Agricultural Pollution Problems and Practical Solutions [New edition] 9780415273404, 0415273404, 0203302028, 9780203302026, 9780203351543

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Agricultu ral Pollution

Spon’s En viron mental Scien ce and En gin eering Series

This new series cover s a wide range of water, waste and contam inated land issues in the context of current best practice, perception and legislation. Series coverage is broad. On the water and wastewater side it includes r esource m anagement, treatment, distribution and collection, monitoring and regula tion. In relation to waste m anagem ent it extends across the waste hierarchy, covering a range of technical and regulatory issues in areas such as waste m inim isa tion, separation and sorting, recycling and disposal, and the determ ination of appropriate waste strategy. The series is targeted at en gineers an d scien tists in the process, waste, an d environ m en tal sectors. Titles will also be of interest to econom ists, lawyers, legislators, regulators, and a dvanced students. S e r i e s E d i to r J erem y J oseph, J BJ Environment, Hydrogeologica l and Waste Managem ent Consultant, C/ o Spon Press, 11 New Fetter Lane, London EC4P 4EE, UK To p i c s u n d e r co n s i d e ra ti o n fo r th e s e ri e s i n cl u d e • • • • • • •

Potable/ usable water Waste and polluted waters and sludges Water control an d m anagem ent issues Contam inated la nd Non-disposal str ategies for waste Waste disposal routes Wastes from agr iculture

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Agricultural Pollution Environm ental problems and practical solutions

Graham Merrington, Linton Winder, Robert Par kinson and Mark Redm an

Lon don an d New York

First p ublish ed 20 0 2 by Spon Press 11 New Fetter La ne, Lond on EC4P 4EE This ed ition p ublish ed in th e Taylor & Fran cis e-Librar y, 20 0 5. “To pu rch ase your own copy of t his or an y of Taylor & Fr ancis or Routledge’s collect ion of t housan ds of eBooks please go to www.eBookstor e.tan df.co.uk.” Sim ultaneously p ublish ed in t he USA an d Can ada by Spon Press 2 9 West 35th Str eet, New York, NY 10 0 0 1 Spon Press is an imprint of the Taylor & Francis Group © 20 0 2 Gr aham Merr in gton, Lint on Win der, Rober t Par kinson and Mark Red m an Publisher’s Note Th is book has b een pr ep ared fr om cam er a-ready copy su pplied by t he aut hors All r ight s reser ved . No part of th is book m ay be reprint ed or r ep roduced or u tilized in an y form or by an y electron ic, m ech anical, or ot her m ean s, now kn own or h er eaft er in vented , in clud in g ph otocop ying an d recording, or in an y in form ation storage or retrieval syst em , with ou t per mission in writ ing from th e pu blishers. British Library Cataloguing in Publication Data A catalogu e record for th is book is available from th e British Lib rary Library of Congress Cataloging in Publication Data A cat alog record for t his b ook has been requ ested ISBN 0 -20 3-30 20 2-8 Master e-book ISBN

ISBN 0 -2 0 3-34177-5 (Adobe eReader For mat ) ISBN 0 -415-27340 -4 (h bk) ISBN 0 -419-21390 -2 (pbk)

Contents

1 1.1

List of figures

xii

List of tables

xiv

Acknowledgements

xvi

Agriculture and pollution

1

Setting the scene

1

Defining agricultural pollution 1.2

2 2.1

Solving the problem

2.3

5

Monitoring impacts

5

Cost—the driver for pollution control

7

Nitrates and nitrogen loss

11

Introduction

11

Nitrogen and agriculture 2.2

3

The nitrogen cycle

11 13

Atmospheric deposition

13

Biological fixation

14

Fertilisers

15

Animal manures

16

Nitrogen transformations

16

Nitrogen mineralisation

17

Nitrogen immobilisation

17

Nitrification

18

Adsorption and fixation

18

Nitrogen uptake and losses

18

Nitrate leaching

18

vi

Nitrous oxide and nitric oxide emissions from nitrification and denitrification

20

Ammonia volatilisation

21

2.4

Nitrogen fertiliser use in agriculture

22

2.5

The causes of nitrate pollution

24

2.6

Problems caused by nitrate pollution

27

Nitrate and drinking water quality

27

Eutrophication

28

Nitrates in food

29

2.7

3 3.1

Reducing nitrate losses from agriculture Source management—optimisation of fertiliser N use

32

Cultivation and crop management

33

Phosphorus

41

Introduction

41

Phosphorus and agriculture 3.2

3.3

3.4

The phosphorus cycle

41 42

Phosphorus inputs and outputs

43

Fertilisers

44

Atmospheric inputs

45

Animal manures

46

P transformations in soil

47

Distribution of P in soil

47

P sorption

50

Phosphorus losses

51

Surface run-off

52

Leaching and drain flow

54

Environmental implications of P pollution N and P balance in natural waters

3.5

29

Environmental problems caused by P pollution

56 56 57

P in ground and surface waters

57

Eutrophication

59

vii

3.6

Practical solutions

60

Source management—optimisation of fertiliser P use

61

Animal feeds

62

P transfer management by soil conservation

63

4

Soil erosion

69

4.1

Introduction

69

4.2

Vulnerability of soil to erosion and degradation

69

4.3

4.4

5 5.1

Extent of erosion

69

Processes and products of erosion

71

Water erosion

72

Wind erosion

74

Agricultural land management practices and accelerated erosion

75

Environmental impacts of soil erosion Impacts of sediment on aquatic environments

77

Physical impacts

77

Chemical impacts

78

Impacts of sediments on land

80

Practical solutions to soil erosion

80

Assessment of erosion risk

81

Field and crop based erosion control strategies

81

Farm and catchment based erosion control strategies

84

Organic wastes

89

Introduction

89

Organic wastes and agriculture 5.2

77

Farm wastes

89 92

Changes in farming systems

92

Pollution risks from farm wastes

94

Ammonia in watercourses

95

Pathogens from farm wastes

97

Veterinary products in farm wastes

97

viii

Water pollution incidents 5.3

Non-agricultural organic soil amendments Sewage sludge or biosolids

5.4

5.5

5.6

6 6.1

99 99

Pollution risks from applying sewage sludge to agricultural land

10 1

N and P pollution from sewage sludge

10 4

Odours

10 5

Pathogens

10 5

Other non-agricultural organic wastes

10 6

Detrimental effects of applying non-agricultural wastes to land

107

Practical solutions

10 9

Farm waste handling, storage and disposal

110

Solid manures

110

Silage

111

Slurry

111

Dirty water

112

Good agricultural practice

113

Non-agricultural organic wastes

119

Avoiding pollution problems by potentially toxic elements in sewage sludge

120

Avoiding crop contamination by pathogens

121

Alternative technologies for farm waste treatment

123

Biogas production

123

Reed bed treatment

123

Gaseous emissions

131

Introduction

13 1

Agriculture and gaseous emissions 6.2

98

13 1

Ammonia emissions

133

Pollution problems

133

Sources of ammonia pollution

134

Practical solutions

13 5

Minimising ammonia emissions by good practice

13 5

ix

6.3

6.4

6.5

6.6

7 7.1

Methane emissions

136

Pollution problems

136

Sources

136

Practical solutions

137

Nitrous oxide emissions

139

Pollution problems

139

Sources of pollution

139

Practical solutions

141

Carbon dioxide emissions

142

Pollution problems

142

Sources of pollution

143

Practical solutions

143

Farm odours

14 4

Sources of pollution

14 4

Practical solutions

145

The location and design of new livestock units

147

Minimising odour nuisance by good management practice

148

Technologies for reducing odour nuisance

148

Pesticides

153

Introduction

153

Pesticide history

153

Pesticide use

154

7.2

Pesticides and their application

155

7.3

Mode of action

156

7.4

Insecticides

157

7.5

Organophosphates

158

Carbamates

158

Pyrethroids

159

Herbicides

159

Triazines

159

x

7.6

7.7

7.8

7.9

7.10

7.11

7.12

Phenoxyacetic acids

16 0

Carbamates

16 0

Fungicides

16 0

Traditional inorganics

161

Phenylamides

161

Carboxamides or oxathins

161

The causes of pesticide pollution

161

Pesticides as pollutants

161

Resistance and resurgence

162

Behaviour and fate of pesticides in the environment

16 4

Persistence

16 4

The fate of pesticides in soil

165

Loss of pesticides from soil

167

Environmental problems caused by pesticide residues

168

Pesticides in surface and ground waters

168

Effects on non-farmed organisms

170

Practical solutions

174

Good practice

174

User training

175

Planning and preparation

175

Working with pesticides

176

Disposal of waste

177

Keeping records

177

Other considerations

177

New technology

178

Adjuvants

179

New formulations

180

Alternatives to synthetic pesticides

180

Naturally derived pesticides

180

Microbial pesticides

180

xi

8

Non-chemical weed control

181

Non-chemical pest control

183

Genetic modification

193

8.1

Introduction

193

8.2

Risks and benefits

19 6

Herbicide-resistant crops

19 6

Insect-resistant crops

2 02

8.3

Assessing environmental risk

20 6

8.4

The future for GMOs

20 8

Policy strategies for reducing pollution

213

9.1

Introduction

213

9.2

The emergence of contemporary agri-environmental policy

213

9

9.3

9.4

9.5

Agri-environment regulation 2078/92

214

Rural development regulation 1297/1999

216

Policy instruments, strategies and implementation

216

Policy strategies

217

Policy instruments

217

Implementation and dissemination

2 20

Learning from experience

221

Policies for reducing nitrate pollution

221

Nitrate pollution control policies in the UK and Europe

223

Policies for reducing pesticide pollution

225

Pesticide reduction programmes in Europe

2 26

Future directions

2 29

Adoption of sustainable farming practices

2 29

Legislative changes

230

Author index

23 5

Subject index

237

Figur es

1.1 Annual use of nitrogen fertiliser in the UK 1.2 Water pollution in ciden ts recorded within a ran ge of econ om ic sector s 1.3 Num ber of agricultural pollution incidents recorded between 1979 and 1999 in England a nd Wales 2.1 The agricultural N cycle 2.2 Relative yield from the addition of N 2.3 Per cen tage of tota l N available to crop after application of m an ur e 2.4 Variation in soil nitrate un der a cereal cr op 2.5 Nitrogen fertiliser use in the UK 2.6 Mean concentration of nitrate in river waters in Easter n England 2.7 The change in mean nitr ate concentrations in ground waters in Eastern UK 2.8 Measured N inputs, outputs an d sur pluses for two types of dair y farm 2.9 Location of nitr ate vulnerable zones in England a nd Wales 2.10 Variation in nitra te loss from a pplication of m anur es 3.1 Agricultural P cycle 3.2 The contribution of farm ing types to the UK annual P sur plus 3.3 Transfers and tr ansformations of P in soil 3.4 Influence of pH on the retention of P 3.5 P in puts from n on -point sources and m ain hydrological flow pathways 3.6 The enr ichm ent ratios of sedim ent run-off as a fun ction of fer tiliser P application 3.7 The relationship between P input and biodiversity in fr eshwater systems 3.8 The potential cycling of P between various form s in aquatic system s 3.9 Nutrient a nd tran sport m anagem ent option s for the control of P loss 4.1 Rill erosion in a winter cer eal crop 4.2 Soils at r isk of wa ter erosion in England 4.3 Catchm ent based soil conservation schem e on cultivated land 5.1 Deep litter housing system 5.2 Distribution of total pollution in ciden ts from organic waste in 199 9 5.3 The fa te of sewage sludge in the UK in 1996/ 7 5.4 The change in selected m etal content of sewage sludges 5.5 Slurry/ sludge in jection 5.6 Manur e spreading 5.7 The effect of adding organic residues with high C/ N ratio 5.8 A typical walled silo 5.9 An above-groun d cir cular slurry store 5.10 A weeping wall slurry store 5.11 An earth-banked store suitable for solid, liquid or sem i-liquid wastes

2 3 5 14 16 17 20 23 25 25 26 30 32 43 45 49 51 52 54 58 59 63 72 73 85 93 10 0 102 102 10 6 107 10 9 112 112 113 113

xiii

5.12 Cut out colour key to aid identification of high risk areas 6.1 Cumulative N2 O em ission s from gra ssland cut for silage 6.2 The carbon cycle 6.3 Spreadin g slurry usin g a vacuum tan ker 6.4 Altern atives to con ven tion al slurry spreaders 7.1 Developmen t of resistance 7.2 Aquatic pollution incidents from pesticides and sheep dips 7.3 Num ber of grey partridges 1966 to 1998 8.1 Worldwide use of genetically m odified crops 8.2 Assessm ent of potentia l for tra nsgene flow

116 141 142 14 6 14 9 163 16 9 174 19 4 20 1

Tables

1.1 Changes in agr icultural pr actice in the UK between 1930 and 1990 2.1 Typica l nutrient rem oval in winter wheat 2.2 Nitrogen containing chemicals used in fer tiliser s 3.1 Phosphorus con tent of plant parts 3.2 Com m on P containing fer tilisers 3.3 Effect of carbon a nd phosphorus ratios on organ ic P transforma tion s 3.4 Distance moved by ma cronutrient ions by diffusion 3.5 Fertiliser P r ecom mendations for cereal crops in the UK 3.6 Factors r espon sible for con trollin g P loss by erosion and lea chin g 3.7 Some pr actical m anagement techniq ues for the reduction of P loss 4.1 Typica l erosion rates under natural vegetation and cultivated land 4.2 Land m anagement practices which increase vulner ability of soils to erosion 4.3 Physical impacts of increased sedim ent load on aq uatic organism s 4.4 Physical impacts of sedim ent transport and deposition 4.5 Erosion risk classes for areas of England and Wales 4.6 Field/ soil er osion r isk definitions 4.7 Recom m en ded croppin g strategy for high an d very high risk sites 5.1 Exam ples of am ounts of excreta pr oduced by livestock 5.2 Value of organ ic m anures produced in the UK 5.3 Problems caused through or ganic waste application for crops a nd livestock 5.4 Char acter istics of silage effluent 5.5 Relative toleran ce to oxygen depletion by river or gan ism s 5.6 BOD of farm wastes 5.7 Chem ical com position of farm yard slur ry 5.8 Classification of water quality based on suitability as fish ery 5.9 Typica l N and P content of sewage sludges 5.10 Gener al characteristics of industr ia l wastes applied to agricultural land 5.11 ADAS far m waste managem ent plan 5.12 Areas needed to spread slurry during m onths when livestock are housed 5.13 Perm issible total m etal concentrations in sludge amended soil in EC a nd US 5.14 The safe sludge m atrix 6.1 Changes in soil carbon and nitrogen under different cultivation practices 6.2 Points of good pra ctice for reducin g odour emission s from livestock production systems 7.1 Global agrochem ical usage an d market va lue in 1999 7.2 Com m on pesticides and their tar get organisms 7.3 Gener al properties of som e com m on insecticides 7.4 Gener al properties of som e com m on her bicides

1 12 15 42 46 47 50 50 53 65 71 75 77 80 82 82 82 90 91 91 94 95 95 98 10 0 10 1 10 8 115 116 12 1 122 143 148 154 155 158 159

xv

7.5 7.6 7.7 7.8 7.9 8.1 8 .2 8.3 8.4 8.5 8.6 9.1

Gener al properties of som e com m on fungicides Proportion s of pesticide applied foun d in drain age water Pesticides detected in ground waters in the UK Precaution s n eeded by pesticide application m ethods Wind speed conditions and the suitability of pesticide spraying Potential hazards a ssocia ted with GM cr ops Potential benefits and risks of transgenic-resista nt crops Herbicide applied in Roundup Ready and conventional soybean crops Yields and yield dr ag US crop acr eage planted with Bt crops Envir onm ental indicators m easured during UK far m scale evaluations National pesticide reduction targets in Sweden, Denm ark and the Netherlands

16 0 168 171 176 177 197 19 9 19 9 20 3 20 4 20 7 227

Acknowledgements

The authors would like to thank the following for the provision of m aterial, helpful suggestions and com men t: Tom Misselbrook, Institute of Grassland and Environmen tal Science, North Wyke Research Station. Helen Stokes, Matt Lobley and Mick Fuller at Seale Hayne, Departm ent of Ecology, Lithuanian Univer sity of Agr iculture, Noreikiskes. Environment Agency (North West) for data for case studies: J eremy J oseph, J ohn Quinton, National Soil Resources Institute, Cranfield Un iver sity (Figure 4.1) Ca m Grant and Brian William s, Department of Soil Water, Adelaide Univer sity. Cr own copyr ight m aterial is reproduced under Class Licence Num ber C02P0 0 0 0 0 70 with the perm ission of the Controller of H MSO and the Queen’s Printer of Scotland. Finally, the authors express person al than ks to their respective fam ilies for their en cour agem en t an d support throughout the pr oject.

1 Agriculture and Pollution

1.1 SETTING THE SCENE During the latter half of the twen tieth centur y, the globa l human population doubled from less than 30 0 0 m illion to 60 0 0 million. As the global population in cr eases, demand for food contin ues to rise. This leads to the intensification of agr iculture which in turn places increasing demands on the natur al envir onm ent (Brown et al., 20 0 0 ). Agricultur e is of fun dam ental im portance to any n ationa l econ omy a nd the lifeblood of r ural com munities throughout the world. It occupies 35% of the wor ld’s lan d surface with 11% un der dir ect cultivation and 24% m anaged as per manent pasture (UNEP, 1992). In the United Kingdom m or e than 76% of land is under agricultural production (MAFF, 20 00a). The first evidence of agricultural activity in the UK can be traced back to 50 0 0 BC (Reed, 1990 ), although our book is concer ned with changes occurr ing only in the last 70 year s or so. The UK agricultural industry has under gone a m ajor revolution sin ce the 1930 s; progress in an imal an d crop br eedin g, the availability of pesticides and fertilisers, and ever-advancing technology has resulted in a substantial increase in productivity an d levels of na tional self-sufficien cy. Successive gover nm ent policies, notably the farm support m easures of the Agriculture Act 1947 and adher ence to Eur opean policy (Box 1.1) have provided UK far mers with m arkets for their pr oducts and a price structure that has encouraged agricultural intensification (MAFF, 1995, 20 00a). This drive towards intensive pr oduction (Table 1.1) has led to a m ajor increase in the use of agrochemicals like fertilisers (Figur e 1.1) an d pesticides. In parallel, there have been man y technological advances adopted by the agriculture industr y in recent years, such as the genetic m odification of crop plan ts and precision farm in g. Table 1.1 Chan ges in agr icultur al pr actice in th e UK between 1930 an d 199 0 (Ed ward s an d With er s, 19 98).

2

CH APTER 1

Figure 1.1 Ann ual use of n itrogen fertiliser in th e UK (‘0 0 0 ’s of ton n es) (Win ter , 1996).

The area of land affected by agricultur e, its reliance on natura l processes and the use of technology to intensify production results in a unique system both econom ically and ecologically with the following four key characteristics (OECD, 1997): • Economic viability of agr iculture is influenced gre atly by the natural environm ent. Productivity depends upon factors such as clim ate, soil fertility and water supply. • Agricultur al activities affect the quality of th eenvironment. Crops and livestock for m part of the agroecosystem , utilising na tural resources for growth. En vironmen tal benefits such as the m aintenance of traditional landscapes m ay be apparent, but eq ually there m ay be costs such as deterioration in soil, water an d air q uality or the loss of ha bitats importan t for conservation . • The r elationship between agricultural activity a ndthe environm ent is complex and site-specific. In teracting factor s include the physical, chemical an d biologica l attributes of the local environm ent, the mix of far m enterprises, managem ent practices, a nd the production technologies adopted. • Cultur al and political influences affect the agricu lture impacts on the environm ent. Most developed nations support food production by public subsidy and governm ent intervention (OECD, 1996; MAFF, 20 00a). Such measures in evitably affect the level of food production , its location and m anagement.

BOX 1.1 THE EUROPEAN UNION AND AGRICULTURAL INTENSIFICATION

Many of the changes in UK agricultur e have been influenced by initiatives within the EU. The Com mon Agricultural Policy (CAP) was conceived by the original m embers of the European Com m un ity in order to encoura ge food production at a tim e when food shortage was still a recent m em ory. The original objectives contained in the 19 57 Treaty of Rome included increasing productivity, sta bilising m arkets, ensuring r easonable food pr ices for consumers, and maintaining a fair standar d of living for farm ing com m unities. The plentiful supply of food and the extent of EU exports are a testam ent to its success. However, the CAP has been criticised because it created an econ omic clim ate in which food pr oduction was encouraged at the expen se of other consideration s such as the

AGRICULTURE AND POLLUTI ON

3

environm ent. Food surpluses were created that were dealt with either by being destroyed, or by expor t with subsidies (damaging international tr ade and a ffecting agriculture in d eveloping cou n tr ies). Th e CAP has also rein for ced th e post-war tren d in m any Eu ropean coun tries towards the expan sion, in tensification and sp ecialisation of crop and livestock pr oduct ion. Th e p rice stru ctur e of m arkets redu ced th e econ om ic risks associat ed wit h specialisation, stim ulated the use of high er in pu ts t o incr ea se outpu t, and en cour aged th e expan sion of produ ction in to p reviously un cult ivated ar eas. Farm s grew lar ger, m or e cap ital-inten sive an d eager to ad op t fu rth er im pr oved t ech n ologies with which to in tensify p roduction (Clu nies-Ross an d H ild yard, 19 92). Escalating fin ancial costs, en viron m en tal concer ns, t he ap proach in g en largem ent of th e EU an d th e dist or tin g effect of th e CAP on world tr ade con tin ue to m ove t owards fun dam ental r eform . The latest ‘Agen da 20 0 0 ’ refor ms will r esult in the in tern al m arket pr ices for cer eal, beef an d da iry pr od ucts bein g reduced t o th e level of global m arket p rices by 20 0 6. In add ition , th e Rur al Develop men t Regulation (th e so-called ‘secon d pillar’ of the CAP) pr ovides fur ther evidence of a long-t er m desire to sh ift supp or t fr om pr od uction to ru ral developm en t an d en vir on m en tal m an agemen t (although close to 90 % of the CAP bud get is still devoted to com m od ity r egim e su pport an d com pensation paym ent s).

This book focuses upon the UK, and is illustrated prima rily with Eur opean examples. The need to identify, under stand and solve the problem s caused by agricultural activity is clear fr om statistics that show that it can be a m ajor cause of pollution (Figure 1.2). Seven key types of agricultural pollution : n itrates, phosphates, sedim ent loss, organ ic wastes, gaseous em ission s, pesticides and genetic m odification are included. We describe why pollution m ay occur a nd how such problems can be overcome. Before we investigate these topics individually, the underlying causes of agricultural pollution and responses to them are considered. The final chapter reviews the role of agricultural policy, and reflects on how it may be used as a tool to deliver environmentally sensitive agricultural system s.

Figure 1.2 Water pollu tion in cid en ts r ecor ded with in a r ange of econ om ic sectors (Envir onm ent Agen cy, 20 0 1).

Defining agricultural pollution At its most inclusive, the term pollution can be used to describe a ll unwanted envir onm ental effects of hum an activity. The Oxford English Dictionar y defines pollution as ‘the presence in the environm ent, or the introduction into it, of products of human activity which have harm ful or

4

CHAPTER 1

objectionable effects’. This definition could include ‘visual’ pollution such as unsightly farm buildings (Conway and Pretty, 1991) but we use a definition whereby a pollutant is a substance, including those: • deliberately introduced into the environment (e.g. pesticides, fertilisers, genetically modified crops and sewage sludge); • pr oduced by agricultur al processes as wastes (e.g.silage effluent and livestock slurry); • pr oduced by the en hancem en t of natural processes ni the course of agricultural activity (e.g. increased nitrous oxide em issions fr om cultivated soils or soil erosion). A further distinction between a ‘contaminant’, which is, an y substance introduced by human activity into the envir onm ent with no evidence of ha rm, a nd a ‘pollutant’, which is causing dam age or harm may be m ade (Crathor ne et al., 1996). Substa nces enter ing the environment m ay only cause pollution if they a re: • pr esent in excessive qua ntities—‘concentration effe cts’; • in the ‘wrong place a t the wrong time’; • transform ed into h armful ‘secondary pollutants’ asa result of biological or chem ical processes. Pollutants from agr icultural systems have the potential to have a m ajor impact, ranging fr om the im mediate on-farm environm ent to food products a t the point of sale, and from local groundwater sources to the stratosphere. The im pacts of agricultura l pollution can be categorised into the followin g areas (OECD, 1997): • quality of natur al resour ces, notably the physica,l biologica l and chemical condition of soil, water a nd air; • composition a nd functioning of ter restrial, aquatic and m arine ecosystem s, including issues of biodiversity and habitat qua lity; • other environmental im pacts such a s public nuisance caused by odours from livestock pr oduction. The occurrence of pollution in agricultural system s is well documented (Baldock et al., 1996; OECD, 199 7; Edwards an d Withers, 1998; Ish erwood, 20 0 0 ; EFMA, 2 0 0 1) an d the precise effect on the natural environment is m ediated by the interaction between environm ental factors and farm m ana gem ent. The m ove towar ds m ore intensive far ming m ethods has led to a mar ked increase in the number of pollution incidents recor ded (Figur e 1.3) and this has driven the developm en t of pollution control strategies designed to reverse this trend. This book addresses agr icultural pollution although the full environm enta l and socio-econom ic im pact of post-war agriculture also includes declines in farm la nd wildlife, the loss of tra ditional landscapes, rural depopulation, animal welfare, and health related issues (Baldock et al., 1996; Corpet, 1996; DoE, 1996; J on es, 1999; MAFF, 2000 a, 20 0 1). In this book we consider the following environm ental issues:

AGRI CULTURE AND POLLUTION

5

Figure 1.3 Num ber of agr icultur al pollu tion in cid en ts r ecor ded between 1979 an d 199 9 in En glan d a nd Wales (NRA, 199 2; En vir on m en t Agency, 2 0 0 1).

• con tamin ation of groun d and surface water by n itr ates (Chapter 2), phosphates (Cha pter 3), organ ic wastes (Chapter 5) and pesticides (Chapter 7), all of which can disrupt aquatic and m arine ecosystem s a nd have significant effects on drinking water quality; • disruption of agr oecosystems by pesticides (Chapter 7) or genetically m odified crops (Chapter 8 ), including flora and fauna in crops a nd sem i-natural habitats; • the effect of sedim ent loss due to erosion (Chapter 4); • con tamin ation of soil and crops by metals, organ ic m icropollutan ts and pathogens fr om the application of livestock wastes and non-agricultural industrial wastes (Chapter 5); • atm ospheric contam ination by a mm onia, m ethane andnitrous oxide which play various roles in acid r ain production, global war ming and ozone depletion (Chapter 6 ). 1.2 SOLVING THE PROBLEM Monitoring impacts Assessment of the environmental consequences of agricultural pollution is needed before solution s to the problem s can be formulated. Agriculture is only one of m an y econ omic activities that cause pollution a nd so its ‘share of r esponsibility’ for any envir onm ental im pact incurred m ust be identified before control priorities ar e set. For example, pesticide pollution is not just caused by agriculture. Most of the herbicides com m on ly detected in water also have n on agricultural uses with local authorities and public utilities regular ly spraying public parks, roadside verges, r ailway lines and playing fields to control weeds. Until their prohibition from nonagricultural use in 1992, this included the use of atrazine and simazine, two of the m ost comm only detected h erbicides in drinking water . Further investigation of the effects of agr icultural pollution is com plicated by: • The com plexity of contamination and pollution pathways: for example, pesticides m ay affect wildlife by direct contact with the chem ical or its breakdown pr oducts, indirectly by contamination of food sources, or by the destruction of habitats and resour ces upon which species depen d.

6

CHAPTER 1

• Spatial and tem poral variations in the occurrenceof har m or damage: There are essentially two m ain sources of pollution from agriculture: ‘poin t’, in volvin g discrete an d easily iden tifiable incidents such as leakage from a slurry store or silage clam p, and ‘diffuse’, involving the lea ching and r un-off of pollutants from large areas of agricultural land to ground a nd sur face water . • The un predictable nature of pollution due to variations caused by clima te, soil type an d other environm ental factors. The challenge is to identify the effects of pollution and then translate this understanding into appropr ia te action to manage the pr oblem . Comm on responses (OECD, 1997) include: • government action thr ough changes in policy and law, research and development, tr aining and inform ation pr ogram mes and economic instrum ents such as financial subsidies and taxes; • responses by the agr icultura l industry, such as the voluntary adoption of new quality standards and the im position of stricter quality; • m odified behaviour by farm er s, including changes ni the use of agr ochemical inputs and other farm m anagement practices; • consumer reactions expressed via changing patternsof purchase and consum ption. Pollution m anagem ent can be achieved in two ways. Fir stly, we can attempt to ‘cure’ the problem by acting against the pollutants them selves (e.g. by wa ter treatm ent). Secondly, we m ay ‘pr event’ th e problem by addressing the underlying causes of pollution (e.g. by encouraging the a doption of alter na tive agricultural practices tha t are less polluting). In both cases, the actions that can be taken are technical; whether they ar e adopted depends upon the presence of appropr iate knowledge, effective legal regulations and adeq uate financial incentives (Conway and Pretty, 1991; MAFF, 1998a , b). This book includes solutions that a re essentially far m-based, m odifying m anagem ent practices and business decisions via: • • • •

the availability of n ew technologies; the provision of inform ation and advice to encourage ‘good agricultur al practice’; statutory controls and regulations that enforce change; agri-environm ental policy and the pr ovision of financia l incentives in the form of taxes or subsidies; • the emergence of alternative agricultur al system s with a m arket linkage offer ing financial incentives in the form of price.

For each pollutant we describe current pr actical solutions to contr ol their impact by considering new technology, good agricultural practice and regulation. We also descr ibe the developm ent and im plem entation of contem porary agri-environmental policy in the UK and western Europe, including the em er gence of alternative a gricultural systems such as Integrated Crop Managem ent (Chapter 9).

AGRICULTURE AND POLLUTI ON

7

Cost—the driver for pollution control It has been recogn ised by govern men t an d in dustry that the ‘extern al costs’ or ‘extern alities’ caused by economic activity should be considered when m aking decisions r egar ding pollution control (DoE, 1994). An externality is a side effect (or by-pr oduct) of agr icultura l practice which is unpriced within the econom y of the far ming system but which nonetheless incur s a cost for som eone (or som ethin g) else by reducing their profit or welfare (Han ley, 1991). These extern alities include: • the depr eciation of natural capita l thr ough the use of non-renewable natural resources such as oil and coal or the loss of other natural assets such as biodiversity and landscape; • declines in per sonal or collective ‘welfare’ suchas public health; • the cost of environm enta l degr adation including the cost of clea ning up dam age; • the cost of defensive expenditure including the cost of pr eventa tive action to avoid environ m en tal dam age. The principal challenge when accounting for agricultural externalities is not in their identification, but the a ssignment of monetar y value. Costs m ay be financial (e.g. incurred in water treatment) or econom ic (e.g. due to the loss of a landscape feature valued by people). Exam ples showing how costs ma y be evaluated are given in Box 1.2. It is a cha llenge for agricultur al policy makers to facilitate the tr ansition of farm practice to those which are environmentally sensitive, allowing farmers to m odify their farm ing practices whilst m aintaining the econom ic viability of their businesses.

BOX 1.2 THE COSTS OF AGRICULTURAL POLLUTION

Nitrates in drinking water

Th e EC m axim u m ad missib le con centr ation (MAC) of n itr ate in hu m an dr in king wat er is 50 m gl −1 (50 p arts per m illion). H owever, an in creasin g nu m ber of r aw water sour ces in th e UK exceed t his con cen tr ation and water su pply com pa nies ha ve been for ced to in trodu ce treatm en t pr ogr am m es (DWI, 20 0 0 ). Treatm ent option s (Croll an d Hayes, 19 88 ) in clu de blend ing high n itr ate cont en t water with th at which is less polluted, biological den itr ification , ion exchan ge, rever se osmosis an d electrodialysis. Estim ates of t he in vest men t in cu rred in the installation of denitrification equip men t ra nges fr om £ 148 m illion to £ 20 0 million (DoE, 198 6; Ofwat, 19 92), with an n ual ru nn in g costs of at least £ 10 million per year . Phosphates in waters Th e t ota l pollut ing load s fr om sewage tr eat men ts works in t he UK ha s fallen by between 30 an d 40 % dur in g the 1990 s a nd p hosphat e loads specifically by 37%. Th is has been due to im proved sewage t reatm ent with in vest m en t of £ 250 m illion com bined with a redu ction of phosph ate u sa ge in detergen ts (En vironm en t Agen cy, 20 0 0 a, b). Agr icu ltur e is a diffu se sou rce, accou nting for over 50 % of th e Eur op e-wide ph osp hate in sur fa ce water s (En vir on m ent Agen cy, 20 0 0 b). Estim at ed an nu al costs for the removal of ph osph ate fr om surface waters fr om agricult ure in th e UK a re in the or der of £ 55 million (ENDS, 20 0 0 b). Organic wastes

8

CHAPTER 1

Th e produ ction , storage an d disposal of an im al waste an d sila ge efflu en t can present sign ifican t risks to t he aqu atic life of st ream s and river s, althou gh t he total nu m ber of farm p ollution in cid en ts du e t o organ ic wastes is n ow declin in g in th e UK. Noneth eless, the Envir onm ent Agen cy still sp en ds app roxim ately £ 5 m illion per year on su rveying and corr ectin g r iver pollution in cid en ts caused by agr icultur e (Nation al Audit Office, 19 95). Gaseous emissions In th e UK, agricultu re is r espon sible for ap proxim ately 8 % of all gr een house gas em issions, in par ticu lar n it rous oxide an d m etha ne (MAFF, 20 0 0 a). Predictin g th e effects of increased global war m in g on clim at e chan ge in the UK over th e next 50 year s is frau ght with difficu lty, especially wit h regar d to chan ges in rain fall, storm in ess an d extrem e events such as drough t (MAFF, 20 0 0 b). The possible costs of clim ate ch an ge upon agr icult ura l pr od uction ar e also on ly pr ed iction s, wit h con siderable un certainty related to weed , p est an d disease outbr eaks, global com m odit y p rice effects an d ch anges in yield and quality of ar able crops. Never theless, th e comb ined cost of th e detrim en tal m an ifestat ions of nit rous oxide and m et han e em issions from agricult ure ha ve been estim ated to be well over £ 1,0 0 0 m illion a year (ENDS, 20 0 0 a). Pesticides in drinking water Und er th e 19 80 EC Direct ive on Drinkin g Water Quality, th e MAC for any p esticide in drinkin g wat er , ir respective of its toxicity, is 0 .1µg l−1 (0 .1 pa rt per billion ). This is ackn owledged as on e of the m ost stringen t p esticide stan dar ds in the wor ld a nd is ar guably ver y difficult to enforce. Since the early 199 0 s an in creasin g n um ber of UK groun dwaters pr ovidin g sources of d rin king water have pesticide concen trat ions in excess of the EC stan da rd an d it is cur rent ly estimat ed that 8% of all th ose wat er s tested exceed t he lim it set (MAFF, 20 0 0 a). This h as triggered h uge in vest men t in tr eat men t p lant s by wa ter com pan ies in an at temp t t o reduce pest icid e levels in the water su pplied to th eir custom ers (ENDS, 20 0 0 b). Removal of tr ace p esticides fr om wat er is com plex and expen sive; tech nology availa ble in clud es gr anu lar a ctivated carbon (GAC), ozon e treatm en t an d an activated car bon san dwich between layers of slow san d filters (ENDS, 20 0 1). Estim at es of th e t ota l capital in vest men t un dert aken by UK wat er com pan ies on pesticide treatm en t plan ts ran ge from £ 8 0 0 m illion to £ 10 0 0 m illion (Ofwat, 1992). An nu al run n in g cost s ar e also expected to have risen by abou t 10 % of capital expen dit ure i.e. £ 80 m illion to £ 10 0 m illion per year .

The control of pollution can be viewed in the broader context of ‘susta inable development’. This concept is founded upon the idea that the environm ent is a finite entity that is incapable of absorbing the im pact of everything released into it or rem oved fr om it. In other words, the environm ent only has a certain ‘capacity’ to accom m odate the im pact of hum an activity. As people depen d upon the en viron m en t for their survival an d quality of life they have a duty of care to look after it for the benefit of them selves and future generations (J acobs, 1991). Agr iculture occupies a central place within the sustainability debate since it occupies m ore land tha n any other econom ic activity. To m eet the challenge of sustainability, a griculture m ust reduce its envir onm ental im pact by minimising or eliminating pollution. In doing so it m ust also rem ain econom ically viable in order to survive an d play its part in the life an d econ om y of rural ar eas. REFERENCES Baldock, D., Bish op, K., Mit chell, K. an d Ph illips, A. (1996) Growing Greener: Sustainable Agriculture in the UK. Coun cil for the Protection of Ru ral En glan d and Wor ld Wid e Fun d for Nat ure, Lond on .

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Brown , L., Flavin , C. and Fren ch, H . (20 0 0 ) State of the World 2000. Wor ldwat ch Institute, 262p p. Clun ies-Ross, T. an d Hildyar d, N. (1992) The Politics of Industrial Agriculture. Earth scan Pu blications, Lond on . Con way, G. an d Pr et ty, J . (199 1) Unwelcome Harvest: Agriculture and Pollution. Ear thscan Pu blications, Lond on . Cor pet, D.E. (1996) Micr obiological h azards for hum an s of ant im icr obial gr owth p rom ot er use in an im al pr od uction . Revue de Médecine Vétérinaire 147, 8 51– 8 62. Crath or ne, B., Dobbs, A.J . and Rees, Y. (19 96) Chem ical Pollu tion of the Aqu atic Envir onm ent by Prior ity Pollut ant s an d its Con trol. In : Pollution, Causes, Effects and Control (Ed . R.M.H arr ison). 3r d Edit ion. Th e Royal Societ y of Ch em ist ry, Camb rid ge, pp. 1– 25. Croll, B. and H ayes, C. (198 8 ) Nitr ate an d water supp lies in the Un ited Kingd om . Environmental Pollution 50, 16 3– 18 7. DoE (198 6) Nitrate in Water: a Report by the Nitrate Co-ordination Group. Depa rtm ent of En viron men t Pollut ion Paper No. 26, H MSO, Lond on . DoE (199 4) Sustainable Development: the UK Strategy. Depart m en t of Environm en t Com man d Paper 2426, H MSO, Lond on . DoE (1996) UK Indicators of Sustainable Development. HMSO, Lon don . DWI (20 0 0 ) Overview of Water Quality in England and Wales, Drinking Water 1999. Drin king Water In sp ectorate, Dep artm ent of th e En vir on m en t, Tran spor t an d the Region s, Lon don. Edwar ds, A.C. an d Wither s, P.J .A. (1998 ) Soil ph osp horus m anagem en t an d wat er qu ality: a UK perspective. Soil Use and Management 14, 124– 130 . EFMA (20 0 1) Sustainable Soil Management: an Achievable Goal. European Fert ilizer Man ufactu rers Association , Brussels. ENDS (20 0 0 a ) The diffuse pollution challenge. ENDS Report 310 (Novem ber ). ENDS (20 0 0 b) Farming’s environmental costs top £1.5 billion per year says Agency. ENDS Rep or t 309 (Oct ober). ENDS (20 0 1) Water firms urged to end chlorine addition. ENDS Report 314 (Mar ch). Environm en t Agency (20 0 0 a) Achieving the Quality, the Environment Agency’s Views of the Benefits to the Environment of Water Company Investment over the Next Five Years. Envir on m ent Agen cy, Dep artm ent of th e En viron m en t, Tran sp or t an d the Region s, Lon don. Environm en t Agen cy (20 0 0 b ) Aquatic Eutrophication in England and Wales: a Management Strategy. Environm en t Agen cy, Depar tm ent of the En viron m en t, Tr an sp or t an d the Regions. Environm en t Agen cy (20 0 1) Water Pollution Incidents 1999. URL: h ttp:/ / www.en viron m en t-agency.gov.uk. H anley, N. (Ed.) (1991) Farming and the Countryside: An Economic Analysis of External Costs and Benefits. CAB Int er nat ional, Wallingfor d. Isherwood, K.F. (20 0 0 ) Fertilizer Use and the Environment. (Revised Ed ition ). In ter n ational Fertilizer In dustr y Association , Paris. J acobs, M. (1991) The Green Economy. Plu to Press, Lon don . J on es, D.L. (199 9) Escherichia coli O157 in th e envir on m ent. Soil Use and Management 15, 76– 8 3. MAFF (1995) European Agriculture: the Case for Radical Reform. Min istr y of Agricu ltu re, Fisher ies an d Food , Lon don . MAFF (199 8a) Code of Good Agricultural Practice for the Protection of Water. Revised 1998 . Min istry of Agr icultur e, Fish er ies an d Food , Lon don. MAFF (1998 b) Guidelines for Farmers in Nitrate Vulnerable Zones. Minist ry of Agr icu ltur e, Fisheries an d Food , Lon don . MAFF (20 0 0 a) Towards Sustainable Agriculture: Pilot set of Indicators. Minist ry of Agricu ltur e, Fisheries an d Food, Lond on . MAFF (20 0 0 b) Climate Change and Agriculture in the United Kingdom. Min istr y of Agr icu ltur e, Fisheries an d Foods, Lon don .

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MAFF (20 0 1) Foot and Mouth Disease: Public Information Factsheet 1. Min istr y of Agricultu re, Fisheries an d Food, Lond on . Nation al Au dit Office (199 5) National Rivers Authority: River Pollution From Farms in England. H ouse of Com m ons Pa per 235. HMSO, Lon don . NRA (1992) The Influence of Agriculture on the Quality of Natural Waters in England and Wales. Nation al Rivers Aut hority, Bristol. OECD (1996) Agricultural Policies, Markets and Trade in OECD Countries—Monitoring and Evaluation 1996. Orga nisation for Econ om ic Co-operat ion and Developm ent, Paris. OECD (1997) Environmental Indicators for Agriculture. Organisat ion for Econom ic Co-op er ation an d Developm ent, Paris. Ofwat (19 92) The Cost of Quality—a Strategic Assessment of the Prospects for Future Water Bills. Ofwat, Birm ingh am . Reed, M. (1990 ) The Landscape of Britain: from the Beginnings to 1914. Routledge, Lon don. UNEP (1992) The World Environment 1972–1992: Two Decades of Challenge. Cha pm an and H all (on beh alf of th e Un ited Nations En vir on m ent Program m e), Lon don . Wint er , M. (199 6) Rural Politics: Policies for Agriculture, Forestry and the Environment. Routledge, Lond on .

2 Nitrates and Nitrogen Loss

2.1 INTRODUCTION The present structure and output of agricultur al system s could not be m aintained without the advent and widespr ead use of synthetic or mineral fertiliser s. Of the m ajor plant nutr ients N, not only provides the grea test responses in crop yield fr om fertiliser addition but is also the m ost readily lost from the agroecosystem . In parallel with increased agricultural production over the last 50 years has been the increase in nitrate (NO 3−) con cen tration s in rivers, lakes an d undergroun d aq uifers. Ther e is str on g eviden ce to suggest that this is due to pollution from agriculture. Furthermore, this loss of N fr om agriculture as lea chate in the form of NO 3 −, but also to a lesser extent as gaseous form s of N or erosion a s N associations with soil particles, repr esents an econom ic shortfall, in that the applied N is not being utilised for food production. This chapter tackles three m ajor questions: • Why ha ve NO3 − levels increased in waters? • Is this increase harm ful to the environment? • What can be done to reduce and stop fur ther incr eases in NO 3− pollution an d losses fr om agricultural systems? Nitrogen and agriculture Nitrogen is an essential constituent of all nucleic acids, am ino acids and proteins, and therefore fun dam enta l to the reproduction and growth of all organ ism s. In a gen eral in troduction to the global N cycle, J enkinson (1990 ) estimated that in 1990 the world’s hum an population conta ined a total 10 m illion tonnes of N. Although this is sm all compared to the total am ounts of N in the atm osphere (3.9×10 9 million tonnes), soil or ganic m atter (1.5×10 5 m illion tonnes) or pla nts (1. 5×10 4 million tonnes), the hum an population is increasing and is inextricably linked via agriculture and the food industry to the global N cycle (Schlesinger, 1997). Nitrogen occur s na turally in soils and is closely associated with soil organic m atter . H owever , it is the sim ple ionic form s of a mm onia (NH 4 + ) and NO3− which plants can easily absor b and utilise (Whitehead, 2 0 0 0 ). Of the elem en ts essen tial for pla nt gr owth, N is requir ed in the greatest quantity by agricultural crops (Table 2.1). The exceptions to this rule-of-thumb are those crops

12

CHAPTER 2

Table 2.1 Typ ical nu trien t r em oval (kg h a −1) in a h arvested wint er wh ea t cr op n utr ient offtake per tonn e of pla nt m aterial.

which form large un der groun d storage organ s, n otably potatoes, since these also requir e large quan tities of phosphorus (P) an d potassium (K). Nitrogen -con tain in g compoun ds are in volved in virtually a ll of the biochem istr y of the crop plant. This includes chlorophyll that is essential for photosynthesis, the nucleic acids in which the pa ttern for the pla nt’s growth an d developmen t is encoded, an d a var iety of plan t proteins ranging from lipopr otein membranes to enzym es such as ribulose 1, 5-biphosphate ca rboxylase-oxygenase (Rubisco) which plays a key role in the conversion of atm ospheric carbon dioxide (CO 2 ) into organic carbon durin g photosyn thesis. Nitrogen deficien cy is gen erally char acter ised by the yellowin g or chlorosis (i.e. loss of chlorophyll) of the lower leaves of cr op plan ts, star ting from the tip and extending to the whole lea f with increasing deficiency. In severe cases, the whole plant is stunted and the leaves rem ain sm all. While the tim ely application of r elatively sm all am ounts of additional NH 4 + and NO3 − to the soil will often relieve the symptom s of sickly-looking crops, m uch higher levels of min er al N are requir ed to sign ifican tly in crease crop yield. An abun dan t supply of min eral N in cr eases crop yield by influencing leaf ar ea in two distinct ways: •

By en couragin g the rapid growth of above-ground vegeta tion : This can have a n im portan t influence upon crop yield, since any increase in the size of the crop canopy promotes both the interception of sunlight and the absorption of CO2 ther eby increasing photosynthetic efficiency. • By prom oting the duration of the crop canopy: Nitro gen is very m obile within the plant and any shortage of N in youn g tissue is usually met by the mobilisation of N from the older leaves resulting in their chlorosis and eventual dea th. An abundant N supply avoids this problem and m aintains more leaves to carry on photosynthesising for longer.

In certain crops, a good supply of m ineral N may a lso help to im prove the quality, as well as the yield, of the final har vested produce. For exam ple, bread-m aking wheat var ieties need to contain at least 11– 12 % protein in or der to for m a sa tisfa ctory dough. This requires a sufficiently high N supply (late in the growing sea son) to achieve an N concentra tion in the grain of around 2.0% on a dry m atter (DM) basis. In other crops, increased supplies of m ineral N are not so welcome. The availability of too much N for pota to plants can produce too m an y over-large tubers, while barley grown for m alting needs gr ain with as m uch starch and as little protein content as possible (ideally not in excess of 1.6% N). Too much N may also reduce the sugar content of sugar beet (Isherwood, 20 0 0 ). Excessive am ount of N also pr oduces vegetative gr owth with large succulent thin-walled cells. This can cause two problem s: • the lea ves and stem s are more readily attacked byinsect pests and fungal diseases;

NITRATES AND NITROGEN LOSS

13

• the stem s ar e less mechanica lly strong and crops are subsequently pr one to ‘lodging’ i.e. being blown over in wet and win dy weather. Despite bein g the most im portan t crop nutrient in agr iculture, N is also an en vir onm ental pollutant ca using significan t ecological disturbance. This cha pter will consider the ‘leakage’ of N from a griculture into the wider environment in detail and discuss those factor s influencing the occur rence and behaviour of N within agricultural system s. 2.2 THE NITROGEN CYCLE Nitrogen is a transient nutr ient and the am ount available in the soil at any one tim e to meet the dem an ds of a growin g crop is the pr oduct of the com plex n etwork of physical, biologica l and chem ical pathways thr ough which the various for ms of N m ove: • into the soil (inputs); • within the soil (transform ations); and • out of the soil (losses). Together these pathways are kn own as the soil N cycle (Figure 2.1) and are an integr al par t of the over all cycling of N within nature. A full discussion of this cycle and its significance for agricultural production is beyon d the scope of this book, but it has been exten sively reviewed elsewhere (e.g. Wild, 1988; Powlson, 1993; Tisdale et al., 1993). Nitrogen can enter the soil N cycle in a num ber of ways and in different form s. Atm ospheric deposition, biological N 2 fixation, fertiliser s and anim al feeds effectively im port N from outside of the far m, whilst anim al m anures typically tra nsfer N fr om one part of the farm (e.g. where livestock which have been overwintering in sheds and barns) to another. Atmospheric deposition Agricultur al land r eceives sign ifican t quan tities of N via the deposition of N oxides an d NH 3 fr om the atmospher e. The m ajor source of N oxides in the atmosphere is fuel com bustion and the socalled NOx emissions from power stations and motor vehicles. Atmospheric NH 3 is derived fr om a num ber of sources including industrial em issions, coal burning, livestock wastes and other agricultural sources (Schlesinger, 1997) (Chapter 6 ). Since m ost atm ospheric N com pounds are highly soluble in water, deposition may occur in rainfa ll, although the dry deposition of gaseous and par ticulate ma terial may also occur. An estim ated 30– 50 kg N ha −1 year −1 are deposited on agricultur al land from the atm osphere in southern and easter n England (and the sam e for som e parts of Germ any). Of this, when deposited onto land in cereals, it is estimated that 5% is lea ched, 12% denitrified, 30% imm obilised and the rem ain ing taken off by the crop (Goulding et al., 1998). Total N deposition in intensive livestock pr oduction area of the Netherlands m ay range from 40– 80 kg N ha −1 year −1 (Whitehead, 20 00). For the farm er this ma y r epresen t a useful N input, but may be particularly detrim en tal in natural

14

CHAPTER 2

Figure 2.1 Th e agricultu ral N cycle (adap ted fr om Rowell, 1994 ).

and sem i-natural ecosystems (Box 2.1). Even in areas r em ote from intensive production agriculture deposition may still be in the region of 15 kg N ha −1 yea r −1 (Brady and Weil, 1999). Biological fixation Certain species of bacteria and algae are ca pable of reducing atm ospheric N to NH 3 . The m ost im portant agr icultura l example is the Rhizobia bacter ia that form a close symbiotic r elationship with leguminous crops such as peas, beans and clover. Legumes are mainly grown on the farm as either arable or fora ge crops. Arable legum es (e.g. field peas and beans) are grown and ha rvested to pr oduce a dr ied grain for inclusion as a pr otein source in the diet of farm anim als, while forage legum es (e.g. red and white clover ) are comm only grown mixed with grass, and are either used for livestock grazing or cut for hay and sila ge. Legume crops usually leave large quantities of high pr otein content crop residues which can m ake a significant contr ibution to levels of organic N in the soil and, upon decomposition and mineralisation, to soil m ineral N levels and the growth and yield of subsequent crops. Some legume crops are gr own specifically for incor poration into the soil as green manures although in r ecent tim es their use in conventional a griculture has dim inished (Parson s, 1984). In the UK crop rotations including legum inous cr ops form ed the traditional base of agr iculture for m any centur ies. One of the best known rotations in the 170 0 s was the ‘Norfolk four course’. This originally took the form of roots, barley, seed and wheat. The seed component of the course of the rotation was some form of legum e, notably a one-year red clover ley (som etim es with ryegrass) or an ar able legume crop. In som e par ts of the country, the 1-year seeds crop wa s extended into a short-term 2 -year ley, or into the m edium to long-term leys which form the basis of tra ditional ley/

NITRATES AND NI TROGEN LOSS

15

arable farm in g system s e.g. 4– 5 year forage legum e or grass/ clover ley followed by up to 3 years cereals (La ity, 1948). Biological N 2 fixation by legum es is highly var iable depending upon the num ber of a ctive nodules, their size and longevity, and the bacterial stra ins occupying them . These factors in turn are affected by the complex interaction of legum e species and cultiva r, cr op m anagem ent and conditions of gr owth (notably water availability and soil nutr ient status). In an extensive review of North Am erican work, La Rue and Patterson (1981) quoted estim ates of annua l N 2 fixation in the ran ge of 10 – 10 0 kg N ha −1 for arable legumes and 100– 250 kg N ha −1 for forage legum es. In Europe, productive grass-clover swards would be expected to fix between 10 0 a nd 30 0 kg N ha −1, without the addition of N fertiliser (Whitehead, 200 0). Fertilisers The industrial fixation of atm ospheric di-nitr ogen gas (N 2 ) is directly a nalogous to biological fixa tion since it also involves the r eduction of N 2 to NH 3 . During the com mercial ma nufacture of NH 3 (Haber -Bosch process), hydrogen and atm ospheric N 2 are combined at high temper ature (30 0 – 50 0 °C) and pressure (40 0 – 10 0 0 atm ospheres) in the presen ce of a catalyst. The NH 3 pr oduced m ay be used as a fertiliser m aterial itself (Table 2.2), but is more com mon ly processed to fertiliser ma terials such as am monium nitrate or ur ea (so-ca lled ‘straight’ N fertilisers), or monoand di-am monium phosphates used for the m anufa cture of ‘com pound’ NPK fertiliser s. Compared to the use of legum es as an N source, fertilisers directly supplem ent soil m inera l N levels with NH 4 + and/ or NO3 −, and therefore rapidly increase the a mount of N availa ble for crop uptake. As a guide (and assuming that the fertiliser is applied at an appropr iate tim e), the yield of a wheat crop m ay in cr ease by a m axim um of 24kg for each addition al kilogram of fertiliser N applied, up to where the r esponse begins to pla teau (Figure 2.2) (Mackenzie and Traureau, 1997). As application r ates are increased, a point of ‘optim um ’ application is reached at which the availability of extra N ceases to be worthwhile for crop growth and is thus surplus to requirem ent (Figure 2.2). Table 2.2 N-con tain in g chem icals an d m aterials com m on ly used in ‘st raight’ and ‘com poun d’ fertiliser s (Wh ite, 1997).

16

CHAPTER 2

Figure 2.2 Relative yield from th e add ition of a gr owth factor, su ch as N, as described by t he Mitsch er lich Equat ion (d ata fr om Tisdale et al., 1993).

Animal manures Of the N consum ed by livestock in the form of herbage and concentrate feeds, a relatively sma ll pr oportion is actually utilised for the production of meat or m ilk (Whitehea d et al., 1986). In stead, the majority (typically 70 % of the N consum ed by cattle an d 8 0 % consum ed by sheep) of N is excreted as dung and urine (Hayga rth et al., 1998). When cattle an d sheep ar e gr azin g this excreta is voided directly to the soil surface, but on ce the an im als a re housed in the win ter it accum ulates in yar ds and buildings as slurry and manure which needs to be stored and spr ead on the land a t an appropr ia te time (Chapter 5). The amount of N availa ble to crops following the surface application of manure va ries with a range of factors, including soil type, form and sour ce of m anure and tim e of year. Figure 2.3 shows the percentage of total N available to the next crop following th e winter and autumnal application of differ ent m anure types to a sandy soil. The usefulness of this type of inform ation to farm ers attempting to account for the fertiliser value of m anur es a nd r educe N leakage will become evident as the chapter continues (Box 2.2). Manure N com prises two m ajor fraction s of agron omic in terest, readily availa ble in organ ic N (mainly NH 4 + ) a nd organic N (Chapter 5) (MAFF, 200 1). Amm oniacal N is water soluble, com prising of urea and NH 4 + , and when the m anur e/ slurry is applied to the soil supplem ents soil m ineral N levels in a similar man n er to fertiliser N. The organ ic N fra ction m ust un dergo m ineralisa tion before being available for crop uptake. 2.3 NITROGEN TRANSFORMATIONS Within the soil N cycle, a num ber of important biological and chem ical transformations occur (Figure 2.1) which influence both the am ount of N available for crop uptake and that which is at risk of ‘leaking’ into the wider envir onm ent. Two of the most im portant of these are m ineralisation an d im mobilisation that involve the tr an sfor mation of N between orga nic an d inorganic form s in the soil.

NI TRATES AND NITROGEN LOSS

17

Figure 2.3 Th e percen tage of total N availab le to a crop followin g su rface ap plica tion of man u re. Num bers in paren theses are percen tage DM con ten t of the r espective m anu re (data from Cham bers et al., 20 0 0 ).

Nitrogen mineralisation Soils natur ally contain N in two discrete pools: organic and inorganic. Depending upon the cropping history of the soil, m ineral com position and prevailing envir onm ental conditions the sum of these two pools would be between 20 00 and 600 0 kg N ha −1. Alm ost all of which would be in the organ ic for m and therefore unavailable for crop uptake (Powlson, 1993). Soil orga nic m atter is colonised by a variety of heterotrophic soil organisms that derive their energy for growth from the decom position of organic m olecules. During decom position essential nutrient elements, including N, are also converted fr om orga nic to inorganic form s. This is ter med mineralisation and occurs whenever soils are moist and warm enough for m icrobial activity, with a ‘flush’ of intense activity usually occur ring in the spring and to a lesser extent in the autum n. The mineralisation of organic N involves the degra dation of proteins, a mino acids, nucleic acids and other n itrogenous compoun ds to NH 4 + . Once formed, NH 4 + joins the m in eral N pool (a lon g with NH 4 + derived directly from fer tiliser, manure a nd atmospher ic deposition) and is subject to a num ber of potential fates, including fur ther tr ansfor mation by im m obilisation, nitrification or adsorption/ fixation, as well as direct loss fr om the soil by crop uptake or volatilisation. Mineralisation is the key process controlling N availa bility for plants and leaching loss (Goulding, 20 00). Nitrogen immobilisation

Nitrogen imm obilisation is defined as the transformation of in organ ic N com pounds (NH 4 + , NO3 − and NO2−) into organic N form s (J ansson and Per sson, 1982). It occurs when there is a readily available source of carbon (C)-rich m ater ial, such as following the addition of crop residues like cereal straw which are low in N, but high in C (C/ N ratio > 8 0:1, Chapter 5). Both m ineralisation an d imm obilisation are greatly in fluen ced by the availability of C (Powlson , 2 0 0 0 ). As m icroorganisms attack and decompose the C-r ich material, they also absorb NH 4 + from the soil and

18

CHAPTER 2

rapidly convert it to microbial biomass. Subsequently, the microbial biom ass dies, enters the active phase of the soil orga nic m atter and becomes liable to decom position again. Som e of the m icrobial N m ay en ter the passive phase to form hum us, or it may be released as m ineral N. The con tinuous turnover of NH 4 + in this m anner for ms a sub-cycle within the overall N cycle of the agricultural soil and produces a net effect—n et m inera lisation or net im mobilisa tion—which influences the supply of NH 4 + for other N cycle processes. Nitrification When (as is usual in m ost soils) the microbial population is limited by available C, m ost of the m ineralised NH 4 + is oxidised rapidly to NO3 − by the process of nitrification. This is a two-stage pr ocess m ediated by two im portant groups of bacteria: Nitrosomonas that oxidise NH 4 + to NO2 − and Nitrobacter which oxidise NO2− to NO3 −. Since the oxidation of NO 2 − is more r apid than that of NH 4 + , there are on ly ever trace amoun ts of NO 2 − in the soil. Im portan tly, the m inera lisation of organic N in soil and in cr op residues is seen as one of the m ajor sources of NO 3 − in agr iculture (Powlson, 2000). Adsorption and fixation Am m onium can be adsorbed onto the surface of clay minerals a nd soil orga nic matter (the ‘cation exchange sites’), from where it is freely exchangeable with other cations in the soil solution. Am m onium is also approxim ately the sam e size a s the K+ potassium ion a nd readily enters the interlayer portions of clay minerals (e.g. verm iculite). The collapse of this interlayer spa ce, for example by drying, effectively fixes the NH 4 + , m akin g it on ly ver y slowly available to en ter the soil solution . This is thought of as bein g a partial r eversal of the weathering processes undertaken by alum ino-silicate minerals. Nitrogen uptake and losses There are a number of pathways by which N is lost from an agricultural soil. The m ost desirable route is via cr op uptake and subsequent removal by grazing or harvest since this pr oduces both an econom ic return for the farm er, and is n ot a direct cause of pollution . The proportion of applied N taken during a growing season should, under controlled conditions, be between 50 and 70 % (for phosphor us 40 kg N ha −1 in m any UK soils) (Lor d and Anthony, 20 0 0 ). Major leaching losses occur when two conditions are m et: •

Soil water m ovem ent is la rge i.e. the in flux of water (either rain fa ll or irrigation ) is greater than the evapotranspiration. This is further influenced by soil texture and structure that may affect the hydraulic conductivity and water storage capacity of the soil. Nitrate leaching losses are gen er ally greater from poor ly str uctured sandy soils than well structured clay soils. • Soil NO3 − levels are high due to the min er alisation of organic N (possibly from the application of animal manures), or the presence of excessive or unused fer tiliser . For example, a Septem ber application of animal m anur e at the maxim um recom mended rate of 250 kg (total) N ha −1 (MAFF, 1998a) on a sandy soil would supply just 14 kg N ha −1 to a winter cereal crop, but 9 kg N ha −1 m ay be lost by leaching (MAFF, 2001). The pr oportion of leached NO 3 − derived from organ ic or fertiliser N sour ces will depend upon environmental conditions, the rate of applied N and the crop m anagem ent system s employed. However, considera ble quantities of NO3 − leached fr om arable soils can or iginate from the r apid m iner alisation of or ganic N rather than dir ectly fr om applied fertiliser. Particular ly under h orticultural cr ops, where large N residues may be left in the soil (>30 0 kg N ha −1 m ay rem ain after some brassicas) (Rahn et al., 1996). Figur e 2.4 gives an indication of the variation in NO3 − levels in a soil un der arable cr opping conditions in the UK and illustrates the tim e leaching and leakage is m ost likely to happen (Davies, 2000 ). A comparison of this Figure with Figur e 2.3 highlights the im portan ce of tim e of year in regar d NO3 − leaching from manure or fertiliser applications.

Seasonal rainfall and evapotranspiration patterns inter act with soil NO 3 − levels to affect leachin g losses. Although these vary greatly fr om year to year, and between regions, som e general statem ents can be made ba sed upon research in the UK and Den mar k (Powlson, 1988; Addiscott, 1996; Sim melsgaard, 1999; Chambers et al., 20 0 0 ): • in sum mer evapotra nspira tion gener ally exceeds rain fall and leaching is usua lly m inim al. However, fertiliser N losses can occur if application coincides with intense, heavy rainfall; • a particularly dry sum m er can limit cr op N upta keand lead to the accum ulation of NO3 − in the soil which is then susceptible to a utumn/ winter leaching;

20

CHAPTER 2

Figure 2.4 Var iation in soil NO3 − levels u nder a cer eal crop over a per iod of one year (Davies, 20 0 0 ).

• pr ovided that the soil is approaching or has r eturn ed to field capacity, autumn rainfall will leach any NO3− rem aining from pre-ha rvest fertiliser application or der ived from late summ er / autum n m iner alisation. Applications of fertiliser N to the seedbed of autumn sown crops are also very susceptible to leaching; • in win ter ther e is a large excess of rain fall overevapotra nspira tion an d any NO 3− present in the soil profile is readily leached; • spring applied fertiliser N is susceptible to leach ing if application coincides with hea vy rainfa ll or NO3 − is not ra pidly removed fr om the available pool by cr op uptake. The leaching of soluble organic nitrogen forms has also been suggested as being a major pa thway of N loss fr om agricultural soils. It is thought that in soils this pool may be of similar size to the m iner al N pool, but it is less subject to fluctuation and change (Mur phy et al., 20 00). Nitrous oxide and nitric oxide emissions from nitrification and denitrification Another route by which significant amounts of N may leak from an agricultural soil is via gaseous emissions of nitric oxide, NO, nitrous oxide, N 2O, and molecular n itrogen, N 2 (Fowler et al., 1996). The predom inant source of these em issions is micr obial nitrification (see above) and denitrification. Denitrification is the dissim ilatory reduction of NO 3 − is the m ajor natural process by which oxidised N is returned to the atm osphere (Chapter 6 ) (Royal Society, 1983). It occur s under anaerobic soil conditions, when NO3− replaces O 2 a s the term inal electron acceptor in microbial respiration. Conditions favouring denitrification ar e the pr esence of: adeq uate NO 3 − levels, den itrifyin g organ ism s, high soil water con tents (low soil oxygen) and poor soil structure (both of which r esult in low air filled porosity a nd potential anaerobic conditions) (De Klein et al., 2001). The

NITRATES AND NITROGEN LOSS

21

m ain rate determ ining fa ctors are soil tem perature and the amount of readily available carbon substrate presen t (Sm ith an d Arah, 1990 ) (Chapter 6). Denitrification can be a very im portant N loss process in agricultural soils, particularly fr om heavy soils in wet conditions, but its m easurement is frequently complicated by high spatial and tempor al variability and difficulties of determ ining total denitr ification loss fr om the m easurem ent of N 2 O emission s (Folorunso and Rolston , 1984; Arah et al., 1991; De Klein et al., 20 0 1). As well a s representing a further econom ic loss to the farm er by reducing the availability of m ineral N for crop uptake, emission s of NO and N 2O are pollutants that may pose an en viron m en tal hazard (Chapter 6) and also cause dam age to natural and sem i-natur al ecosystems (Box 2.1).

BOX 2.1 CRITICAL NITROGEN LOAD

Critical load is the m aximum atmospher ic ‘pollutant’ load that sensitive ecological systems can toler ate without incur ring long-term har mful effects (RCEP, 1996). The concept has m ainly been applied to acid pollutants, indicating the capacity of soils in sensitive environm ental areas to buffer atm ospheric acid inputs. Critical loads for N in puts ar e less easy to determ in e (N com poun ds, su ch as NH 3 , ca n both a cidfy an d act as a nu trien t), bu t have been estim ated at 5– 45 kg h a −1 year −1 for a r an ge of ecosystem s from heath lan d to com m er cial forestry; th e poorer t he soil and the sparser t he vegeta tion, th e sm aller the cr itical load. An an n ual deposition of up to 40 kg N h a −1, as estim ated by Gould in g et al. (1998 ), su ggests th at some n atur al and sem i-n atur al ecosystem s ar e likely to be receivin g m uch m or e N th an t heir critical load (Asm an et al., 19 98). This will cause ch anges in th e flora and fau na of th e ecosystem s du e to the in creased N supply, an d incr ea sed soil acidification as th e NH 3 deposited is n itr ified and relea ses H + ion s in to th e soil solu tion . Localised N deposition and soil acidification h as, for exam ple, been n oted in field crops due t o NH 3 volatilisation from a nearb y in tensive p ou ltry un it (Speir s an d Frost, 198 7).

Ammonia volatilisation +

Although NH 4 is not gener ally at risk of leaching because of its r etention in the soil on negatively charged cation ion sites, gaseous losses of NH 3 do occur in agr icultura l systems, including emissions from soil (Figure 2.1). In Eur ope, the largest source of atm ospheric pollution by NH 3 is agriculture, although oceans and biom ass burning are also important and an estimated 60% of global em issions are from anthropogen ic sources (Fowler et al., 1996; Asm an et al., 1998). Gaseous losses of NH 3 are most significant from agricultural system s involving livestock (especially intensive production systems) due to the breakdown of urea in anim al urine and faeces (Whitehead et al., 1986). Indeed, 92% of all am m onia in Western Europe originates fr om agriculture of which about 30% is fr om livestock (Isherwood, 20 0 0 ). Am m onia loss is likely to be greatest where high concentrations of am m oniacal N occur. For exam ple, from urine or slur ry. Losses of N as NH 3 m ay be as great as 80% of the total when slurr y is surface applied to gr assland (depending on weather and sward conditions) (Whitehead, 20 0 0 ).

22

CHAPTER 2

In ar able soils, the greatest losses occur when a mm oniacal fertilisers or urea ar e applied under alkaline conditions. Am m onia losses r esulting from surface vola tilisation a re aggr avated by high soil tem peratures an d drying conditions, but can largely be preven ted by placing fertiliser s below the soil surface or workin g them in thoroughly with the top soil (Tisdale et al., 1993). 2.4 NITROGEN FERTILISER USE IN AGRICULTURE Sustained agr icultural pr oduction depends upon the continual fixation of atmospher ic N 2 to replenish the N lost from the soil in harvested crops, livestock production an d the n atural soil loss pr ocesses, such as leaching and denitrification, already described above. Since the late 1940s agriculture has under gone r apid moder nisation as politica l and econom ic support cr eated a favourable econom ic climate in which technical efficiency and technological advancem ent wer e encouraged and flourished. In the m odern agricultural system s now typical of areas such as nor thern Europe, reliance upon legum es has long been superseded by the use of industrial N2 fixation and the application of synthetic or mineral N fertiliser s. For example, in 1850, wheat yields in France where 1000 kg ha −1, by 1950 this h ad reach ed 160 0 kg h a −1, with a fertiliser input of 1.1 m illion tonnes. By 1994– 1996 yields had reached 6772 kg ha −1 with an input of total fertiliser of 4.8 m illion ton n es (of which 2.4 million ton nes wa s N) (Isher wood, 2000 ). A similar pattern has occurred in the UK (Figure 2.5), but im portantly since 198 2 N fertiliser application to winter wheat have rem ained relatively constant (18 0 kg N ha −1), wher eas as yields have increased significantly (from 7.5 t ha −1) (FMA, 20 00). This is thought to be lar gely due to techn ological im provements in other aspects of crop production in cluding advan ces in seed strategies. Accor ding to Wild (1993), three developm ents were im portant in establishing the place of fertilisers in m odern agriculture: • Long-term field exper iments (e.g. those a t the Roth amsted Experim ental Station) showed that crop yields could be ma intained with continuous cropping when the required plant nutrients wer e applied as fertiliser. In practice, good m anagem ent also requir es at least some crop rotation (e.g. inclusion of br eak crops for pest and disease control), as well as the addition of organic ma tter in som e for m. • The industrial synthesis of N fertilisers in the 19 20s, based upon the ‘Haber pr ocess’, m ade large-scale fer tiliser production possible. Later developm ents in the oil industry provided cheap energy for the process. • The in troduction of higher -yieldin g, m ain ly short-s temm ed, varieties of cerea l crops which has m ade it profitable to add greater a mounts of fertiliser. Im proved cultivations, irrigation and the use of pesticides ha ve also justified the use of m ore fertiliser. Global fertiliser N use has dram atically increased since the mid-1940 s and by the late 1980s, average fertiliser N rates in the m ost intensively farm ed coun tries of nor thern Eur ope an d eastern Asia were in the region of 120 – 550 kg N ha −1 (FAO, 198 8 ). In contrast, the rates of total fertiliser applications in the countries of sub-Saharan African were in many cases less than 5% of those used

NITRATES AND NITROGEN LOSS

23

Figure 2.5 Nitrogen fertiliser use in t he UK (‘0 0 0 s tonn es) (FMA, 1998 ).

in intensive agricultural system s. Between 1993/ 4 and 1997/ 8 world fertiliser nutrient usage increased by 13% (Isherwood, 20 0 0). An estimated 9.6 m illion tonnes of N fertiliser are applied to 140 m illion ha of agricultural land across Europe, of which half is used on wheat, ba rley, oats, m aize and rye and a quarter on grassland (Aldinger, 20 0 1). However, fer tiliser N usa ge reached a peak in the mid to late 1980 s an d is expected to continue to decline to 20 0 6 (by approxim ately 7% of the cur rent total), from when it is thought that dema nd will rem ain stable (Aldinger, 20 0 1). Wor ldwide fer tiliser N use has had an enor mous impact upon agricultural productivity. It is difficult to separate out precisely the contribution of fertiliser N to increased output fr om that of the other technological inputs iden tified above by Wild (19 93). Estim ates in the UK (H ood, 198 2 ) suggested that fertiliser N had been responsible for 30– 50% of crop yield incr eases, with the rem ain ing 50 – 70 % due to im proved varieties, increased a grochem ical inputs an d better husban dr y techniques. Increases in fertiliser N use have not occur red uniform ly acr oss a ll regions. In the UK, increased fertiliser use has tended to be concentrated in those areas of the country most suited to intensive agricultural pr oduction. These areas have seen significant changes in agricultural land use, notably a decline in traditional ley/ ar able r otational system s and an increase in continuous arable cropping. In m uch of eastern England agricultural land use is now dominated by intensive cer eal pr oduction (Edwa rds and Withers, 1998). Sim ilar changes have been observed in m any par ts of the world as the availability of N fertiliser has helped to facilita te a profound change in cropping patterns. This has lead to a shift from m ixed and multiple-cropping system s with r elatively closed and self-sustaining N cycles to intensively m anaged m onocultures with large N inputs in the form of synthetic fertilisers (Rosswall and Paustian, 198 4). H owever, in som e low-incom e developing countries this shift has also been welcom ed, in that the cycle low-input low-output technologies that are perceived to perpetuate hum an drudgery with the ever present r isk of hunger is broken (Isherwood, 20 0 0 ). With these increases in fer tiliser usage and the accompanying changes in global cr opping pa tterns, incr easing a ttention ha s focussed on N leakage from the agroecosystem . Som e concer ns can be linked directly to fertiliser use, others r elate to the overall intensification of m odern

24

CHAPTER 2

agriculture r esulting in part from the increased use of fer tiliser N and the m iscalculation of requir em ents for crop growth. 2.5 THE CAUSES OF NITRATE POLLUTION The application of more N to a soil tha n can be assim ila ted by the soil or ta ken up by a crop creates a surplus of N. In the UK, in the early 198 0 s this surplus was in the or der of 70 kg ha −1, but is now down to 25 kg ha −1 (FMA, 1998). This sur plus is particular ly evident on dairy farm s where inputs are often in the region of 400 kg N ha −1 but offtakes may only be in the order of 60 –80 kg N ha −1 (Peel et al., 1997). H owever, it is im portant to stress that even under well-m anaged arable lan d the N surplus may still be in the order of 20kg ha −1 due to the mineralisation of organic N (Goulding, 20 0 0 ). The tran sfer and fate of this surplus has created many en vir onm en tal concerns in Europe. The m an ifestation s of the effect of the N sur plus produced through agricultural production are con sider able, especially as this surplus often behaves so con servatively in the soil system . The com plexity of the behaviour of N in the agroecosystem has been illustr ated in Figure 2.1 and with so many potential diffuse pathways for its loss from the agr oecosystem, it is little surprise that the efficiency of applied N used by farm ers never approaches 100%. Indeed it may even be as low as 10% in som e gr assland system s (Davies, 20 0 0 ; J arvis, 20 0 0 ). Over 8 0% of the total N in river water s is found in the for m of NO 3−, and in the last 30– 40 years, NO3 − levels in ma ny European groun d, surface a nd coastal water s have been gradually r isin g (House of Lords, 1990; DoE, 1986; Pau Vall and Vidal, 1999). An analysis of 12 UK rivers for which data was available over a 20 -year period showed incr eases in concen tration of between 50 an d 40 0% (Wilkinson and Greene, 1982). Figure 2.6 shows the mean concentrations of NO 3− in the Anglian region of eastern England from 1980 to 1998 and all are ver y close to the EU limit for NO 3 − in water s of 50 mg l−1 (DETR, 20 0 1). The concentr ation of nitrate in rivers and reser voirs on the Channel island of J ersey h ave been found to consistently exceed 50 m g NO3 −l−1, due m ainly to the large proportion of agricultural land that is devoted to the production of early potatoes. Losses from this crop can exceed 100 kg N ha −1 year −1 (Lott et al., 1999). Rivers with the highest NO3− concentr ations are found in the Midlands and south-east England, with the lowest levels in the mounta inous region s of Wales, n orther n England and Scotlan d. The tempor al a nd regiona l tren ds in river wa ter q uality are also reflected in lakes an d reservoirs. For example, there has been a n oticeable increase in reservoir NO3-N levels in south-east Englan d. Between 1992 and 1996 over 65% of European rivers had average a nnual NO 3 − concentrations greater than 1 m g l−1, and of those 15% were gr eater tha n 7.5 m g l−1 (Pau Vail and Vidai, 1999). Considerable qua ntities of freshwater are stored in groundwater aquifers, the most im portant of which ar e the chalk and Triassic sandstones. Long-term da ta on the NO3− N concentration of groundwater are less com mon than for surface water s. Nevertheless there is a m arked upwar d trend in m any catchm ent ar eas (Figure 2.7), par ticularly in the dr y easter n areas of England (Wilkinson and Greene, 1982; Da vies, 200 0; DETR, 20 01). This is probably due to the fact that the NO3− concentrations in water dr aining from agricultural land are dependent upon both the am ount leached but also the dilution rate. For exam ple, catchments in the wetter west of the UK

NITRATES AND NI TROGEN LOSS

25

have con siderable leachin g (50 kg N ha −1 year −1) but a lso greater r ainfall, so reducing a verage NO 3 concentrations (Davies, 200 0).

−

Figure 2.6 The m ean concen trat ion of NO 3 − (m g l−1) in river wat er s from th e Anglian region of Eastern England between 19 80 and 1998 (DETR, 20 0 1).

There is little doubt that the intensification of agriculture has been r esponsible for the increasin g levels of NO 3− in UK ground and surface waters (e.g. Royal Society, 1983; DoE, 1986; House of Lor ds, 1990 ; MAFF, 1993; Isherwood, 2000 ). It ca nnot be assumed, however, that this n itrate pollution is lin ked directly or solely to the incr eased use of N fertilisers; a correlation between the increase in NO 3 − pollution and the increase in fertiliser N use does not im ply causality. On the contra ry, NO3 − loss from agricultural lan d is a com plicated process tha t involves m any factors.

Figure 2.7 Th e cha nge in mean NO3 − concen tra tions (m gl−1) from 198 0 to1998 in t wo gr ou nd water sources in Eastern En glan d (DETR, 20 0 1).

For exam ple, intensive livestock production systems, produce large q uantities of anim al m anure which can pr esent disposal pr oblem s with excessive or un tim ely la nd applica tion leading to both diffuse and/ or point source NO3 − pollution (Chapter 5). Furthermore, the sur plus N in these

26

CHAPTER 2

Figure 2.8 Measu red N inp uts, outp uts an d surp luses (kg N ha −1) for two system s of dairy farm m anagem ent in t he Min im al Im pa ct Dairy Syst em s (MIDAS1) experimen t (Peel et al., 19 97).

systems can be con siderable even when a low in put m in im al-loss strategy is taken (Figure 2.8) (Peel et al., 1997). The balance of the two m ajor N inputs to agricultural soils, mineral fertilisers and livestock m an ures varies greatly across Europe, with the for mer bein g of greater impor ta nce in Den m ark, Germ any, Gr eece, France, Luxem bour g, Finland and Sweden and the latter in Belgium and The Netherlands (Pau Va ll and Vidal, 1999). It is widely acknowledged that a m ajor factor contr ibuting to the incr eased pollution of the aquatic envir onm ent by NO 3 − has been the specia lisation and intensification of agricultural en terprises (Chapter 3) (Edwards and Withers, 1998). Notably, the decline in tr aditional m ixed farming system s (i.e. crops and livestock on the sam e far m) and the increase in specia list arable and livestock far ms. These increase with the intensity of agr icultura l production and arise where N inputs (e.g. livestock feed) exceed N outputs (e.g. sales of crops and livestock). Nitrogen surpluses m ay be evident at a farm , r egional or national level, although their occurrence a nd extent differs consider ably between farm ing types a nd differen t coun tries. Specialist livestock farm s (notably dair y, pigs and poultry) ar e particula rly prone to accum ulating excessive amounts of N in the form of m anure because they use relatively high levels of N fer tiliser and/ or im port large amounts of concentrate feed, and are comm only found in geographically concentrated areas (Chapter 5) (Lor d et al., 1999). A specific trend observed in m any areas of the UK has been the transition (stimulated lar gely by UK and subsequently EU agricultural policy) from pastor al and balanced rotational cropping to intensive ar able production. The initial ploughing of gr assland is a significant cause of NO 3 − leaching and elevated nitrate levels in groundwater ha ve been specifica lly linked to the increased cultivation of perm anent grassla nd to pr oduce arable cr opland during the 1940 – 1950s. Guidance exists to a id the farmer in taking into account the potential relea se of N following the ploughing of grass, which is dependent upon th e am ount of fertiliser N th e gra ss received each year, the intensity of gr ass utilisation and tim e since ploughing (MAFF, 1994a). Importantly, it has also been established that even with this guidance it is possible to gr eatly over estim ate N fertiliser requir em ents for cer eal crops following grass leys. It is suggested that a reduction in N surplus

NI TRATES AND NITROGEN LOSS

27

would more likely be achieved if crop yield was removed as a determinant and an assessm ent of m ineralisa ble soil N was m ade during the growing season (Withers and Sylvester -Bradley, 1999). Furtherm ore, in ten sive ar able croppin g system s are in tr in sically leaky (Powlson , 1988; MAFF, 1993) and prone to nitrate leaching since: • they include periods of incom plete cr op cover which are vulnerable to leaching events; • regular cultivations stimulate N m ineralisation and increase the NO 3 − availa blity; • repeated applications of fer tiliser N incr ease th elevel of potentially mineralisable N in th e soil, thereby increasing the am ount of or ganically-derived NO 3 − that can be available for leaching; • there is the risk of direct leaching losses from excessive (sur plus due to poor N a ccounting) or poorly tim ed fer tiliser N applications and or ganic m anures (Richards et al., 1999; Cham ber s et al., 20 0 0 ; Goulding, 200 0; MAFF, 20 0 0 ) (Cha pter 5). Nitrate pollution of ground, surface a nd mar ine water s is a major environm ental issue in many European countries, with potential im plications for human health, which has attracted considerable public and political attention. In contr ast to pesticides, however, the control of NO3 − pollution is not simply a question of r ationalising and reducing an agrochem ical input as outlined above. While the steady increase in the use of N fertilisers is undoubtedly significant in causing NO 3 − pollution, the overa ll relationship between the rate of N applied a nd the incidence of NO 3 − leaching is not always clear and direct. Nitrate lea ching is caused by a number of factor s and arises from a variety of sources, and this potentia lly complicates the control of NO 3− pollu tion. 2.6 PROBLEMS CAUSED BY NITRATE POLLUTION

There are two key con cern s regardin g the NO 3 − pollution of water resources: the quality of drinkin g water a nd public health, and the eutrophication of surface waters. There is also som e concern about increased levels of NO3 − in foods and, although this is not strictly an agricultural pollution issue, it is given som e consideration below. Nitrate and drinking water quality Much of Europe’s drin king water is sour ced from rivers an d groundwaters. However, exten sive leachin g of NO3− from soils into these sources r esulted in waters in ma ny areas that approach or exceed the EU m axim um m anda tory limit concentration of 50 mg NO 3− l−1. An estima ted 1 million people in the UK regularly drink water with greater than 50 m g NO3−l−1 concentration (Pa ckham , 1996). Drinking water normally provides approxim ately up to 30% of the daily intake of NO 3 − ingested by humans, with the rest com ing from fruit, vegetables and meat products (Isherwood, 20 00). Nevertheless, the ingestion of large amoun ts of NO 3 − in drinking water may be harm ful to huma ns since, a lth ough nitr ate is relatively non-toxic and rapidly excr eted from the b ody it can be reduced to potentially toxic nitrite in th e m outh and gut (Magee, 198 2). The m ain alleged health h azards of NO3 − in drinking water are methaem oglobinaem ia in young babies (0 .125 mg P l−1) in European rivers are in Northern Irelan d, southern England and acr oss centr al Europe through Rom ania to the Ukraine (Steén, 1997). Agricultural non-point source losses of P account for less than 50% of the total (of

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which 16% is thought to or iginate from fertiliser s and 34% from livestock), but this is the hardest and most costly to control (Environm ent Agency, 20 0 0 a). It is estimated alm ost 90 % of the P enterin g rivers in East An glia is from sewage wor ks effluen t. Areas which have been particularly susceptible P enrichm ent from this sour ce also include Loch Leven in Scotland, Lough Neagh in Northern Ireland, the Shropshire and Cheshire m er es and a n um ber of water supply reser voirs in southern En glan d (Foy a nd Bailey-Watts, 1998). In 1999, 23% of all substantiated water pollution incidents (those in which a prosecution followed) in Englan d an d Wales were caused by sewage, on ly those caused by fuels an d oil wer e greater at 26% (Envir onm ent Agency, 2000 a). H owever , the inputs of P from point sources are easier to iden tify an d so have seen m ore attem pts at control than diffuse or nonpoint sources as industrial and sewage discharges ar e brought into line under the Urban Waste Water Trea tm ent Directive (Heathwa ite, 1997; Foy and Bailey-Watts, 1998). Yet there is lim ited eviden ce to suggest that the reductions of P input from these point sour ces has had any dr amatic effect on P content in aquatic system s, suggesting diffuse sources of P (or even P release from sedim ents) m ay be pla ying a critical role (Steén, 1997). 3.5 ENVIRONMENTAL PROBLEMS CAUSED BY P POLLUTION The over riding concern associated with P loss from agricultural land is the pollution of sur face water s causing eutr ophica tion . However, a further issue is the loss of indigenous vegetation types in naturally nutr ient poor grassland system s which ha ve received increased P inputs. The human health effects of excessive P ar e, un like N, pr imarily lim ited to those effects resultin g from im pacts on surface water q uality. P in ground and surface waters Phosphorus concentration in gr ound waters tends to be relatively low and subsequently of lim ited en viron m en tal im portan ce (Sharpley et al., 198 7). Nevertheless, the total concentration of P in surface water s (often considered critical in ter ms of the potential to accelerate biological pr oductivity a nd potentially eutrophication) is only 0 .0 1 m g l−1 (cf. N 1–0 .5 m g l−1) (Daniel et al., 1998). This is 10 tim es lower than the concentration in soils comm only thought to be required for crop growth. This difference is of fundamental impor ta nce in regard to issues of water quality and the managem ent of P usage in agricultural system s. The loss of P from these system s is of lim ited agron omic value, but crucial in regard to en viron men tal q uality. The effects of these losses depend upon the characteristics of the receiving water. For exam ple in rivers, wher e ecology is la rgely determ ined by flow regime, residence times m ay be too short for phytoplankton to develop. Favoura ble conditions m ay be reached when the speed of the water is checked, such as in reservoirs, dam s or slower movin g r eaches (Gibson , 1997). The biological responses in str eam s and rivers to increased P loading are site and season specific (Edwards et al., 20 0 0 ). In the UK, rivers flowing through ar eas dominated by arable farm ing tend to have greater P concen trations than sim ilar lowla nd areas in which pastor al activity is the m ain use (Environm ent Agency, 20 00a).

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In creases in P in sur face waters may cause rapid r ises in the growth of algae an d other water plants. Indeed, a linea r relationship is thought to exist between biom ass production and concentrations of total P in waters of less than 0 .1 mg l−1, but above this other fa ctor s such as light availability become more important (Gibson, 1997). A schem atic of the relationship between prima ry production, the input of P and the trophic status of a freshwater system is given in Figure 3.7. Gener ally, there are four consecutive stages of eutrophication ; oligotr ophic, m esotrophic, eutrophic and hypereutrophic. The progr ession through these stages r esults in changes in the biodiversity, ecology and oxygen status of the water body (Peir szynski et al., 20 0 0 ).

Figure 3.7 Th e relation ship between P in put an d biod iversity in freshwat er syst em s, th e values below each tr op hic status are exam ples of possible wat er qua lity chara cteristics (adapt ed from Cor rell, 1998 a nd Mason , 1996).

On reaching surface waters the for m of P can change, so while dissolved P m ay be the for m in which it reaches a lake, it m ay well be sorbed on to sedim en ts an d organ ic matter in the water colum n (Figure 3.8). Soluble P ma y represent an im media tely bioavailable source on reaching a receiving water body. Particulate P is a long-term source to the aquatic system and the rate of P release is determ ined by the physico-chemical char acter istics of the specific sedim ent and water colum n (Sharpley et al., 1987). In nutrient poor aquatic system s, for exa mple a freshwater lake, the P will rem ain predom inantly in the bottom sediments. However, if this system r eceives a large input of P it m ay become eutrophic. The resulting increase in primar y production can prom ote anoxic conditions in the water close to the bottom which causes the increased relea se of P from the sediments back into the water column (Corr ell, 1998). A com mon pattern of P cycling in lakes is that in winter most P is in the dissolved form , but as spring begins the increased light favours

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photosyn thetic activity and phytoplan kton population s in crease and dissolved P con centrations fall (Gibson, 1997).

Figure 3.8 The p ot en tia l cycling of P between various form s in aquatic systems (ad apted from Cor rell, 1998 ).

The P tran sform ation s occur rin g in feeder chann els may impact greatly upon the biological pr oductivity of the receivin g lake, reservoir or strea m. This is very impor tant, as n et P in put to a lake or strea m from agr icultural non-point sources may incr ease the cha nces of water quality pr oblem s such as eutrophication, but the transform ations of P between the var ious for ms a ffecting bioavailabity will determ ine the extent to which nutrient enrichment will occur. Eutrophication The eutr ophication of surface waters is a globa l issue and is a result, in par t, of elevated P concentrations. It can be defined as ‘an increase in the nutr ient status of natural waters that causes an a ccelerated growth of algae or water plan ts’ (Pierzynski et al., 2000). It is essentially a natur al process, but over stim ulation through an thr opogenic inputs (cultural eutrophication ) results in an increase in algal growth a nd plant production (Figure 3.7). It is comm on for lakes to have a pea k in the population of d iatoms (Bacillariophyta) in spring with a late spring flush of green algae (Chlorophyta). H owever, in eutrophic system s on e m ay expect a lar ge sum mer peak in blue-green algae (Cyanophyta) (Mason, 1996). Approxim ately 6% of river lengths in the UK ar e designated as Eutrophic Sensitive Ar eas and in stan ces of blue-gr een algal blooms are reported fr om over 20 0 sites a year (Environ m en t Agency, 200 0a, b). The long-term change in the trophic status of freshwater system s in the UK is difficult to deter mine, but between 199 4 and 19 98 a doubling in the num ber of designa ted sites sensitive to eutr ophication has occur red. Accelerated eutrophica tion of aquatic system s is associated with surface rather th an subsur face inputs of P, a consideration of im portance in the following discussion of m anagem ent objectives to

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reduce P leakage. The detrimental effects associated with eutrophication of surface waters in the UK were outlined in Chapter 2 and include restr ictions in water use for fisheries, r ecr eation, drinkin g for hum an s and livestock an d in dustry. Further discu ssion of these issues are given in other texts (Harper, 1992; Sakam oto, 1996; Moss, 1998). 3.6 PRACTICAL SOLUTIONS The optim isa tion a nd effective use of soil P surpluses and fer tiliser inputs are r equired for the sustaina ble use of P in productive agricultural system s. An improvem en t in un dersta nding in these areas will enable the developm ent of measures to reduce the loss of P from agricultural land and the subsequent transfer to surface wa ters (Sharpley et al., 20 0 0 ). As the prim ary m echanism of P tr ansport is via run-off, both dissolved and par ticulate P form s, the reduction of non-point sources of P requir es managem ent strategies that addr ess b oth the sources a nd transfer mechanism s. Source ma nagem ent, such as better farm -scale nutrient budgetin g and balan cing P in puts for livestock and cr op production m ay offer con siderable financial sa vings along with environm ental benefits. Transport m anagement focuses on the reduction of the m ain P transfer pathways through the use of soil conservation, land-use m anagem ent and buffer zones (MAFF, 1998). It is clear fr om the preceeding chapter on N, that m any of the m anagem ent practices set out as pa rt of the Code of Good Agr icultural Practice for the Protection of Water (MAFF, 1998) will also apply to the reduction of P leakage. However, the Code also recomm ends that for soils conta ining over 25 m g kg−1 Olsen P the rates of m anur e application m atch P crop offtake and that no P is applied to soils conta ining >45 mg kg−1 Olsen P (soil P Index 3—Table 3.5). Furthermore, n utrien t budgeting targets are given along with guidance for the environm ental suitability of changes in m anagem ent on soils with a high soil P index (Withers et al., 2000 ). Concern in r egar d to eutr ophication a nd the im pa ct of non-point source P losses fr om agricultural systems at the national level is acknowledged by the publication of a national strategy —Aquatic Eutr ophication in England and Wales: a Management Strategy (Environm ent Agency, 20 0 0 a). This docum en t outlin es an holistic approach to aq uatic eutr ophica tion , which requir es action under a range of statutory and international com mitm ents entered into by the UK (EC dir ectives, UK Biodiver sity Plan , the OSPAR Conven tion ). The En vironm ent Agen cy strategy is to focus prima rily on N and P and their impacts upon ‘con trolled waters’ (lakes, streams, can als, reservoirs, estuaries and coastal waters), and the overall objectives are to: •

pr ovide a framework to manage cultural eutrophication , to protect, rehabilitate an d restore water s adversely affected; • deliver on the Environment Agency’s ‘Environm entalVision’, particularly its com mitm ents to eutrophication control; • ensure contribution to the UK Biodiversity Plan, especially implem entation of Ha bitat Action Plans for standing waters; • pr om ote a r eduction in nutr ient content of waters nationally, coordinate catchment based m an agem en t of policies and procedures for the assessment a nd con trol of local eutrophication .

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The document outlines a ser ies of actions that will be per form ed to achieve these objectives which include: • the setting up of eutrophication control action pla ns through the Environment Agency’s catchment-based local Environment Agency Pla n; • pr om ote the use of recycled phosphorus from soa psan d deter gen ts; • refin e and prom ote the uptake of the curren t codesof pr actice; • with other organisations (e.g. LEAF) pr ovide dem onstration farms of good nutrient m anagem ent and com m ercial success; • review cur rent eutrophication m onitoring procedures via the National Collabor ative For um on Environmenta l Monitoring; • further work on assessing the cost-effectiveness of curr ent eutrophication control; • pr ovide interim targets (annual m eans) for P in fre shwaters (eutr ophic in standing water—0 . 0 85 mg Pl−1, r un n ing water—0.200 m g soluble Pl−1). A ran ge of practical stra tegies and option s for reducin g P leaka ge from agricultural systems curren tly exist an d several examples are discussed below a nd shown in Figure 3.9 (Withers et al., 2000 ). At field level, costs are generally high, but the m ethods effective. However on a catchm ent scale losses can be gr eatly reduced at relatively low costs. Incr easin gly approaches focus on an integr ated catchm ent m anagem ent plan , wher eby site an d catchmen t specific factors are taken into account and r isk areas targeted. Source management—optimisation of fertiliser P use The r eduction of P surpluses in soils is a long-term issue and rectification will take m any years. Nevertheless, targeted use of fer tilisers and nutrient m anagement planning (and budgeting) may be effective in the short-term in reducing P losses through run-off (Sharpley et al., 1993). The key aspects of P fertiliser m anagem ent include: • accurate estimates of cr op and forage plant P requirem ents; • use of soil, m anure and plant tissue testing; • m on itoring of soil P levels; • timely a nd efficien t fertiliser practices (Danielet al., 199 7; Pierzynski et al., 2000). Many soils in the UK have a P sur plus and the greatest risks of fur ther accum ulation a re where there is a limited lan d a rea for m an ure disposal. In pa rticular, this occurs where m an ure application s are lim ited solely by N requir em ent—th e ratio N: P is often lower in man ure (e.g. 3 ) than is r em oved in the crop (e.g. 8 ) and so N tends to taken up, where as P m ay build up in the soil (Shar pley et al., 19 93; Sm ith et al., 199 8). It is therefore n ecessar y that farm er s adequately account for the P contr ibution in manures (even basing manure applications on P requirement not N) with the fram ework of an appropriate fertiliser plan (MAFF, 20 0 0 b). Following the determ ination of a P requirement then the rate, tim ing, method a nd a mount of fertiliser application need to be deter mined. The form of P fertiliser is thought to have a lim ited im pact on P loss via run-off if the fer tiliser is incorporated to the soil (Pierzynski et al., 20 00).

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Inor ganic fer tiliser applications m ay be placed and banded with crops to ensure m ore efficient usage in ter ms of cr op uptake and crop recover y of applied P and less chance of loss through run-off (Tisdale et al., 1993; Sharpley and Halvorsen, 1994). Estim ates of P loss in runoff suggest that fields receiving broadca st P ma y be two order s of m agnitude grea ter than from fields in which the fertiliser has been incorpor ated into the soil. Sim ilar r esults have also been observed for broadcast and incorporated m anur e applications (Sharpley et al., 1993). Greater challenges exist in regard to m anur e usage a nd P loss. Best ma nagem ent pra ctice for the recycling of anim al m anur es and organic wastes to land ar e outlined in Chapter 5. H owever , specific measures to reduce P loss include composting, pelletising, tran sport from P surplus to deficient a reas and the use of additions such a s alum which can reduce P solubility (Sharpley et al., 20 00). Im proved utilisation of the residual P conten t in soils and the m or e effective use of fertiliser applications could be possible through cr op selection/ r otation tillage and plant stubble m anagem en t. Animal feeds Livestock feed is often seen as one of the m ajor contributary factor s to excessive P input into soils. Like the reduction and tar geted use of P in fertilisers, the reduction of P inputs into livestock feeds is unlikely to have an im pact in the short-ter m, as soil con cen tration s will change on ly slowly (Withers et al., 2 000 ). Therefore, this may be considered as a long-term preventative m easure in com parison to the soil conserva tion and transfer managem ent str ategies below, which ar e likely to have im mediate measura ble impacts on leakage rates. The m an ipula tion of dietary P in take of livestock in cludes methods for in creasin g P adsorption by th e anim al, and refining animal feed rations. This issue is clouded by the variability in P availability in forages, feeds and supplements and the clear need to balance production and animal per form ance. Purchased feeds account for about 6 0% of the P inputs into intensive dairy operations and ar e mostly consum ed as phytate. It is thought that limiting the am ount of P fed to livestock is only going to be effective in r educing loss in intensive farm ing system s. Reducing stocking densities, the use of inorganic P fertiliser applications a nd using less purcha sed feeds would probably have a larger effect on reducing the P surplus and subsequent losses (Valk et al., 20 00). Nevertheless, the amount of P in livestock feeds need to be balanced with dieta ry requirem ents. Recent studies in the Netherlands suggested that feeds for dairy cows conta ined up to 20 % more P than was required, due to the P con ten t of grass silage. An overall reduction of dietary P to dairy cattle of 10 % could be m ade (20% in the UK) without im pacting upon pr oduction (Valk et al., 20 0 0 ). A further strategy that is cur ren tly being adopted to increase dietary utilisation of P by pigs and poultry is the supplem entation of grain feed with phytase enzym es. This enzyme is absent in the guts of th ese anim als and enables the digestion of a greater proportion of th e P from th e grain (the phytate P). Furthermore, gr ain varieties, low in phytic acid, are now being grown as feeds to reduce the need for the enzyme addition. Both these measur es may reduce P in excreta by over 20 % (Pierszynski et al., 20 00). In the Netherlands there is combination of approaches being adopted to reduce excessive use of P. The policy is to br ing about a stepwise r eduction of the P supply towar ds an equilibrium level

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Figure 3.9 The nut rien t an d t ran sport m an agem en t op tion s for th e contr ol of P lossfr om non -p oint sou rces (adap ted from With er s et al., 2 0 0 0 ).

based on ca lculated rem ovals and known inputs (Van der Molen et al., 1997). The eq uilibrium level is defin ed as the sum of the m an ure an d fertiliser supply that m eets the n eed of the cr ops an d com pensates for inevitable losses. This approach defines losses under ‘good agricultur al practice’ which are often higher than those losses consider ed acceptable in order to m eet envir onm ental standards. It is estimated that a relatively sm all excess of P inputs, in the order of 5 kg P ha −1 year −1, will increase the a rea of land in th e Neth erlands that is defined as strongly P-saturated by 100 % by the year 2050. In som e parts of the countr y where P-induced eutrophication is a severe risk, m anure-P retur ns to land are restr icted based on a manure quota (Hotsm a, 1997), with excess being exported to other r egions with a lower P input. This is just part of a str ategy being em ployed in the Netherlands wher e the intensity of livestock production has lead to serious envir onm ental pr oblem s. Other appr oaches include the restor ation of buffer zones and wetland m argins, lowering groundwater levels and applying ferric and alum inium com pounds to soils (Van der Molen et al., 1997). P transfer management by soil conservation Soil con ser vation practices promote environ mental and econ omic susta in ability, irrespective of the potential influences upon P losses. In a recent study of 13 erosion susceptible ca tchm ents in the UK (MAFF, 1997) er osion occurred in 40 %, of which the m ajority where cropped to winter

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cereals. The m ain factors a ssociated with this er osion were poor crop cover (7°. The assessm ent of er osion risk, and methods for reducing the

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vulnerability of soils to erosion, such as maintaining a high soil infiltration r ate and reducing the catchment area to m inim ise the risk of sedim ent associated r un-off ar e discussed later in this chapter.

BOX 4.1 PROCESSES OF EROSION (MORGAN, 1995)

The two m ajor transport a gents ar e wa ter and wind. Er osion is a two-phase process: detachment of soil pa rticles, followed by transport. When sufficient energy is no longer available to transport the particles, then deposition occurs. Water Detachment Rain sp lash, con tr olled by inten sity of ra infall, soil ch aracter istics (e.g. t ext ure an d organ ic ma tter con ten t) an d soil an tecedent con ditions. Coar se silt and fine san d-sized par ticles (0 .0 5– 0 .2 0 -m m diam et er ) are m ost easily d eta ched by rain. Overland flow, occur s wh en surface stor age a nd in filt ration capacity of t he soil is exceed ed . Flow velocity an d soil coh er en ce con trols par ticle d etachm ent rate. For exam ple, clay par ticles ar e m ore resilien t t o detach m en t b ecau se of the coh esiven ess of in dividu al cla y part icles. Transport On ce a par ticle has becom e en trained, it will be tr anspor ted by water un til th e energy in th e flow redu ces below a thr esh old th at is gr ain size depend en t. Coar ser , san d-sized p articles a re deposited fir st (closest to t he sour ce of t he sedim en t) wh ile silt and clay-sized m aterial (< 0 .0 6-m m diam eter) ten ds to r em ain in suspension for lon ger. Wind Detachment Th e capacity of wind to detach p articles depend s up on th e critica l shear velocity of th e win d, which is con trolled by win d sp eed a nd surface r ou ghn ess. Par ticles between 0 .10 and 0 .20 -mm diam eter ar e most su scept ible to win d erosion . Transport On ce detach ed in m otion , sedim en t tran sport takes place in suspen sion (for p articles < 0 .2-m m diam et er ) or saltation (boun cin g). This saltation process will detach oth er p articles. Deposition occur s wh en win d speed is reduced , for exam ple by a bar rier su ch as a h edge.

Wind erosion Soil erosion by wind is m ost comm on in ar id areas, and on coastlines, where winds are strong and vegetation is sparse due to salt deposition. The severity of wind erosion depends on wind speed an d soil/ vegetation surface r oughn ess (Box 4.1). Once soil particles have becom e detached and are in m otion, they can be tr anspor ted considerable distances. A sequence of drought years in the 1930s in the South West USA r esulted in the transport of fine sand and silt-sized par ticles more than 10 0 0 km from the source of the erosion (Goudie, 20 0 0 ). The resultin g sedim en tation is regarded as on e of the most widespread form s of water pollution (Tivy, 19 90 ). As with water erosion , the disaggregation and tran sport of surface soil horizon s by win d removes particulate m aterial that includes organic, nutrient-rich com ponents associated with the finer fractions of the

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eroded material. These fractions tend to fall within the range 0 .05– 0.5-m m diam eter (Mor gan, 1995). In the UK, the more temper ate clim ate and surface topography tends to restrict the area of soils vulnerable to wind erosion. Mostly, it is the fine sandy and pea t soils in a rable cultivation that are vuln erable to loss. Most ‘win d-blows’ ten d to occur in the sprin g an d early summ er , and affect fields planted to high value crops such as sugar beet, onions and carrots (Evans, 1996a; MAFF, 1998). Estim ates of the spa tial extent and quantity of soil lost during periods of wind erosion are necessarily tentative, due to the difficulties in measur ing deposition ra tes a cross la rge ar eas down win d of a ‘win d-blow’. Evans and Cook (1986) obser ved m oderate and severe blows in the Cam bridge peat fens and Nottinghamshir e sa ndlands in 5 or 6 years in 10 , with soil loss rates of 5– 10 m 3 ha −1 being typical. Agricultural land management practices and accelerated erosion Many land managem ent practices have been found to increase er osion risk. Often these are integr al with the progressive intensification of agriculture tha t has occurred in r ecent decades. Globally, it is acknowledged that the progressive incr ease in area of cultivated land during the twen tieth century has exposed more land, often un suitable for arable cultivation, to the erosive forces of water and wind (Parkinson, 1995; Brown et al., 20 00). Agricultur al land m anagem ent practices tha t ar e known to increase the vulner ability of soils to erosion are given in Table 4 .2. In many cases, it is the com bination of a number of these practices with naturally erodible soils or steep slopes that leads to enhanced risk of serious erosion. Repeated cultivation changes the soil configuration, and has been shown to r educe the organic m atter content of soils such that structure stability can be com prom ised. Surveys conducted by the Nation al Soil Resources In stitute in Englan d a nd Wales ha ve revealed that soil organ ic matter levels in arable and ley-arable cropping have declined by an average 0.5% over the 15-yea r period 1980– 1995 (MAFF, 2000 ). For sandy soils that tend to have a naturally low organic m atter content, the continued loss of organic m atter can lead to capping of the soil surface on ra infall im pact. This reduces in filtration an d allows run off to occur m ore frequen tly. For a deta iled r eview of soil factor s influencing accelerated water erosion, see Evans (1996b). Table 4.2 Land m an agemen t p ractices which in crease vuln erability of soils to erosion .

Crop rotation exerts a significant influence on susceptibility to erosion. Evans (1996a) notes that the shift fr om spr ing to winter cereals in the UK since the 1970 s, encouraged by changing EU

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pr ice support m echanisms, ha s lead to a n increa se in incidence of within-field r iver bank and ditch erosion, accom panied by the siltation of more gently sloping r eaches of rivers. Cultivation and drilling of autum n sown cr ops such as winter wheat and oilseed r ape leaves the soils with a fine surface dur ing the autum n period. Late drilling and heavy autum n ra infall have lead to the increased incidence of water erosion in recent years in the UK (Davidson and Harrison, 1995). Sim ilar ly, spring sown crops such as potatoes and sugar beet have been implicated in serious erosion of vulnerable soils. In a sur vey of erosion and farm ing pra ctices in England and Wales, Skinner and Cham bers (1996) asked fa rmers to consider which land m anagement pra ctices would affect the sever ity of soil erosion . The factors identified in order of priority were arable cr oppin g (20 %), tractor wheelin gs (16 %) an d tram lin es (8%). Rem ova l of hedges, which leads to an in creased catchm en t area for developin g rills an d gullies has also been foun d to be im portan t (MAFF, 1998, 19 99b). In the United States, continuous m aize production in the absence of soil conservation m easures has m ade a significant contribution to soil loss from agr icultural land (NRCS, 1997). Ultimately, it m ay be tha t major changes in methods of cultivation and the intensity of land use are needed to r educe rates of erosion fr om agricultural land. Reganold et al. (1987) r eported on the lon g-ter m effects of con vention al and organ ic farm ing on soil pr oductivity, depth and erosion rates on soils in Washington State, USA. Silt loam soils on two adjacent farm s were cultivated with winter cereals for 37 yea rs. At the end of the per iod the topsoil of the organically farm ed soil was 16cm deeper than the conventionally far med soil. The differ ences were attr ibuted to significantly greater losses of soil by erosion on the conventionally farm ed soil. Measured losses on the conventional field in this area, where soil er osion rates were high, averaged 32 t ha −1, while those from the organically farmed soil were four times lower , typically 8 t ha −1. It was predicted that at the cur rent rate of er osion, a ll the topsoil on conventional far ms growing cer eals would be lost by the th ir d decade of the twenty-first century.

SOIL EROSI ON

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4.3 ENVIRONMENTAL IMPACTS OF SOIL EROSION Impacts of sediment on aquatic environments The transfer of soil particles eroded from farm la nd to a djacent watercourses can have significant im pacts on water quality. Some im pacts, pr im arily those of a physical nature, are tr ansient, as erosion is an episodic process, linked to infr eq uent, high m agnitude clima tic events. Other im pacts, associated with chemical pollutants, tend to ha ve a cum ulative in effect on aquatic organisms. Most authorities (for exam ple, NRA, 1992) cla ssify these im pacts as diffuse pollution, as m ajor erosion events tend to a ffect lar ge areas of land within a given catchment. Successful strategies to control losses and m itigate against these im pacts tend to be catchm ent based (Napier, 1990). Physical impacts Erosion generates par ticulate m aterial, predom inantly of an inorganic nature, that can have deleterious im pacts on both plant and animal life. Despite the fact that er osion events ar e shor tlived, the impacts can be very persistent. Clearly, the frequency of storm events controls the ability of an aquatic ecosystem to r e-adjust to less sediment rich conditions after run-off events. Physical im pacts on aquatic organism s tend to be difficult to assess, as periods of high suspended sedim ent concentration ar e accom panied by oxygen depletion and elevated concentrations of chemical pollutants, such as pesticides bound to silt and clay. Exam ple impacts of lowland r ivers in the UK are given in Table 4 .3. Table 4.3 Physical im pact s of in creased sedim ent load on aqu atic or gan ism s (Leeks, 1995; Evans, 1996a).

The im pacts of soil erosion can be very localised. For exam ple, at a confluence between a sma ll ditch or stream which is supplying the sedim ent, and a larger river inhabited by a wide r ange of

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CHAPTER 4

aquatic organisms. Sand-sized sediment (