was scored
between Sq 1 (friable) and Sq 5 (very compact). As is recommended in the VESS
method, scores were cross-checked regularly between two assessors. The scores were
confirmed by checking for factors that increase the score, such as larger, more
angular, less porous aggregates; and clustering, thickening and deflection in roots.
5.2.3.4 Soil Survey of Ireland assessment
The methods used in the Soil Survey of Ireland to formally classify soils according to
the United States Soil Taxonomy (Soil Survey Staff, 1975 and 1993) classification
systems and to evaluate the suitability of land for specific agricultural enterprises and
125
Chapter 5
crops were applied to each of the 6 study sites. Soil augering was first carried out to
ensure that the selected pit location was as representative as possible of the field as a
whole. A pit was then excavated at each of the 6 field sites so that a fresh vertical soil
face (no smearing) was achieved to the required depth (0.7 - 1.55 m). A detailed
description of the general characteristics (drainage condition and pattern of horizon
development) of the profile was carried out. Properties of individual soil horizons
such as texture, structure, porosity, consistence, colour (assessed using a Munsell
colour chart (Munsell, 2009)), mottling, amount of SOM, stoniness, presence of
hardpans and root development, were described. Other less important criteria assessed
were the formation of saturated zones, changes in soil moisture with depth, anaerobic
zones and the pattern of roots. A bulk sample was taken from each soil horizon for
physical and chemical analysis.
The soil properties measured in each soil horizon as part of the morphological
description were: (1) pH (1:1 soil/solution ratio); (2) PSD by sieve and pipette
analysis; (3) CEC; Bascomb, 1964; (4) SOC determined by sulfochromic oxidation
(ISO, 1998; ISO 14235); and (5) Free iron (Fe) (hydro)oxides using the dithionitecitrate-bicarbonate method. Measurement of SOM was by loss on ignition at 550°C
after Byrne (1979). Only after the analytical data became available were the soils
formally classified by a soil scientist. The full classification for each site included
determination of soil type, soil series (where available) and parent material.
The soil properties described and measured using the Soil Survey of Ireland method
can also be used to estimate the erodibility of the study soils. Some of these properties
influence the soils capacity to infiltrate rain, and therefore help determine amount and
rate of runoff; some influence its capacity to resist detachment and transport by
rainfall and overland flow, and thereby determine the soil content of the runoff
(Wischmeier and Mannering, 1969). The methodology used to determine the study
soils erodibility using the soil-erodibility nomograph of Wischmeier and Smith
(1978): K (ton ha h ha-1 MJ-1 mm-1) = ((2.1 x 10-4 M1.14(12 - OM) + 3.25(s-2) + 2.5(p 3))/100)*0.1317, which ranks the soils erodibility based on the % OM, soil structure
class (s), permeability class (p) and textural factor (M: percentage silt (Si) + fine sand
fraction (fSa) content multiplied by 100 - clay fraction (Cl)) is outlined in Figure 5.2.
126
Chapter 5
Soil survey data
K-factor properties
% Organic carbon
% Organic matter (OM)
K-factor determination
Soil erodibility classes
Low erodibility
1.High clay soils have low K values (about
Soil structure class (s)
0.007-0.02) because of better cohesion
Description of
1-very fine granular
making them more resistant to detachment
structure in the
2-fine granular
by water.
plough horizon
3-moderate or coarse
granular
2.Coarse textured soils also have low K
K-factor equation
4-blocky, platy or massive
K = ((2.1 x 10-4 M1.14(12 - OM) +
Profile description
of texture, colour,
mottling,
structure, porosity
and hardpans
Profile permeability class (p)
3.25(s-2) + 2.5(p - 3))/100)*0.1317
1-rapid
(Wischmeier and Smith, 1978)
2-moderate to rapid
values (0.007-0.026) because of low
runoff potential even though they are
easily detached by water.
Moderate erodibility
Medium textured soils like loams have
3-moderate
4-slow to moderate
K is expressed in SI units of ton ha
5-slow
h ha-1 MJ-1mm-1
moderate K values (0.033-0.053) because they
are moderately susceptible to detachment and
they produce moderate runoff.
6-very slow
Particle size
High erodibility
analysis – silt (Si),
Textural factor (M)
clay (Cl) and fine
M = (% Si + fSa)(100 - % Cl)
Soils having high silt and fine sand content have
high K values (> 0.053) because they are easily
sand (fSa)
detached; tend to crust and are prone to runoff.
Figure 5.2 Methodology for determining soil erodibility using Soil Survey of Ireland data.
127
Chapter 5
5.3 Experimental results from erosion risk and soil quality assessments
The results from the erosion risk and soil quality assessments conducted at the study
sites are presented here. These data provided information on the likelihood of erosion
and the loss of soil quality across the study sites.
5.3.1 Field-Scale Defra Assessment
The application of the Defra (2005) erosion risk assessment to the 6 study sites
(Table 5.5) identified three of the sites as having high or very high erosion risk based
on soil texture (determined by hand as outlined in Appendix D) and slope (determined
using an inclinometer). According to the assessment, rills are likely to form in most
years and gullies may develop in very wet periods in fields classed as very high risk.
In high risk areas, rills are likely to develop in most seasons during wet periods. The
Bunclody and Duleek fields were less erodible due to the clay loam texture of their
soils which makes them more resistant to detachment by rainfall and runoff than the
sandy loam textured soils of the Tullow, Fermoy, Clonmel, and Letterkenny fields.
The erosion risk of the Tullow field was classed as moderate due to a low slope of 2
degrees. The Duleek field was also classed as a moderate risk due to a moderate slope
of 4 degrees and a medium soil texture of clay loam. According to Defra (2005),
sediment may be seen running to roads, ditches or watercourses, and rills may develop
in some seasons during very wet periods in fields classed as moderate risk. The
Fermoy site was classed as having very high risk of soil erosion due to a combination
of erosion-susceptible soil and steep slope. Only the Letterkenny site had a high land
use risk (Table 5.5) because of the planting of potatoes in the rotation, which is
highlighted in the Defra (2005) assessment as a highly erosion-susceptible land use.
The planting of potatoes at the very high risk Letterkenny site should be avoided in
future unless precautions are taken to control erosion.
Based on the erosion features observed in the fields during the study period (Table
5.5) and the anecdotal evidence provided by the farmers (Table 5.5), the Defra (2005)
assessment was effective in identifying fields where more severe erosion is likely to
occur, with the exception of the Fermoy field in which the assigned risk class was too
128
Chapter 5
Table 5.5 Results of the Defra (2005) erosion risk assessment
Study site
Annual
Soil
rainfall
texture
mm
Bunclody
1038
Clay loam
Slope
Slope
Erosion history of sites
Observed erosion
Length

m
2
120
Land use rotation
features
Sheet erosion on longer slopes
Defra erosion
Defra land
risk
use risk
None
OSR, WB, SB, SW
Lower
Moderate
None
OSR, WO, WW
Moderate
Moderate
Greater depth of soil in
Greater depth of soil in
WB, WW, SB
Very high
Moderate
footslope areas
footslope areas
Erosion on slopes > 4; gully
Rilling; gully; runoff in
WW, WO, OSR
High
Moderate
eroded in 2011
tramlines
WB, potatoes
High
High
SB, beans, fallow
Moderate
Moderate
during prolonged heavy rainfall
Tullow
838
Sandy
2
200
loam
Fermoy
1078
Sandy
during prolonged heavy rainfall
9
100
loam
Clonmel
1245
Sandy
3
200
loam
Letterkenny
1097
Sandy
Sheet erosion on longer slopes
5
75
loam
Erosion during heavy rainfall in Rilling; runoff and erosion
particular on tramlines; flooding in tramlines; sand and silt
deposits at field ending
Duleek
815
Clay loam
4
120
No past erosion
None
SB, Spring barley; SW, Spring wheat; WO, Winter oats; Winter barley; WW, Winter wheat; OSR, Oilseed rape
129
Chapter 5
high when compared with the observed erosion levels. The Fermoy site was classed as
very high risk and therefore rilling and even gullying should have been observed;
however, no erosion features were observed during the study or reported by the
farmer with the exception of a greater depth of soil measured in footslope areas
compared to the crest of the hill. In contrast, the Letterkenny and Clonmel fields were
classed as high risk and were the only fields in which rilling and sediment deposits
were observed. Further assessment of the Fermoy field is required to determine what
factors (that can be easily assessed by the farmer), outside of slope and texture can be
used to help explain the observed low erosion levels. Once identified, these factors
should be included in the assessment so that more accurate determination of erosion
risk on fields such the Fermoy field is possible. It is recommended in the Defra (2005)
assessment to upgrade or downgrade a field‟s erosion risk based on the following
factors: soil structure; SOM; valley features which tend to concentrate runoff water;
long, unbroken slopes and very steep slopes (i.e. greater than 11 degrees). However,
no clear method is outlined for the farmer with which he/she can assess soil structure
and upgrade or downgrade the risk class accordingly. These and other factors
influencing soil erosion are investigated in this section using simple farmer friendly
methods that are suitable for inclusion in an improved erosion risk assessment method
for tillage soils in Ireland.
5.3.2 Visual Soil Assessment
Results of VSAs conducted at the 6 study sites are presented in Table 5.6. The
cropped areas of the Bunclody, Fermoy and Tullow sites were shown to be in good
condition (denoted by VSA structure scores of 2; Table 5.6), dominated by friable,
fine aggregates with no significant clodding (irregular blocks created by artificial
disturbance; e.g., tillage or compaction - Figure 5.3). Tramline areas of these sites
were shown to be in moderate-to-moderately good condition (denoted by VSA
structure scores of between 1 and 1.75) and contained a mix of coarse clods and
friable fine aggregates (Figure 5.3). Soil structure is vulnerable to change by
compaction and erosion, and its preservation is key to sustaining soil function
(Mueller et al., 2010). Tramline areas of the Clonmel and Letterkenny soils were in
poor condition (denoted by VSA structure scores of 0), and were dominated by coarse
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Chapter 5
Table 5.6 Results of Visual Soil Assessments
Visual indicators of soil quality and respective weightings in brackets
Study site Land use Texture Structure Porosity Mottling Colour Earthworms Smell
(3)
(3)
(3)
(2)
(2)
(3)
(2)
Potential
Ponding
Cover and
Erosion
Visual
Soil Quality
rooting
(3)
crusting
(1)
score
Index2
depth (3)
Bunclody
Tullow
Fermoy
Clonmel
ranking1
(2)
OSR
1.5
2
1.5
2
2
0
1
2
2
2
2
43
Good
Tramline
1.5
1.5
0.5
2
1.5
0
1
0
1
0
1
23.5
Moderate
OSR
1
2
1.5
2
2
0
1
2
2
2
2
41.5
Good
Tramline
1
1.5
0.5
2
2
0
1
0
1
0
1
23
Moderate
WB
1
2
2
2
1.5
0.5
1
2
2
2
2
43.5
Good
Tramline
1
1.75
1
2
1.5
0
1
0
1
0
1
24
Moderate
WW
1
1.5
0.78
2
1
0.5
1
0.5
1
1
1
28
Moderate
Tramline
1
0
0
1.5
1
0
1
0
0
0
0
10
Poor
WB
1
1.5
1.5
1.5
1
0
1
0.5
0.5
1
1
25
Moderate
Tramline
1
0
0
1.5
1
0
1
0
0
0
0
10
Poor
SB
1.5
1
1.5
1.5
1
0.5
1
2
2
0.5
2
35.5
Moderate
Tramline
1.5
1
1
1.5
1
0
1
0
0.5
0
1
20
Moderate
Letterkenny
Duleek
SB, Spring barley; WB, Winter barley; WW, Winter wheat; OSR, Oilseed rape
1
The score for each visual indicator of soil quality was multiplied by its designated weighting and the products were added to give the visual score ranking, after Shepherd
(2009).
2
Good = > 37; Moderate = 20-37; Poor = < 20. See Appendix E for more details.
131
Chapter 5
Bunclody
UC (VS = 2)
Tullow
UTL (VS = 1.5)
UC (VS = 2)
Fermoy
UC (VS = 2)
Clonmel
UTL (VS = 1.75)
UC (VS = 1.5)
Letterkenny
UC (VS = 1.5)
UTL (VS = 1.5)
UTL (VS = 0)
Duleek
UTL (VS = 0)
UC (VS = 1.0)
UTL (VS = 1.0)
UC, Undercrop; UTL, Under tramline; VS, Visual scoring
Figure 5.3 Visual scoring of soil structure during visual soil assessments at the 6
study sites.
132
Chapter 5
clods with very few finer aggregates. The coarse clods were very firm, angular or subangular in shape, and had very few or no pores (Figure 5.3). The presence of these
larger aggregates reduces intra-aggregate porosity (Shepherd, 2009), thereby lowering
soil permeability. Visual soil assessment structure scores identified poorer soil
structure in cropped areas at the Letterkenny, Clonmel and Duleek sites when
compared to the other sites.
The poorer structure at the Duleek site was most likely due to tillage operations being
conducted when the soil was drier than the optimum water content. Tillage conducted
when the soil is wetter or drier than this optimum, results in the production of a
greater number of large clods that must be broken down in one or more subsequent
tillage operations (Dexter and Birkas, 2004). The Duleek site had the lowest annual
rainfall (Table 5.5) of all the sites and was subject to drought in the past. As a result,
tillage operations may have to be conducted when the soil is drier than the optimum.
In contrast, the Clonmel and Letterkenny sites received high annual rainfall and were
subject to periodic flooding, which may have resulted in tillage operations being
conducted when the soil was wetter than the optimum. The lower degree of
compaction in the tramlines of the Duleek site, when compared with the cropped area
of the Duleek site and tramline areas in other sites, was likely due to the drier
condition of the soil during the course of the year. Vehicle traffic operations
conducted when soil moisture is high result in greater soil compaction. Ideally, soils
should only be trafficked upon when soil moisture conditions are < 60% of field
capacity (Raper, 2005). However, this is generally not possible under Irish weather
conditions and tramline compaction on sandy soils such as Clonmel and Letterkenny
often results. The susceptibility of these soils to compaction is further highlighted by
the moderately poor PRD scores obtained in the cropped area of each soil (Table 5.6).
The apparent resistance of the Bunclody and Tullow soils to compaction in tramline
areas could be due to their better soil structure in cropped areas. Good soil structure
reduces the susceptibility to compaction under wheel traffic (Shepherd, 2009).
The soil quality index determined using visual scores of indicators of soil quality is in
close agreement with VSA structure scores (Table 5.6). This is to be expected given
the strong influence soil structure has on soil quality indicators like porosity, mottling,
133
Chapter 5
ponding, PRD and erosion. Soil structure is known to be a crucial criterion of
agricultural soil quality. The soil quality at the Bunclody, Tullow, and Fermoy sites
was good in cropped areas and moderate in tramline areas where some compaction
had occurred. Soil quality was only moderate in the cropped areas and poor in the
tramline areas of the Clonmel and Letterkenny sites, which may partly explain the
observed higher erosion levels at these sites.
5.3.3 Visual Evaluation of Soil Structure
Visual evaluation of soil structure scores for the study soils ranged from 1 - 2 for the
respective land use being practiced on each soil and from 3 - 5 in tramline areas and
are presented in Table 5.7. These scores reflect the surface compaction that has taken
place in tramline areas as a result of vehicular traffic during spraying and fertilising
operations. This affects the structural quality of the soil by changing it from friable or
intact under crop to firm, compact or very compact in tramlines areas. The reduction
in soil porosity associated with this compaction of the surface soil can significantly
reduce its permeability in these areas, thereby increasing the risk of surface runoff
being generated during wet periods. This risk was particularly high in tramline areas
of the Letterkenny and Clonmel fields as these soils had the poorest structural quality,
with both being classed as very compact. Observed surface runoff from tramline areas
during wet weather at the Clonmel and Letterkenny sites allied with the absence of
runoff from cropped areas shows that soil structure quality, as measured by VESS,
can be indicative of areas where runoff risk is high. Unless traffic is eliminated, good
timing of operations is the most effective way to preserve soil structural quality (Ball
et al., 1997). Surface compaction can be alleviated in the next cultivation cycle.
134
Chapter 5
Table 5.7 Results of Visual Evaluation of Soil Structure
Study site
Bunclody
Tullow
Fermoy
Clonmel
Letterkenny
Duleek
Land use
Structural Quality Score
Structural Quality
OSR
1
Friable
Tramline
3
Firm
OSR
1
Friable
Tramline
3
Firm
WB
1
Friable
Tramline
3
Firm
WW
2
Intact
Tramline
5
Very Compact
WB
2
Intact
Tramline
5
Very Compact
SB
2
Intact
Tramline
3
Firm
SB, Spring barley; WB, Winter barley; WW, Winter wheat; OSR, Oilseed rape; Structural quality as
determined by VESS (see Appendix F); Scores of 1 - 3 are usually acceptable, whereas scores of 4 or 5
require a change of management.
5.3.4 Soil Survey of Ireland assessment
Detailed results of the soil profile examinations carried out to classify the study soils
according to the Soil Survey of Ireland method can be found in Appendix G. Profile
descriptions carried out to the full depth of each study soil are provided in Appendix
G (Table G.1 and Figure G.1). Soils containing high proportions of silt and very fine
sand, identified by the presence of small pockets or thin layers of pale sand grains
within the cultivated layer or concentrated at its base (Batey, 2000), are most erodible
(Smolen et al., 1988). Thin layers of pale sand grains were identified in the cultivated
layer of the Letterkenny soil (Appendix G, Figure G.2). These are the result of the
disintegration of aggregates into their component particles (Batey, 2000) which, due
to their lack of cohesion and low settling velocity, are easily eroded. The observation
in the Letterkenny field of deposits of fine sand particles at the bottom of sloping
tramlines (Appendix G, Figure G.3) was further evidence of the disintegration of
aggregates and the transport of their component particles by concentrated water flow
in compacted tramlines.
135
Chapter 5
The PSD results for the Letterkenny soil (Appendix G - Table G.2) confirmed the
findings of the field analysis. Having 60% fine sand within its sand fraction, the
Letterkenny topsoil was the only topsoil to be classed as a fine sandy loam, with the
other study soils falling in the medium and coarse sandy loam and loam classes.
Medium and coarse sandy loams are normally less susceptible to erosion than fine
sandy loams, as they are very fast draining and contain larger particles with higher
settling velocities. The soils tested in this study, which had topsoil textures ranging
from loam to sandy loam and clay contents between 10 and 21% (Table 5.8), fell
within the range of soils susceptible to structural decline and erosion.
In general, permeability decreases as subsoil colour changes from red to brown to
yellow to dark (black to very dark brown) to grey (light grey to bluish and greenish
greys) (Moore, 1998). Based on the Munsell colour description of each horizon for the
soils tested, the Letterkenny soil had by far the worst drainage class of all the soils,
going from moderately well-drained (yellow-brown) in the topsoil to near permanent
waterlogging (grey/olive) in the lower horizons (Appendix G, Table G.1). This was
indicative of a substantial risk of overland flow due to saturation excess.
The lack of mottles in the horizons of the Bunclody, Tullow, Fermoy and Duleek soils
indicated that they had good drainage and were unlikely to generate overland flow
when managed properly. The presence of few-to-many mottles in the lower horizons
of the Clonmel soil indicated that the soil was moderately well-drained and may be
susceptible to infiltration excess overland flow when subjected to high intensity
rainfall. The presence of many medium, distinct olive-to-dark olive 5Y 5/8 mottles in
the B21g horizon of the Letterkenny soil (Appendix G, Figure G.4) may be indicative
of prolonged waterlogging. Care should be taken when using mottle patterns to assess
the wetness class of a soil profile, as they can persist long after the condition
responsible for their formation is gone (Batey, 2000).
Stable surface soil structure is important for facilitating rapid water infiltration,
controlling soil erosion, and reducing water runoff of soil contaminants to nearby
surface waters (Franzluebbers, 2002). The surface horizon of each soil was friable
(except the Clonmel soil, which was firm) in consistence and had a structural mix of
136
Chapter 5
crumb in the upper few cm grading to fine and medium sub-angular blocky. The
plough horizon of the Clonmel soil had the poorest structure of the 6 plough horizons.
The lower horizons of the Clonmel and Letterkenny soils had structures normally
associated with poor drainage and waterlogging. Hard pans such as those identified in
the Clonmel and Letterkenny soils impede the movement of water through the soil
profile, increasing the susceptibility to waterlogging and erosion by rilling and sheet
wash (Shepherd, 2009). Based solely on soil structure, the Letterkenny and Clonmel
soils had a higher associated erosion risk due the greater risk of runoff being
generated on these soils.
The SOM level in the plough horizon of the study soils was above 3.4% (soils with
SOM levels above this threshold are considered not to be vulnerable or depleted
(DAFF, 2009a)) with the exception of the Letterkenny soil, where it was only 3.23%
(ca. 1.9% SOC) (Appendix G, Table G.2). These low values of SOC are not unusual
for sandy tillage soils. Based solely on the SOC levels measured in the plough horizon
of the study soils, the Letterkenny, Tullow and Clonmel soils posed the greatest
erosion risk.
Both the Clonmel and Letterkenny soils were classified as Gleys by a soil scientist
(Appendix G - Table G.3) based on the results of the profile examinations. Gleys are
soils in which the effects of drainage impedance dominate and which have developed
under conditions of permanent or intermittent water-logging (Finch et al., 1983). In
general, Gley soils have weak structure and poor drainage which makes them difficult
to cultivate. The potential for surface runoff due to saturation excess is higher in these
soils than in the better drained Brown Earth (Bunclody, Fermoy and Duleek) and
Grey Brown Podzolic (Fermoy) soil types.
In Table 5.8, the soil properties determined analytically or estimated by visual and
tactile assessment as part of the soil classification procedure are used to produce Kfactor values for the cropped and tramline areas of the study sites. The K-factor values
for the Letterkenny soil identified it as having the highest erodibility of all the soils,
which was in agreement with the observed erosion across the study sites. The Fermoy
soil, when cropped, had the lowest erodibility as estimated by the K-factor. The low
137
Chapter 5
Table 5.8 K-factor erodibility values and associated risk.
Study site
Bunclody
Tullow
Fermoy
Clonmel
Letterkenny
Duleek
FSa + silt
Clay
SOM
%
%
%
57
21
4.81
53
14
3.6
56
14
4.08
54
16
3.57
69
10
3.23
59
20
4.08
Land use
Permeability
Structure
K-factor
-1
Erodibility Risk
-1
ton ha h ha MJ mm
-1
OSR
Mod. to rapid (2)
Fine crumb (2)
0.0258
Low to moderate
Tramline
Slow (5)
Med/coarse crumb (3)
0.0399
Moderate
OSR
Mod. to rapid (2)
Fine crumb (2)
0.0311
Low to moderate
Tramline
Slow to mod. (4)
Med/coarse crumb (3)
0.042
Moderate
WB
Rapid (1)
Very fine crumb (1)
0.0237
Low
Tramline
Slow to mod. (4)
Med/coarse crumb (3)
0.0422
Moderate
WW
Slow (5)
Med/coarse crumb (3)
0.0452
Moderate
Tramline
Very slow (6)
Massive (4)
0.0528
Moderate to high
WB
Slow (5)
Med/coarse crumb (3)
0.0620
High
Tramline
Very slow (6)
Massive (4)
0.0696
High
SB
Mod. to rapid (2)
Med/coarse crumb (3)
0.0348
Moderate
Tramline
Slow to mod. (4)
Med/coarse crumb (3)
0.0414
Moderate
FSa, fine sand; SB, Spring barley; WB, Winter barley; WW, Winter wheat; OSR, Oilseed rape; SOM, Soil organic matter;
Low risk = low runoff potential (K-factor is 0.007-0.02 (high clay), 0.007-0.026 (coarse texture)); Moderate risk = moderate potential to produce runoff and moderately
susceptible to detachment (K-factor is 0.033-0.053 (medium texture)); High risk = easily detached, tend to crust and prone to runoff (K-factor is >0.053 (silt/fine sand)).
138
Chapter 5
K-factor in this case was largely due to the rapid permeability of the Fermoy soil
profile and the fine crumb structure of its plough layer, both of which lower its
potential to produce runoff.
5.4 An erosion and soil quality screening toolkit for Irish tillage soils
Integration of erosion risk indicators (Defra, 2005) and soil quality indicators with
observed erosion levels in the study soils (Table 5.9) allowed the selection of a
strategic set of indicators for inclusion in a screening toolkit designed to assess fields
for likelihood of erosion and loss of soil quality (Table 5.10). It is proposed that the
specific data required to support indicators deemed appropriate for inclusion in the
screening toolkit be collected by the farmer and an erosion risk class assigned (Figure
5.4 and Figure 5.5). Sites identified as high and very high risk can then be further
assessed by specialist advisors/consultants. The justification for inclusion of each
indicator in the screening toolkit is now provided.
5.4.1 Development of the screening toolkit
The Defra (2005) erosion risk assessment utilises field slope and soil texture primarily
to determine a field‟s erosion risk class. Certain visual indicators of soil quality such
as structure, ponding, and porosity, are also good indicators of erosion risk and were
therefore used in conjunction with SOM and AAR values to upgrade or downgrade
the erosion risk class generated using the Defra assessment (Figure 5.4 and Figure
5.5). Visual indicators receiving scores of 2 were in good condition and indicated that
the risk of soil erosion occurring was lower and hence the erosion risk class was
downgraded. In contrast, visual indicators receiving scores of less than 1 were in poor
condition and therefore required upgrading to a higher risk class. Soil structure is the
critical parameter here and its appraisal was supported by assessment of porosity,
ponding, SOM and AAR. Soil structure was weighted twice as high as the other
parameters used in the methodology for upgrading/downgrading the Defra risk class
(Figure 5.5).
139
Chapter 5
Table 5.9 Integration of Defra (2005) erosion risk rankings with selected soil quality indicators
Study site Defra (2005)
1
AAR
2
SOM
erosion risk
(mm)
(%)
Lower
1038
4.81
Tullow
Moderate
838
3.6
Fermoy
Very high
1078
4.08
Clonmel
High
1245
3.57
Bunclody
Letterkenny
Duleek
High
Moderate
1097
815
Land use
VSA
VESS
Structure
VSA
VSA
K-factor
Observed erosion features
Porosity Ponding Erosion Erodibility Risk
OSR
2
1
1.5
2
2
Low to moderate
None
Tramline
1.5
3
0.5
1
1
Moderate
None
OSR
2
1
1.5
2
2
Low to moderate
None
Tramline
1.5
3
0.5
1
1
Moderate
None
WB
2
1
2
2
2
Low
Greater depth of soil in footslope areas
Tramline
1.75
3
1
1
1
Moderate
Greater depth of soil in footslope areas
WW
1.5
2
0.78
1
1
Moderate
Rills (shallow)
Tramline
0
5
0
0
0
Moderate to high
Rills (moderate), small gully, runoff
WB
1.5
2
1.5
0.5
1
High
Rills (deep), fine sand and silt deposits
Tramline
0
5
0
0
0
High
Rills (deep), runoff carrying sediment,
SB
1
2
1.5
2
2
Moderate
None
Tramline
1
3
1
0.5
1
Moderate
None
3.23
4.08
SB, Spring barley; WB, Winter barley; WW, Winter wheat; OSR, Oilseed rape
1
VSA
AAR, average annual rainfall; 2SOM, soil organic matter
140
Chapter 5
Soil structure determines the porosity, strength and stability of a soil. Improved soil
structure enhances nutrient recycling, water availability and biodiversity while
reducing water and wind erosion, and improving surface and ground water quality
(Bronick and Lal, 2005). Effective erosion control therefore requires the maintenance
of good soil structure. The Defra (2005) erosion risk assessment recommends that
farmers assess their soil‟s structure when assigning an erosion risk class but does not
provide a method with which to conduct the assessment. The methods of soil structure
assessment used in VESS and VSA are proven methods with which the farmer can
score a soil‟s structure. In the present study, poor soil structure scores of 0 (VSA
structure score) and 5 (VESS score = very compact) in the tramlines of the
Letterkenny and Clonmel fields coincided with the observation of erosion features
such as rills and surface runoff carrying eroded soil (Table 5.9). Deposits of fine sand
and silt observed at tramline endings in the Letterkenny field are further evidence of
the effect of poor soil structure on erosion levels. Study field areas receiving good
soil structure scores of 2 (VSA structure score) and 1 (VESS score = friable) were
free of erosion features (Table 5.9) and were not seen to generate surface runoff at
times when surface runoff was recorded in adjacent areas, which had received worse
structure scores. These findings support the inclusion of an assessment of soil
structure, either by VESS or by VSA, in a screening toolkit (an improved version of
the Defra (2005) erosion risk assessment).
In general, VESS scores showed good agreement with VSA structure scores.
However, it is recommended that the farmer assess soil structure using the drop
shatter test from the VSA method rather than breaking the soil by hand as is done in
the VESS method, because the former is less subjective and more effective at
demonstrating to the farmer the effect management decisions have on soil structure.
Guimarães et al. (2011) compared the two methods and concluded that breaking the
soil by hand, if done as recommended, gives a result comparable to the method of
break-up by dropping. However the authors also noted that the user might need some
training in breaking up the aggregates along their natural failure planes (boundaries).
They found that this was generally not an issue after dropping the soil (as is required
in the VSA method) because of well-defined crack lines formed on impact.
141
Chapter 5
Table 5.10 Proposed toolkit indicators for use in screening sites for erosion risk and reduced soil quality.
Screening indicator
Proposed assessment method
Interpretation of results
Soil texture
Defra (2005) hand texturing method
Heavy soils = lower erodibility; medium soils = moderately erodible; sandy and light silty soils = highly erodible.
Slope angle
Measure slope as accurately as possible
< 2 = level ground; 2 – 3 = gentle slope; 3 – 7 = moderate slope; > 7 = steep slope; > 11 = very steep slope.
using a clinometer; record position in
landscape
Erosion features
VSA soil erosion (Shepherd, 2009)
Observation of rilling or gullying is used to upgrade sites classed as lower and moderate risk to high risk,
Soil structure
VSA drop shatter test (Shepherd, 2009)
2 (soil dominated by friable, fine aggregates with no significant clodding) = good;
1 (soil contains significant proportion (50%) of both coarse clods and friable fine aggregates) = moderate;
0 (soil dominated by coarse clods with very few finer aggregate) = poor.
Ponding
VSA ponding (Shepherd, 2009)
2 (no surface ponding of water evident after 1 day following heavy rainfall on soils that were at, or near, saturation) = good;
1 (moderate surface ponding occurs for 2 days after heavy rainfall on soils that were at, or near, saturation) = moderate;
0 (significant surface ponding occurs for 4 days or more after heavy rainfall on soils that were at, or near, saturation) = poor.
Porosity
VSA porosity (Shepherd, 2009)
2 (many macropores and coarse micropores between and within aggregates associated with good soil structure) = good;
1 (fewer macropores and coarse micropores between and within aggregates and moderate consolidation evident) = moderate;
0 (no soil macropores and coarse micropores are visually apparent within compact massive structureless clods) = poor.
% SOM
1
% loss on ignition
< 3.4% (ca. 2% SOC) = soil may be vulnerable or depleted;
> 3.4% soil considered not to be vulnerable or depleted.
AAR
2
Land use
AAR amount (low, moderate, high)
< 850 mm = low rainfall risk ; 850 – 1200 mm = moderate rainfall risk; > 1200 mm = high rainfall risk
Defra (2005) land use risk categories
Avoid erosion susceptible land uses such as late sown winter cereals, potatoes, and field vegetables, on very high or high risk
sites unless precautions are taken to limit erosion. Land uses such as early sown winter cereals, OSR 3 and spring sown cereals,
can be carried out with care on these sites.
1
2
3
SOM, soil organic matter; AAR, average annual rainfall; OSR, oilseed rape
142
Chapter 5
Given that the permeability of soil during infiltration is mainly controlled by coarse
pores (Beven and Germann, 1982), which can be seen by the naked eye, the
assessment of porosity using the VSA approach can give an indication of the study
soil‟s ability to transmit water away from the soil surface. Soils with high porosity are
therefore less likely to generate overland flow and associated erosion than are soils
with low-to-moderate porosity. Furthermore, subsoils with low-to-moderate porosity
are likely to have problems with waterlogging (Moore, 1998). Study soils receiving
VSA porosity scores of < 1 (indicative of low porosity) were also the soils in which
overland flow and rilling were observed (Table 5.9). For this reason, porosity was
deemed suitable for inclusion in the screening toolkit.
Soil organic matter is an important indicator of soil quality and productivity and is
important in influencing soil erodibility (Jankauskas et al., 2007). Furthermore, a
reduction in SOM levels has been highlighted in numerous legislative reports and
scientific papers as contributing to a decline in soil quality, and can result in increased
soil erosion. Increasing SOM content improves the cohesiveness of the soil, reduces
the risk of surface crusting, lowers the risk of soil compaction, increases its water
holding capacity and promotes soil aggregate formation, thereby improving structural
stability and reducing erosion. The primary role of SOM in reducing soil erodibility is
by stabilising the surface aggregates (Stott et al., 1999). Poor stability in surface
aggregates can predispose the soil to: disaggregation under raindrop impact and a
subsequent development of a surface crust, reduction of infiltration rate, and surface
runoff (Quinton and Catt, 2004). Guerra (1994) found that soils on eroded sites in
England had OM contents of 2.1 - 3.8%. In the present study, OM contents ranged
from 3.23 to 4.81% and as such were mostly higher than the range of OM contents
measured at eroded sites by Guerra (1994). The lowest levels of SOM recorded across
the study sites were measured at the Letterkenny (3.23%) and Clonmel (3.57%) sites,
and may in part account for the higher levels of erosion observed at these sites (Table
5.9) when compared to the other study sites. Soil organic matter was deemed to be an
important parameter when screening fields for likelihood of erosion because: (1)
previous studies by Kay and Angers (1999) and Whitmore et al. (2004) have
demonstrated its importance in maintaining soil structural stability and (2) it has been
shown to be one of the crucial factors in determining a soil‟s inherent erodibility. The
143
Flow Chart 1
Chapter 5
Texture assessment
Input texture here
Heavy
Slope assessment
Defra risk class
Risk adjustment
Adjusted risk class
Input slope here
0 - 11
Lower
> 11
High
If rills are observed in
the field in most years or
Medium
0 – 3
Lower
if the field is known to
3 – 7
Moderate
flood at least 1 in 3
> 7
High
years, then upgrade risk
Input risk class here
class (lower or moderate)
to high risk.
0 – 2
Lower
Go to Flow Chart 2
Light sandy
2 – 3
Moderate
to refine risk class
and silty
3 – 7
High
> 7
Very high
Figure 5.4 Methodology for determining erosion risk (adapted from Defra (2005)).
144
using soil quality
indicators and
annual rainfall.
Chapter 5
Flow Chart 2
Modified grading scores
Soil quality scoring
Soil structure
Input score here
2
+2
< 2 and ≥ 1
0
<1
-2
Upgrading/downgrading risk class
When all input
Input here
boxes to the left are
complete, add their
contents to give the
+
Ponding
2
Input score here
+1
< 2 and ≥ 1
0
<1
-1
soil quality/rainfall
adjustment score.
Input here
+
Porosity
2
Input score here
+1
< 2 and ≥ 1
0
<1
-1
Input here
+
> 3.4%
+1
< 3.4%
-1
Input here
Rainfall assessment
Average annual rainfall
Input rainfall here
Input score here
risk from side 1
(e.g. high becomes
If < 0
% Soil organic matter
Input % SOM here
Downgrade Defra
If > 0
moderate)
Upgrade Defra risk
from Flow Chart 1,
or keep at the very
Soil quality is not
high risk class
an issue
+
< 850 mm
+1
850 – 1200 mm
0
> 1200 mm
-1
Soil quality may be
Input here
Input final risk class
an issue
Figure 5.5 Methodology for upgrading/downgrading Defra risk class using soil quality indicators and average annual rainfall.
145
Chapter 5
inclusion of SOM in the screening toolkit will not increase the burden on Irish tillage
farmers as they are already required to monitor SOM levels on long-term (6 years or
more) continuous tillage land to ensure that levels stay above a threshold of 3.4%.
Soils with SOM levels above 3.4% (ca. 2% SOC) are considered not to be vulnerable
or depleted (DAFF, 2009a). In Ireland, where OM levels are found to be below the
threshold value of 3.4%, the farmer is obliged to seek advice from a specialist advisor
and, where appropriate, follow the programme of remedial actions. The programme of
action required depends on soil type and on-going practices.
Land use is included in the screening toolkit so that high and very high erosion risk
crops in each field‟s cropping regime can be identified by the farmer and the
necessary precautions taken to ensure that soil erosion is kept to a minimum. The
precautions taken in each case will depend on the land use risk class of the planted
crop and the erosion risk class of the field under assessment. The farmer should
consult with his/her specialist advisor/consultant to determine the best course of
action. If the precautions taken are ineffective and erosion persists, then the land use
should be ceased. The AAR in areas where tillage land is mainly concentrated (i.e.,
the midlands, south and east of Ireland) varies from ca. 657 mm to ca. 1400 mm. As
such, the risk of erosion occurring in a farmer‟s field is strongly influenced by its
location. Previously, Unwin (2001) observed that erosion problems in England were
worse in areas where the AAR exceeds a threshold of 800 mm. This resulted in the
country being divided into areas above and below a mean annual rainfall threshold of
800 mm when assessing the risk of erosion occurring. Boardman et al. (2009) pointed
out that such an approach ignored the problem of rainfall variability (both spatially
and temporally) and failed to consider „at-risk‟ periods where high daily/monthly
totals, for example, may have coincided with the most vulnerable periods for erosion
(sowing and post-harvesting); areas that normally fall below 800 mm could be at
serious risk in wetter years. While the problems identified by Boardman et al. (2009)
are valid, the use of an AAR threshold in erosion risk assessment is still appealing, as
it accounts for the role of rainfall amount (if not intensity and duration) in soil
erosion. As such, a similar AAR threshold as that used by Unwin (2001) is proposed
for use in Ireland to upgrade or downgrade a field‟s erosion risk class. Average annual
146
Chapter 5
rainfall thresholds are included among the proposed toolkit indicators (Table 5.10)
and can be obtained by the farmer for his location, from his/her specialist advisor.
5.4.2 Toolkit testing
In Table 5.11, the methodology developed in Figure 5.4 and Figure 5.5 is applied to
the study soils examined in the present study. There was better agreement between
erosion risk classes assigned using the screening toolkit developed in this study and
the observed erosion in study fields than there was between Defra (2005) erosion risk
classes and the observed erosion in study fields. The Fermoy soil of the present study
received scores of 2 for each of structure, porosity and ponding (Table 5.6), which are
assigned modified positive gradings of 2, 1, and 1, respectively, in Table 5.11. These
gradings are indicative of the good drainage class and lower runoff potential of the
Fermoy soil. Similarly, SOM received a positive grading of 1 because it is above the
threshold of 3.4%, while AAR received a neutral grading of 0. The total grading score
for the Fermoy soil of + 5 is evidence of good soil quality and can be used to
downgrade the Defra erosion risk class (determined by the farmer using Figure 5.4) of
the Fermoy field from very high to high, which is more in line with observed erosion
in the field. In the case of fields still flagged as high risk after downgrading, the
farmer should seek advice from his/her specialist advisor/consultant. Detailed
quantitative assessment of the Fermoy soil using the Soil Survey of Ireland method
and the K-factor equation (Table 5.8) confirmed that its low erodibility was largely
due to the rapid permeability of the soil profile and the fine crumb structure of its
plough layer, both of which lower its potential to produce runoff.
The same grading system was used to downgrade the risk class of the Tullow and
Duleek fields from moderate to lower risk, and to upgrade the Letterkenny and
Clonmel fields from high risk to very high risk, thereby ensuring the prioritisation of
areas where detailed investigation is required. According to the Defra (2005)
assessment, sediment is rarely seen to move in fields classed as lower risk, but
polluting runoff may enter ditches or watercourses. Good soil structure (+2), an
absence of surface water ponding (+1), moderate porosity (0) and healthy SOM (+1)
levels in both the Tullow and Duleek fields (Table 5.11) and a low AAR (+1) in the
147
Chapter 5
former, provides the evidence to support the downgrading of the Defra (2005) risk
class from moderate to lower in both fields. No erosion was observed in the Tullow or
Duleek fields; therefore, their downgrading from a Defra (2005) erosion risk class of
moderate to one of lower, using Chart 2 of the screening toolkit (Figure 5.5), is
justified. In the case of the Letterkenny field, moderate soil structure (0), the presence
of surface water ponding (-1), moderate porosity (0), and lower than desirable SOM
levels (-1) (Table 5.11) provides the evidence required to support the upgrading of the
Defra (2005) risk class from high to very high. Similarly, the moderate soil structure
(0), low porosity (-1), and high AAR (-1) recorded at the Clonmel field is evidence
enough to support the upgrading of the Defra (2005) risk class from high to very high.
The upgrading of the Letterkenny and Clonmel fields to very high risk is justified by
the observation of rills in most years in both. The K-factor erodibility values (Table
5.8) for the cropped areas of the Clonmel and Letterkenny fields were higher
(indicating greater risk) than in other study fields due to the poorer structure of their
plough horizons and slow permeability of their soil profiles. The newly developed
screening toolkit for refining Defra erosion risk classes proved effective (after a
preliminary validation) and has the added advantage of bringing to the farmer‟s
attention the soil quality status. The next step would be to make the screening toolkit
part of a larger scale erosion study in which measured edge-of-field sediment losses
can be used to further validate the erosion risk rankings assigned by farmers using the
screening toolkit. Comparisons of the amounts of SS initially mobilised by rainfall
and overland flow with the amounts measured in flow at appropriate monitoring
points (for example in drain flow, at the edge of fields or at catchment outlets)
provides a means of assessing the delivery factor in sediment transport (Beven et al.,
2005). The larger-scale erosion study should include a workshop where farmers and
specialist advisors can test the toolkit and make suggestions on how it can be further
refined based on their experience with soil erosion and soil quality.
148
Chapter 5
Table 5.11 Application of the methodology for upgrading/downgrading Defra risk classes to the study soils.
Modified gradings
Risk indicators
Fermoy
Tullow
Bunclody
Duleek
Clonmel
Letterkenny
Structure (-2,0,+2)
+2
+2
+2
0
0
0
Ponding (-1,0,+1)
+1
+1
+1
+1
0
-1
Porosity (-1,0,+1)
+1
0
0
0
-1
0
% SOM (-1,0,+1)
+1
+1
+1
+1
+1
-1
0
+1
0
+1
-1
0
+ 5 (downgrade)
+ 5 (downgrade)
+ 4 (downgrade)
+ 4 (downgrade)
- 1 (upgrade)
- 2 (upgrade)
Very high
Moderate
Lower
Moderate
High
High
↓
↓
↓
↓
↓
↓
High
Lower
Lower
Lower
Very high
Very high
AAR (-1,0,+1)
Total grading score
Defra (2005) risk class
(from Table 5.5)
New refined erosion risk class
SOM, soil organic matter; AAR, average annual rainfall
149
Chapter 5
5.5 Summary
This study developed a screening toolkit of erosion risk and soil quality indicators for
use by farmers and specialist advisors in assessing sites for likelihood of erosion and
reduced soil quality. The toolkit was developed by comparing results from quick and
easily conductible onsite assessments and more detailed quantitative assessments with
observed erosion features. Indicators proposed for inclusion in the toolkit included:
texture, slope, erosion features, structure, ponding, porosity, % SOM, AAR and
current land use. Assessment of these indicators at each study site using the grading
system developed in this chapter made it possible to correctly identify the sites where
the erosion risk was high enough to justify further investigation by a specialist
advisor/consultant. The grading system also informs the farmer of the impact of soil
quality on the assigned erosion risk class.
The screening method developed in this study can, if adopted by Irish tillage farmers,
ensure that the highest priority sites are identified expeditiously and with little cost
incurred by the farmer. Resources can then be focused on these sites ensuring the cost
effectiveness of mitigation measures employed to reduce SS and associated P loss,
such as buffer zones and minimum tillage. The screening method will have the added
benefit of bringing to the farmer‟s attention any issues with soil quality on his
holding.
150
Chapter 6
Chapter 6 Conclusions and recommendations
6.1 Overview
The aims of this study were to develop methods capable of identifying tillage fields
where STP and erosion risk levels are sufficiently high to cause concern over potential
losses of DRP, PP and SS in surface runoff following heavy rainfall, and to develop a
methodology for farmers/specialist advisors to identify tillage fields which may be
susceptible to erosion.
6.2 Integration of RDPRI findings with results from the screening toolkit
Integration of the results from the RDPRI and the screening toolkit, developed in this
study, is necessary in order to estimate the risk of sediment and phosphorus
mobilisation from Irish tillage soils, and subsequent delivery to surface water bodies.
The RDPRI and screening toolkit developed in this study are concerned primarily
with estimation of the risk of phosphorus and sediment being mobilised from Irish
tillage soils by rainfall and overland flow. The RDPRI goes somewhat further by
determining the soil test phosphorus threshold above which dissolved reactive
phosphorus mobilised from the soils, by simulated rainfall and overland flow, has the
potential to cause eutrophication of a surface water body. Therefore, tillage fields
identified by routine soil testing as having soil test phosphorus levels in excess of this
threshold can, based on the findings of the RDPRI, be termed critical source areas of
dissolved reactive phosphorus if the fields in question are hydrologically connected to
a surface water body. Sharpley and Tunney (2000) observed that threshold soil P
values have little meaning unless they are used in conjunction with an estimate of a
site‟s potential for surface runoff and erosion. The screening toolkit developed in this
study provides such an estimate by combining site factors (slope angle, land use and
annual average rainfall) and soil properties (texture, porosity, ponding, % soil organic
matter, and structure) that are known to affect runoff and erosion rates. As such,
integration of the RDPRI and the screening toolkit provides a more complete picture
of the risk of phosphorus and sediment being transported from tillage fields.
151
Chapter 6
The screening toolkit informs the farmer of the risk of erosion occurring on a
particular field. It does not inform him/her about the risk of an eroding field impacting
on a surface water body. The same is true of the Defra (2005) erosion risk assessment.
To address this, Boardman et al. (2009) proposed a simple extension of the Defra
scheme to account for the issue of connectivity between an eroding field and a river.
They proposed that the Defra scheme be adapted by taking account of high or very
high risk fields with a river within 200 m downslope. This resulted in certain fields no
longer being classed as at risk and others being confirmed as at risk of delivering
sediment to the river (i.e. high connectivity). Applying the same connectivity
extension to the study fields that were identified as high and very high risk using the
screening toolkit, results in the Fermoy and Letterkenny fields no longer being classed
as at risk (no river 200 m downslope) and highlights the Clonmel field as being a
critical source area of sediment and phosphorus (river within 200 m downslope). The
Morgan‟s P, water extractable phosphorus and Mehlich-3 phosphorus measured in the
Clonmel field greatly exceeded the respective thresholds determined for each by the
RDPRI and therefore the Clonmel field is also a critical source area of dissolved
reactive phosphorus. Given that reducing soil phosphorus to environmentally
acceptable levels in the Clonmel field will take many years of restricted fertiliser use,
methods of flow path manipulation (such as buffer zones or sediment traps) that
reduce connectivity between phosphorus source and receiving waters should be used,
as these can immediately reduce the amount of phosphorus reaching surface waters.
6.3 Conclusions
The main conclusions of the study are as follows:
1. Tilled soils, when subjected to high intensity, simulated rainfall and overland
flow, may produce DRP concentrations in excess of 0.03 mg L -1 (the value
above which eutrophication of rivers is likely to occur) if their P m, WEP, and
M3-P concentrations exceed 7.83 mg L-1, 4.15 mg kg-1, and 61.2 mg kg-1,
respectively.
152
Chapter 6
2. Water extractable P in both rainfall only and rainfall and overland flow
simulations was identified by statistical methods as being a better indicator of
the study soils potential to release DRP in surface runoff than P m, the national
soil P test.
3. This study showed that, while tillage soils with higher levels of soil P generate
higher losses of DRP in surface runoff, an important mechanism of P loss is by
detachment and transport of eroded soil particles. Therefore, the identification
and remediation of sites susceptible to erosion and associated PP loss should
be a high priority.
4. This study developed a methodology that can be used by Irish tillage farmers
with minimal training to identify fields at possible risk of soil erosion and
reduced soil quality. These sites can then be assessed in more detail by
specialist advisors. The method, which uses a combination of erosion risk and
soil quality indicators, correctly identified the study fields in which erosion
was observed as high risk.
5. The adoption of the screening method, developed in this study, by tillage
farmers in Ireland will enable the quantification of the extent of erosion risk
and soil quality degradation associated with tillage soils. This will allow
remediation measures to be prioritised for the most vulnerable sites, which is
likely to result in cost and resource savings for farmers and advisory services.
6.4 Recommendations for future work
1. Cereal growers in Ireland will face pressure to increase yields substantially
into the future to maintain margins and satisfy the demand for home grown
concentrate feed due to the projected expansion in the dairy, drystock and pig
sectors by 2020 (DAFF, 2010). The intensification required to meet these yield
demands may impact negatively on soil quality and increase erosion risk,
particularly where the soil type is prone to damage and/or the site is in longterm tillage. These problems can lead to a reduction in productivity unless the
153
Chapter 6
specific sites are identified and remedial action is taken. It is proposed the
screening toolkit developed in this study can be used by farmers to identify
sites with soil quality/erosion issues. This should be done systematically in
conjunction with regular soil testing so that any soil quality problems
identified by farmers can subsequently be discussed with their advisors when
soil tests are being interpreted. The required course of action to reverse any
degradative trend including any identified in the soil test results such as
excessive soil P status (> 10 mg L-1 Pm) can then be set out in a management
plan for the farm.
2. If predicted increases in the magnitude and frequency of storm events in
Ireland due to climate change are accurate, then soil erosion may become a
significant threat to soil and water quality in tillage areas by the 2020s, through
removal of topsoil, decline of SOM and loss of nutrients. The proposed SFD
has recognised erosion as a threat to soil quality and, if ratified, will require
member states to identify areas prone to soil degradation processes like
erosion. The screening toolkit may be used to assist with this work and also
help identify the causes of the erosion.
3. Soils in long-term tillage and/or under intensive management conditions are
prone to problems relating to soil quality. Further research is needed to
determine „best management practice‟ for inclusion in remedial plans that will
improve soil quality and productivity, and reduce erosion levels and associated
P loss on fields identified as high risk by the farmer using the screening
toolkit. In order to develop comprehensive best management strategies that can
control P loss from fields and/or catchments, all hydrologic implications,
particularly variable source area concepts of runoff generation, must be
incorporated. These best practice measures will also be required for inclusion
in the POM, which Ireland will be required to implement in identified risk
areas should the SFD be ratified.
154
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202
APPENDICES
203
Appendix A List of publications
JOURNAL PAPERS (Accepted)
Regan, J.T., Rodgers, M., Healy, M.G., Kirwan, L. and Fenton, O. 2010. Determining
phosphorus and sediment release rates from five Irish tillage soils. Journal of
Environmental Quality 39: 185-192.
Regan, J.T., Fenton, O. and Healy, M.G. 2012. A review of phosphorus and sediment
release from Irish tillage soil, the methods used to quantify losses and the current state
of mitigation practice. Biology and Environment: Proceedings of the Royal Irish
Academy, 112B: 157 - 183. DOI: 10.3318/BIOE.2012.05.
Creamer, R.E., F. Brennan, O. Fenton, M.G. Healy, S.T.J. Lalor, G.J. Lanigan, J.T.
Regan, and B.S. Griffiths. 2010. Implications of the proposed Soil Framework
Directive on agricultural systems in Atlantic Europe – a review. Journal of Soil Use
and Management 26: 198-211.
MANUSCRIPTS IN PREPARATION
Regan, J.T., Fenton, O., Walsh, M. and Healy, M.G. Physical, chemical and visual
evaluation of 6 Irish tillage soils to assess soil quality and susceptibility to erosion
(Target journal: Soil and Tillage Research).
Regan, J.T., Fenton, O., Grant, J. and Healy, M.G. Estimating phosphorus and
sediment release rates from five Irish tillage soils when subjected to increasing
overland flow rates (Target journal: Science of the Total Environment).
INTERNATIONAL CONFERENCE PAPERS
Regan, J., Rodgers, M., Healy, M.G., Fenton. O. 2008. Sediment and nutrient loss
from five Irish tillage soils at a 30 mm hr -1 rainfall intensity. EUROSOIL Congress,
University of Technology Vienna, Austria, 25 -29 August, 2008 (Oral presentation).
204
NATIONAL CONFERENCE PAPERS
Regan, J., Rodgers, M., Healy, M.G., Fenton, O., Walsh, M. 2008. Soil erosion and
nutrient loss from Irish tillage soils. Agricultural Research Forum, Tullamore, Co.
Offaly, 10-13 March, 2008 (Oral presentation).
Regan, J., Rodgers, M., Healy, M.G., Fenton. O. 2008. Sediment loss and surface
runoff from Irish tillage soils. ESAI Colloquium, Dundalk IT, February, 2008 (Oral
presentation).
Regan, J., Rodgers, M., Healy, M.G., Fenton, O. 2007. Soil erosion and surface runoff
from Irish tillage soils. Teagasc Walsh Fellowship Seminar, RDS, Dublin, November
14th, 2007 (Poster presentation).
Regan, J., Rodgers, M., Healy, M.G., Walsh, M., Fenton, O. 2007. Enrichment ratios
for a range of Irish soils. ESAI colloquium, IT Carlow, January, 2007 (Oral
presentation).
OTHER PUBLICATIONS
Regan, J.T., Rodgers, M., Healy, M.G., Kirwan, L. Fenton, O.. 2008. Sediment and
nutrient loss from five Irish tillage soils at a 30 mm hr -1 rainfall intensity. Teagasc
grassland and EU Water Framework Directive conference, Johnstown Castle,
Wexford (Poster presentation).
Regan, J., Healy, M.G., Rodgers, M. and Fenton, O. 2008. Soil erosion and surface
runoff from Irish tillage soils. Farmfest, Teagasc, Athenry, Galway (Poster
presentation).
Fenton, O., Hyde, B., Ó hUallacháin, D., Healy, M.G., Regan, J., and Rodgers, M.
2007. Tackling nutrient loss head on: catching the nutrients that got away. TResearch
2(2): 32-34. ISBN: 1649-8917.
205
Regan, J., Rodgers, M., Healy, M.G., Fenton, O., Walsh, M. 2007. Soil erosion and
surface runoff from Irish tillage soils. Pp. 25-26. In: J.A. Finn, K. Richards, G. Shortle
(eds). Ireland‟s rural environment. Johnstown Castle Environment Research Centre.
ISBN: 18 4170 477 6.
206
Appendix B
Appendix B Modelling soil erosion and P and SS delivery to
surface waters at the catchment scale
Many different kinds of models are available for use to simulate soil erosion and
sediment and P delivery to waterways at the catchment-scale. In general, these models
fall into three main categories: (1) empirical (2) conceptual and (3) physical or process
based. However, the difference between the model categories is not always clear, and
making the distinction can be somewhat subjective (Merritt et al., 2003). For example, it
has been argued by Lowe (2006) that the Hydrological Simulation Program - FORTRAN
(HSPF) (Bicknell et al., 1996), which has been classed as a conceptual model by many
studies is, in fact, a physically-based model. Previous work by Merritt et al. (2003)
provides a comprehensive review of erosion and sediment transport models. For the
purposes of this review, the focus will be on catchment-scale models that have been used
in Ireland to estimate soil erosion, and P and sediment delivery to waterways. These are
empirical models (Revised Universal Soil Loss Equation (RUSLE) and Sediment
Distribution Delivery (SEDD)) and physically-based models (HSPF (Bicknell et al.,
1996), Soil Water Assessment Tool (SWAT) (Arnold et al., 1998), Système
Hydrologique Européen TRANsport (SHETRAN) (Ewen et al., 2000)) and a modified
version of TOPMODEL (Scanlon et al., 2005). Where possible, the losses estimated
using these models are compared with losses from the same models applied in other
countries and with measured losses from Irish and international catchments (Table B.1
and Table B.2). Caution is required when comparing results from these tables given the
effect of catchment size on SS yield.
Empirical models
These models are generally considered to be the simplest of the three model types and are
frequently used in preference to more complex models as they can be implemented in
situations with limited data and parameter inputs, and are particularly useful as a first step
in identifying sources of sediment and nutrient generation (Merritt et al., 2003). They are
209
Appendix B
derived from the analysis of field observations and endeavour to characterise response
from these data.
The Universal Soil Loss Equation (USLE) (Wischmeier and Smith, 1978) is a program
used widely in America and worldwide that estimates the long-term water erosion from
inter-rill and rill areas. It is represented by the equation: A = RKLCSP, where A is the
estimated soil loss per unit area, R is the rainfall erosivity factor, K is the soil erodibility
factor, L is the slope-length factor, S is the slope-steepness factor, C is the cover and
management factor, and P is the support practices factor. The USLE was revised
(RUSLE) (Renard et al., 1991) and revisited (Renard et al., 1994) to take into account
additional information that had become available since its development. Although
developed for application to small hillslopes, the USLE and its derivatives have been
incorporated into many catchment-scale erosion and sediment transport modelling
applications (Merritt et al., 2003). The SEDD model is based on the USLE model. It
discretizes a catchment into morphological units (areas of defined aspect, length and
steepness) and determines a SDR for each unit (Fernandez et al., 2003). The SDR is the
ratio of sediment reaching a continuous stream system to the total amount of sediment
eroded by sheet and channel erosion. The magnitude of the SDR for a particular
catchment will be influenced by a wide range of geomorphological, hydrological,
environmental and catchment factors (Fu et al., 2006).
210
Appendix B
Table B.1 Erosion and sediment yield in a selection of Irish and international studies.
Catchment
Area
Land use
Assessment method
Study Period
2
(Km )
Bush, Co. Antrim, Ireland
< 340
Arable (36.6%)
In stream sampling and
Precipitation
Runoff
Erosion
(months)
(mm)
(mm)
(ton km
12
-
-
-
Sediment yield
-2
-1
yr )
-2
(ton km
Dripsey Co. Cork, Ireland
14
Grassland
Clarianna, Co. Tipperary, Ireland
29.8
dominated
Oona Water, Co. Tyrone, Ireland
88.5
Gelbaek, Central Jutland, Denmark
11.6
Arable
In stream sampling
Evans et al.
(2006)
12
In stream sampling
yr )
51.4 – 107 †
sediment fingerprinting
Reference
-1
1833
1037
-
13.60
Jordan et al.
(2005a)
12
1091
434
-
8.40
12
1366
817
-
41.0
12
932
369
-
7.1
Kronvang et
al. (1997)
Belmont, Herefordshire, England
1.5
Arable (61%)
In stream sampling,
12
660
268
466.6
81.9
Walling et
Lower Smisbey, Leicestershire
2.6
Grassland/arable
turbidity sensors, sediment
12
660
-
400.5
80.3
al. (2002)
24
647 - 706
-
507 (ploughed)
2.4
Walling et
Pang, Berkshire, England
166
Arable dominated
fingerprinting,
137
Cs
measurements
Lambourn, Berkshire, England
234
140 (grass)
Grassland/arable
24
698 - 793
-
437 (ploughed)
al. (2006)
3.7
95 (grassland)
Ireland
-
Arable
RUSLE and SEDD models
-
-
-
1978
22.0 ††
He et al.
Dripsey Co. Cork, Ireland
14
Grassland
RUSLE and SEDD models
-
-
-
-
909.0
(2010)
Pataha Creek Watershed,
327
Arable
RUSLE and SEDD models
-
250 - 1000
-
1766
711
Washington, USA
Lawyers Creek, Watershed, Idaho,
308
Arable
RUSLE and SEDD models
USA
†
-
533 - 737
-
2150
660
Fernandez et
al. (2003)
-1
This value is given in ton yr because the exact catchment area was not given in this paper
††
Fu et al.
(2006)
Model prediction for the total tillage area of Ireland
211
Appendix B
Table B.2 Phosphorus delivery to streams in a selection of Irish and international catchments.
Catchment
Area
Land use
2
(Km )
Study period
Annual
SWAT
HSPF
-1
GOPC
-1
Measured
-1
Reference
-1
(months)
Precipitation (mm)
(kg P yr )
(kg P yr )
(kg P yr )
(kg P yr )
Dripsey, Co. Cork
15
Grassland
12
1833
1371
1530
1389
1719
Clarianna, Co. Tipperary
23
Grassland
8
1091
231
136
243
289
Oona water, Co. Tyrone
96
Grassland
12
1366
33285
25717
12519
27496
Gelbaek, Central Jutland, Denmark
11.6
Arable
12
932
-
-
-
371.2
Kronvang et al. (1997)
Sagån, Sweden
864
Arable
36
618
31104
-
-
36288
Ekstrand et al. (2010)
Lake Fork, East Central Illinois, USA
365
Arable
72
960
-
-
-
17650
Gentry et al. (2007)
Belmont, Herefordshire, England
1.5
Arable (61%)
48
660
-
-
-
405
Withers and
Jordan et al. (2005a)
Hodgkinson (2009)
212
Appendix B
The combined use of Geographical Information Systems (GIS), RUSLE and SEDD
has been shown to be an effective method for estimating water erosion and sediment
yield by Fernandez et al. (2003) and Fu et al. (2006), and for estimating the impacts of
no-till practice on soil erosion and sediment yield by Fu et al. (2006). A case study in
Ireland by He (2010) to estimate soil erosion and sediment yield using GIS, RUSLE
and SEDD, predicted that the average SS delivered from arable land to waterways
was 0.22 ton ha-1 yr-1. However, this finding should be treated with caution because
the catchment-specific parameter β was only estimated for the Dripsey catchment
(using an inverse modelling approach employing observed SDR values from fields)
and sensitivity analysis of β by increments of 1.0 to a maximum of 20.0 was carried
out to infer possible values of β for the Bandon, Dromcummer, Duarrigle and Mallow
catchments. Fu et al. (2006) tested β between 0.5 and 2.0 with an increment of 0.1 and
found that the SDR was very sensitive to the values of β, varying from 0.6 (β = 0.5) to
0.27 (β = 2.0).
Physically-based models
Physically-based models are those in which the model equations are based on physical
laws and relationships. They are more complex than empirical models and require
more measurement and calibration of model parameters. Complex physical models
applied with the necessary expertise or user support can be far superior where there is
a need to address spatial and temporal complexities (Irvine et al., 2005). According to
the DPSIR conceptual framework (Drivers, Pressures, State, Impact and Response)
(Irvine et al., 2005) that guides the selection of modelling techniques in Ireland, it is
likely that the most useful models will be of the physically-based or mechanistic-type
(Nasr et al., 2007).
Nasr et al. (2007) tested three widely used physically-based models (SWAT, HSPF
and SHETRAN), coupled with the grid orientated P component (GOPC) (Nasr et al.,
2005) of diffuse P pollution, in three Irish grassland catchments, to explore their
suitability in Irish conditions for future use in implementing the WFD. These models
range from semi-empirical (SWAT) to fully physically-based (SHETRAN/GOPC) in
how they represent the relevant hydrological, chemical and bio-chemical processes
213
Appendix B
transforming the P compounds both in the soil and during its transport by water (Nasr
et al., 2007). The catchments were selected based on data availability and different
climate, land use and soil type. Soil water assessment tool is a continuous model
working at the basin-scale to look at the long-term impacts of management and also
timing of agricultural practices within a year on water, sediment, and agriculture
chemical yields in large un-gauged basins (Arnold et al., 1998). Hydrological
Simulation Program - FORTRAN is a lumped-parameter model that simulates
hydrology and water quality processes on a continuous basis in natural and man-made
water systems (Im et al., 2003). SHETRAN/GOPC is a fully physically-based model
which relies on relationships derived from physical and chemical laws. The three
models also differ in their representation of the spatial variation within the catchment
and the time steps at which they can simulate. In order of ability to match the
measured discharge hydrographs from each catchment in this study, the models
performed (from best to worst) as follows: HSPF, SWAT, SHETRAN. The best
simulation for daily TP loads in the study catchments was by SWAT. In the short
term, Nasr et al. (2005) recommended using SWAT for TP load estimation. The
SWAT model recently showed good potential for predicting TP losses from arable
land in a Swedish study by Ekstrand et al. (2010) (Table B.2). While computer models
such as the SWAT, can give detailed output on P losses at the catchment scale, they
require large quantities of detailed input data as well as knowledge on how to run the
computer software and interpret results, and do not necessarily represent P loss
processes any better than P-indices (Veith et al., 2005).
TOPMODEL is a process based semi-distributed catchment model (Irvine et al.,
2005) in which the major factors affecting runoff generation are the catchment
topography and the soil transmissivity, which diminishes with depth (Parsons et al.,
2001). In this model, overland flow generation follows the variable source area
concept, while groundwater discharge is from a permanent water table. TOPMODEL
is not intended to be a traditional modelling package, but a collection of concepts to
help in understanding and predicting the hydrological behaviour of basins (Parsons et
al., 2001). It was used for this purpose by Scanlon et al. (2005) when a modified
version of TOPMODEL was developed and applied to a 14.2 ha grassland catchment
in Ireland in order to infer the significant pathways of soil-to-stream P transport. In
214
Appendix B
this study, a physically-based hydrological model generated pathway-specific
information for three components of discharge: overland flow, shallow subsurface
flow and groundwater discharge. An independent comparison of the hydrological
model output and stream water P measurements allowed the authors to infer the
relative contributions from individual pathways to the overall P transport. They found
that the fraction of modelled stream discharge deriving from overland flow and
shallow subsurface flow was a reliable descriptor of the observed TP concentrations.
Shallow subsurface flow was inferred to be the dominant P transport mechanism,
primarily due to much greater volumetric contributions to stream discharge deriving
from it than from overland flow. These model results challenge the commonly held
assumption that the majority of P transport occurs via surface runoff and could have
important implications for the design and implementation of remedial measures
(Scanlon et al., 2005).
215
Appendix C
Appendix C Relationships between DRP in runoff and soil
extractable P methods
Time zero
1-hr interval
24-hr interval
5
Log FWMDRP
4
3
2
1
0
5
4
3
2
1
0
0
2
4
6
8
10
12
0
2
4
6
8
10
12
0
2
4
6
8
10
12
Water Extractable P (mg kg-1)
Figure C.1 Log FWMDRP against water extractable P for selected tillage soils at
a 10 degree slope under a 30 mm hr-1 rainfall and subjected to two distinct
overland flow rates (225 and 450 ml min-1).
216
Rainfall/overland (450)
0
Rainfall/overland (225)
1
Simulated rainfall only
2
Runoff rate = 650 ml min-1
3
Runoff rate = 425 ml min-
4
Runoff rate = 200 ml min-1
5
Appendix C
Time zero
1-hr interval
24-hr interval
5
Log FWMDRP
4
3
2
1
0
5
4
3
2
1
0
0
10
20
30
40
0
10
20
30
40
0
10
20
30
40
Soil P Saturation (%)
Figure C.2 Log FWMDRP against soil P saturation for selected tillage soils at a
10 degree slope under a 30 mm hr-1 rainfall and subjected to two distinct
overland flow rates (225 and 450 ml min-1).
217
Rainfall/overland (450)
0
Rainfall/overland (225)
1
Simulated rainfall only
2
Runoff rate = 650 ml min-1
3
Runoff rate = 425 ml min-
4
Runoff rate = 200 ml min-1
5
Appendix C
Time zero
1-hr interval
24-hr interval
5
Log FWMDRP
4
3
2
1
0
5
4
3
2
1
0
0
1
2
3
0
1
2
3
0
1
2
3
Calcium Chloride Extractable P (mg L-1)
Figure C.3 Log FWMDRP against calcium chloride extractable P for selected
tillage soils at a 10 degree slope under a 30 mm hr-1 rainfall and subjected to two
distinct overland flow rates (225 and 450 ml min-1).
218
Rainfall/overland (450)
0
Rainfall/overland (225)
1
Simulated rainfall only
2
Runoff rate = 650 ml min-1
3
Runoff rate = 425 ml min-
4
Runoff rate = 200 ml min-1
5
Appendix C
Time zero
1-hr interval
24-hr interval
5
Log FWMDRP
4
3
2
1
0
5
4
3
2
1
0
0
20
40
60
80
100
0
20
40
60
80
100
0
20
40
60
80
100
-1
Mehlich 3 P (mg kg )
Figure C.4 Log FWMDRP against Mehlich 3 P for selected tillage soils at a 10
degree slope under a 30 mm hr-1 rainfall and subjected to two distinct overland
flow rates (225 and 450 ml min-1).
219
Rainfall/overland (450)
0
Rainfall/overland (225)
1
Simulated rainfall only
2
Runoff rate = 650 ml min-1
3
Runoff rate = 425 ml min-
4
Runoff rate = 200 ml min-1
5
Appendix D
Appendix D Soil texture classification (Defra, 2005;
Environment Agency, 2007).
220
Appendix E
Appendix E Visual soil assessment score card, tables and
visual aids (Shepherd, 2009).
Soil texture
221
Appendix E
Soil structure
Soil porosity
Number and colour of soil mottles
222
Appendix E
Soil colour
Earthworms
Soil smell
Potential rooting depth (a)
223
Appendix E
Potential rooting depth (b)
Surface ponding
Surface cover and surface crusting
224
Appendix E
Soil erosion
225
Appendix F
Appendix F Field procedure for visual evaluation of soil
structure (Ball et al., 2011; Guimarães et al., 2011)
226
Appendix G
Appendix G Soil classification data
In order to formally classify the study soils according to the United States Soil
Taxonomy (Soil Survey Staff, 1975, 1998) and FAO (2006) classification systems,
analytical data (Table G.2) is needed to supplement the field-based observations
(Table G.1). The full classification for each soil is presented in Table G.3 along with
soil type, soil series (where available) and parent material. The soil series is defined as
a ‘collection of soil individuals essentially uniform in differentiating characteristics
and in arrangement of horizons’ (Gardiner and Radford, 1980). Only four of the six
study sites were located in counties previously mapped by the Soil Survey of Ireland.
Therefore, no soil series was assigned to the Clonmel and Letterkenny soils.
227
Appendix G
Table G.1 Soil profile descriptions
Horizon
Depth
Description
(cm)
Ap
0-23
Gravelly loam; yellowish brown to strong brown (10YR 5/6- 7.5YR 5/6); moderate fine crumb and
Bunclody
sub-angular blocky; many very fine discontinuous random inped pores; friable; many very fine roots
in top 10 cm, very few very fine roots from 10-23 cm; clear smooth boundary to:
B
23-55
Gravelly sandy loam; reddish yellow (7.5YR 6/8); medium crumb; common very fine discontinuous
C
55-100+
Stony gravelly sandy loam; reddish brown (5YR 5/4); weak coarse sub-angular blocky; compact in
random pores; very few very fine roots; gradual wavy boundary to:
situ; common very fine discontinuous random pores; no roots
Ap
0-27
Sandy loam; brown to dark brown (10YR 4/3); moderate fine to medium crumb; grading with depth
to moderate fine to medium sub-angular blocky with crumb as above; friable; common very fine to
fine random inped pores; common very fine to fine roots (winter wheat growing); clear smooth
Tullow
boundary to:
B21t
27-50
Sandy clay loam; yellowish brown (10YR 5/6); Weak to moderate medium and coarse sub-angular
blocky with some medium and coarse crumb; friable; common very fine random inped pores; few
very fine roots gradual wavy boundary to:
B22t
50-60/75
Sandy clay loam; yellowish brown (10YR 5/6); moderate fine and medium sub-angular blocky;
many very fine to fine random inped pores; few very fine roots; clear wavy boundary to:
C
60/75+
Sandy loam; gravelly; weak red (2.5YR 5/2); glacial till.
Ap
0-30
Sandy loam; stony; dark yellowish brown (10YR 4/4); weak to moderate fine and medium subangular blocky breaking to very fine and fine crumb; friable; common very fine random inped
Fermoy
pores; common very fine roots; smooth gradual boundary to:
B
30-70
Sandy loam; stony; strong brown (7.5YR 5/6); weak medium to coarse sub-angular blocky breaking
to very fine and fine crumb; friable; few very fine random inped pores; few very fine roots gradual
smooth boundary to:
C
70+
Sandy loam; stony; brown (7.5YR 5/4); B material mixed with C to a depth of 110 to 120 cm;
stone/pebble ghosts of limestone common below a compact layer 120 to 140 cm, which contains
green shale fragments.
Ap
0-32
Sandy loam; dark brown (10YR 4/3); coarse crumb in top 5cm; moderate medium sub-angular
blocky (5-32cm), firm, common very fine vertical pores; ploughed-in stubble evident to depth of
30cm; common very fine roots; clear smooth to wavy boundary to:
B21
32-57
Sandy loam to loam; strong brown (7.5YR 5/6); localised patches of fine distinct sharp black 10YR
Clonmel
2/1 (Mn) mottles; few medium random pores; weak coarse angular blocky; firm in situ friable when
broken (fragipan); few very fine roots and very few medium roots; gradual smooth boundary to:
B22
57-90
Sandy loam; stony; dark yellowish brown (10YR 4/4) matrix; common coarse faint diffuse reddish
brown (5YR 4/4) mottles; weak coarse sub-angular blocky with traces of clay skins; firm and
compact; few very fine random pores; no roots; gradual diffuse boundary to:
C1
90-160+
Gravelly loam to clay loam; gravely; yellowish brown (10YR 5/6) with many fine distinct sharp
black (10YR 2/1) (Mn) mottles; massive slightly sticky; slightly plastic; rare clay skins; many
decaying pebbles; no roots
Letterkenny
Ap
0-25
Sandy loam; common decaying fragments of shale; yellowish brown (10YR 5/4); weak fine crumb
in top 3cm grading to weak medium sub-angular blocky below; friable; common very fine pores;
common fine and very fine roots and ploughed in stubble; clear wavy boundary to:
B21g
25-56
Loam; greyish brown (2.5Y 5/2); many medium distinct olive to dark olive 5Y 5/8 mottles; massive;
firm; many very fine random pores; very few very fine roots; gradual smooth boundary to:
B22g
56-96
Loam; grey (5Y 5/1); many coarse distinct diffuse strong brown (7.5YR 5/6) mottles; localised
228
Appendix G
areas of common medium black (10YR 2/1) manganese mottles; massive; compact; few very fine
pores; no live roots; common medium fossil roots; gradual smooth boundary to;
B23g
96-130
Loam; olive (5Y 5/3); massive; sticky; few fine pores; common coarse fossil roots; no live roots;
Cg
130-155+
Loam; olive (5Y 5/4); massive; very sticky and plastic; few medium pores; no roots.
Ap
0-30
Loam; brown to dark brown (10YR 4/3); weak fine crumb and weak medium sub-angular blocky;
patches of soil with similar matrix and mottling to B22g; gradual smooth boundary to;
friable; many very fine continuous random inped pores; common very fine roots; clear smooth
boundary to:
Duleek
B2
30-45
Loam; yellowish brown (10YR 5/8); weak medium crumb with some weak medium sub-angular
blocky; very friable; many very fine continuous random inped pores; few very fine roots; gradual
wavy boundary to:
B3
45-60
Loam (more clay than above); brown (10YR 5/3); weak coarse sub-angular blocky and weak coarse
crumb; friable; common very fine continuous random inped pores; very few very fine roots; diffuse
wavy boundary to:
B3/C
60-115+
Loam (gritty in some parts clayey in others); yellowish brown (10YR 5/4); structureless; massive;
very firm; compact; common very fine discontinuous pores; no roots;
229
Appendix G
Bunclody pit to a depth of 100 cm
Tullow pit to a depth of 100 cm
Fermoy pit to a depth of 100 cm
Clonmel pit to a depth of 160 cm
Letterkenny pit to a depth of 155 cm
Duleek pit to a depth of 115 cm
Figure G.1 Soil profiles
230
Appendix G
Figure G.2 Thin horizontal layers of pale sand grains within the cultivated layer
resulting from disintegration of aggregates into their component particles.
Figure G.3 Fine and very fine sand deposited at tramline ending.
231
Appendix G
B21g
Figure G.4 Observed mottling in the B21g horizon of the Letterkenny soil.
232
Appendix G
Table G.2 Analytical data for use in soil classification
Soil
Horizon
Sand
(FSa)
Silt
Clay
(FSi)
pH
CEC
Ca
Mg
(1:1)
Tullow
Fermoy
Clonmel
Letterkenny
Duleek
Fe
%
%
(cmol kg )
% (%)
Bunclody
SOC
-1
Ap
42(20)
37(28)
21
7.8
13.7
12.4
1.0
2.8
1.6
B
71
24(16)
5
7.7
4.7
2.5
0.3
0.8
1.2
C
54
37(26)
9
7.5
4.0
2.1
0.3
1.0
1.2
Ap
58(25)
28(18)
14
6.5
12.8
7.1
1.3
2.0
0.8
B21t
50
28(18)
22
7.2
4.7
4.9
1.3
1.0
1.3
B22t
54
26(16)
21
8.2
9.4
0.7
0.1
0.8
1.3
C
64
22(14)
14
8.8
4.7
1.5
0.0
0.7
0.8
Ap
56(26)
30(17)
14
6.1
13.3
7.3
0.7
2.4
0.8
B
63
31(16)
6
6.9
5.5
4.3
0.3
1.4
0.8
C
61
29(16)
10
7.0
4.5
1.5
0.0
0.6
1.0
Ap
53(23)
31(21)
16
6.4
13.0
7.1
0.3
2.1
1.0
B21
52
31(21)
17
7.0
7.2
3.7
0.3
1.2
1.1
B22t
54
29(19)
17
7.1
4.8
2.5
0.3
0.6
1.4
C1
41
32(22)
27
7.1
2.0
4.1
0.3
0.7
1.7
Ap
54(33)
36(21)
10
5.6
11.1
4.3
0.3
1.9
0.9
B21g
47
40(25)
13
7.4
5.4
5.5
0.3
0.8
1.5
B22g
46
42(29)
12
7.4
7.7
5.1
0.7
0.6
1.4
B23g
45
42(27)
13
7.4
7.6
4.6
0.7
0.8
1.4
Cg
45
42(28)
13
7.8
6.3
3.9
0.7
0.7
1.0
Ap
40(19)
40(28)
20
6.7
12.7
12.2
0.3
2.4
0.9
B2
48
35(23)
17
6.8
11.9
6.1
0.3
1.6
1.0
B3
51
36(25)
13
6.9
6.9
4.7
0.3
1.0
1.2
B3/C
47
34(24)
19
6.7
8.0
6.5
0.3
1.0
1.3
FSa, fine sand; FSi, fine silt; CEC, cation exchange capacity; Ca, calcium; Mg, Magnesium; SOC, Soil
organic carbon; Fe, Iron
233
Appendix G
Table G.3 Soil classification at selected tillage sites
Study site
Parent material
Soil type
American
WRB
classification
classification2
Soil Series
system1
Bunclody
Glacial drift (Ordovician
BEHBS
sandstones, slates,
`Tullow
Typic
Chromi-Eutric
Soil Series
Eutrochrept
Cambisol
Variant of the
limestone, and igneous
Ballindaggan
rocks)
Series3
Glacial drift (limestones,
GBP
igneous rocks and
Typic
Haplic Luvisol
Kellistown
Series4
Hapludalf
sandstones)
Fermoy
Glacial drift (ORS, shales
BEHBS
and sandstones, and
Typic (Arenic)
Chromic- Eutric
Variant of the
Eutrochrept
Cambisol
Knocknaskeha
Series5
limestones)
Clonmel
Glacial drift (limestones,
Gley
shales and sandstones)
Letterkenny
Glacial drift (schist and
Glacial drift (Ordovician
shales sandstones and
Epigleyi-Eutric
Cambisol
Aeric
Gleysol
Unmapped
Eutric Cambisol
Kells Series6
Gley
gneiss)
Duleek
Aquic
Eutrochrept
Unmapped
Fragiaquept
BEHBS
Typic
Eutrochrept
igneous rocks)
WRB, World reference base; GBP, grey brown podzolic; BEHBS, brown earth with high base status; 1Soil Survey
Staff, 1993,1999; 2FAO, 2006; 3Gardiner and Ryan,1964; 4Conry and Ryan, 1967; 5Finch, 1971; 6Finch et al., 1983.
234
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