Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Prepared for: Defra

Prepared by: David Harris, ADAS Boxworth, Battlegate Road, Boxworth, Cambridge, CB3 8NN Dr. Bob Crabtree, CJC Consulting, Oxford John King, ADAS Boxworth Paul Newell-Price, ADAS Gleadthorpe

Date: July 2006

Copyright

The proposed approach and methodology is protected by copyright and no part of this document may be copied or disclosed to any third party, either before or after the contract is awarded, without the written consent of ADAS. 0936648 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Glossary of Terms ALC Agricultural Land Classification AONB Area of Outstanding Natural Beauty BMP Best Management Practice generally defined by being within the Codes of Good Agricultural Practice for Air, Water and Soil, COGAP (Defra) Brickfield series An imperfectly drained soil with a fine loamy texture CAP Common Agricultural Policy Clifton series An imperfectly drained, medium to coarse-textured soil with a perched water table Cu Copper CV Contingent valuation DE Direct energy used in fuel for field operations Defra Department for Environment, Food and Rural Affairs DoE Department of the Environment (now part of Defra and distinct from the Environment Agency) Dunkeswick series A poorly drained soil with a fine loamy topsoil, and a clay subsoil beginning at between 40 and 80 cm depth ELS Entry Level Scheme of Environmental Stewardship Scheme ESA Environmentally Sensitive Area FIOs Faecal indicator organisms GAEC Good Agricultural and Environmental Condition GHGs Greenhouse gases Hallsworth Series A poorly drained, heavy-textured soil with shrink-swell properties HLS Higher Level Scheme of Environmental Stewardship Scheme IE Indirect energy used in manufacture of inputs, machinery and transport K Potash MAFF Ministry of Agriculture, Fisheries and Food Mineral soil All soils that have less than 6% organic matter in the surface soil horizons (MAFF, 2000) MWDs Machinery work days Ni Nickel N Nitrogen Organic soil A soil with high levels of organic matter – above 6% organic matter in the topsoil (MAFF, 2000). P Phosphorus PSYCHIC Phosphorus and sediment yield characteristics in catchments PV Present value Rhizobium Rhizobium bacteria cause the roots of white clover to develop nodules in which atmospheric nitrogen is fixed SAC Special Area of Conservation SAPs Special Protection Area SOC Soil organic carbon Soil series Soils are categorised into a number of soil series, of which Clifton and Hallsworth are two SPS Single payment scheme SSSI Site of Special Scientific Interest SUDS Sustainable urban drainage systems TP Total phosphorus TSC Total sequestered carbon UKCIP United Kingdom Climate Impacts Programme WFD Water Framework Directive WHO World Health Organisation WSP Water storage potential WTP Willingness to pay Zn Zinc

i Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Acknowledgements The project team would like to thank the steering group at Defra for their input and assistance, particularly since this is a new area of work.

We would also like to thank the following for their assistance with the report:

Professor Ken Willis, University of Newcastle on Tyne

Professor Brian Chambers, ADAS

Dr. Mark Shepherd, ADAS

ii Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Contents 1. Executive Summary...... 1

1.1 Remit ...... 1 1.2 The Project ...... 1 2. Introduction...... 10

3. Marginal Valuation of Soil Functions and Services...... 13

3.1 Valuing Soil Services ...... 13 3.2 Market-based Approaches...... 14 3.3 Valuation of Non-market (environmental) Outputs...... 17 3.4 Cost-based Approaches ...... 18 3.5 Marginal Valuation of Soil Services: Conclusions ...... 19 3.6 Total Marginal Valuation of Soil Services...... 19 4. Carbon Storage and Sequestration ...... 21

4.1 Technical Review...... 21 4.2 Valuation of Carbon Storage and Sequestration...... 25 4.3 Benefits from Carbon Sequestration in Soils...... 27 4.4 Conclusions ...... 29 5. Water Storage and Flow Mediation...... 31

5.1 Technical Review...... 31 5.2 Water Storage Potential (WSP) ...... 32 5.3 Conclusions ...... 35 6. Valuation of nutrient cycling (food and fibre) for crop production ...... 36

6.1 Technical Review...... 36 6.2 Supporting Crop Production...... 38 6.3 Valuation of Nutrient Cycling...... 46 6.4 Conclusions ...... 47 7. Supporting Construction – Accessibility and Soil Handleability in Relation to Particular Soil Types/Situations ...... 49

7.1 Technical Review...... 49 7.2 Valuation of Supporting Construction...... 51 7.3 Economic Analysis...... 51 7.4 Conclusion...... 51 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

8. Natural Attenuation of Pollution/Contamination and Contaminated Soil Remediation ...... 53

8.1 Technical Review...... 53 8.2 Natural Attenuation of Industrial Pollutants...... 53 8.3 Natural Attenuation of Diffuse Pollution ...... 54 8.4 Economic Analysis...... 63 8.5 Summary and Conclusions...... 66 9. Archaeological and Landscape Heritage Protection...... 69

9.1 Technical Review...... 69 9.2 Valuation of Archaeological Site/Artefact Protection...... 69 9.3 Factors that Affect Archaeological Site/Artefact Protection...... 71 9.4 Conclusions ...... 71 10. Support for Ecological Habitat and Biodiversity ...... 73

10.1 Technical Review...... 73 10.2 Valuation of Support for Ecological Habitat and Biodiversity...... 76 10.3 Conclusions ...... 78 11. Conclusions ...... 79

11.1 Methodology ...... 79 11.2 Evidence on the Marginal Valuation of Soil Functions ...... 79 11.3 Appropriate Methodologies for Measuring Benefits ...... 83 11.4 Information Gaps ...... 85 11.5 Policy Implications ...... 86 12. References...... 87

List of tables Table 1.1: Appropriate methods for valuing the benefits from soil-related policy measures...... 7 Table 2.1: Matrix of examples of Drivers of change (x) and resulting pressures on soil that are relevant to the economic valuation of the soil functions...... 11 Table 3.1: Implicit prices of soils in England and Wales (from Maddison, 2000)...... 16 Table 4.1: The final carbon sequestration and saving rates applied to each unit area of land undergoing change, by scenario. Units are kg ha -1 yr -1 CO 2 – C...... 23 Table 4.2: Major changes in arable management practice and cropping patterns in England since 1940, which have resulted in a decline in soil organic carbon (SOC) within the top 30 cm of soil...... 24 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 4.3: Values of carbon sequestration for different scenarios that increase sequestration (£ per ha per year)...... 28 Table 4.4: Value of management changes resulting in losses of soil C (£ per ha) ...... 29 Table 5.1: The % change to the standard peak flow flood expected for the change to the proportion of the indicated land uses in a catchment (+10% of catchment area – for catchments of <2,000ha)...... 33 Table 6.1: Yields of arable crops grown on sandy clay loam soil at four levels of Olsen P (Johnston et al. 2001b)...... 39 Table 6.2: Yields of winter wheat given four rates of N at four levels of Olsen P on the sandy clay loam at Saxmundham, 1981-82...... 39 Table 6.3 Reductions in yield due to a drop in soil P reserves ...... 40 Table 6.4: Reductions in yield due to a drop in soil K reserves from Index 2 ...... 41 Table 6.5: Effect of exchangeable K levels on the yields of four arable crops on sandy loam soils at Woburn (Williams, 1973) ...... 41 Table 6.6: The effect of reduced pH on the yield of three arable crops (Johnston 1975, Bolton 1977, Johnston and Whinam 1980)...... 42 Table 6.7: Reductions in yield resulting from various degrees of soil compaction...... 44 Table 6.8: Reductions in output value due to a drop in soil P or K reserves (£ per ha) ...... 46 Table 7.1: Changes in autumn and spring MWDs due to the given drivers...... 50 Table 8.1: Drivers of marginal change in a soil’s ability to attenuate pollution from heavy metals, phosphorus (P), sediment, pathogens, nitrate and acid deposition - and units and scale of marginal change...... 56 Table 8.2: Estimates of the financial costs of water pollution (£m per year)...... 64 Table 8.3: Cost of least cost measures to reduce N, P losses on arable farms (£ per ha per year) P loss reduction (kg per ha per year)...... 65 Table 8.4: Marginal cost of changes in attenuation capacity for N and P ...... 66 Table 10.1: Benefits of ESA landscapes (£’000s per year) ...... 77 Table 11.1: Methods for valuing the benefits from soil-related policy measures ...... 84 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

1. Executive Summary

1.1 Remit The following remit was requested by Defra.

Research is required to estimate an economic value for the identified soil functions and to suggest a methodology for developing a total soil valuation. This was to involve a number of specific objectives, outlined below. The researchers should develop and, if they feel it necessary to best achieve the overall objective, extend the specific objectives.

The research should consider (at least) the following soil services: • Water infiltration (environmental interaction) – flood defence, sewerage system, roadside swales • Nutrient cycling (food and fibre) – crop production • Carbon storage and sequestration (environmental interaction) – climate change • Supporting construction (platform for construction) – engineering solutions required to cope with particular soil types/situations (this may value one soil type over another) • Natural attenuation of pollution/contamination (environmental interaction) – contaminated soil remediation • Archaeological site/artefact protection (heritage)– rescue of damaged/disturbed archaeology

Benefits of soil should be analysed in terms of their magnitude and on whom they fall.

Monetary estimates should be presented, where possible, together with an appraisal of the functions not reflected in the monetary estimates. Consideration should be given to how these values can be integrated to provide an estimate of the total value of soil functions. Attention should be given to avoiding double counting of values where this is a potential problem.

In order to be relevant for policy-making, monetary estimates should be presented in terms of the marginal value of the soil function i.e. the value yielded by an additional unit of the service provided. The reason for this is that policy-making involves deciding where an additional pound of limited public money should be spent. The proposal should set out explicitly how the values could be applied in a cost-benefit analysis.

1.2 The Project This project was commissioned as a first step in the development of an appropriate methodology to estimate an economic value of soil functions and to include a literature review. The methodology is based on the literature review and broader relevant knowledge of environmental valuation within the project team.

There are many national and European initiatives that are raising the profile of soil protection and use. However, for soil to be fully appreciated in today’s society it must have a socio-economic value. Whilst this is an accepted concept, techniques for deriving the value of a given site have yet to be developed.

Page 1 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

The majority of valuation research carried out to date on soil has concentrated on the costs of soil erosion, the value of which gives some indication of the value of good quality soil. Other hints as to the value of soil are available through studies attempting to value global ecosystem services. Very little research has been carried out to date attempting to value the many ecosystem services provided by soil specifically, such as nutrient, water and carbon cycling. Given that more carbon dioxide is stored below the ground than in the atmosphere and above ground biomass combined, the value of carbon sequestration benefits of soil could be of great importance for future policy.

Soil Functions In addition to the above soil services, the project team looked at support for ecological habitat and biodiversity.

In all cases, the baseline valuation was taken as Best Management Practice (BMP), such that deviations from this should be expressed as a change in the marginal value of the soil function.

Soil Valuation The value of soil is determined by the stream of benefits it provides to society. Benefits are reflected in people’s willingness to pay (WTP) as indicated by real market transactions or through evidence on preferences from which values can be derived.

Private market transactions largely determine the value of soil as an input to food and fibre production, and historically much of the interest in soil has centred in its value for agriculture. This is also true for soil as a base for housing and construction, where development value of land is important. However, some soil services, and especially the support of landscapes and biodiversity and the protection of archaeological interest, are not valued in markets. Here, the public directly benefits from the supply of services that are often freely available as an element of ‘nature’. The value of such soil services, as indicated by the welfare gain to the public, is more difficult to determine because of the absence of observable prices.

1.2.1 Summary of Methodology A two-stage approach was used to derive marginal values for soils from existing datasets. First the technical effects were quantified. This was done by identifying the main drivers of change in soil functions and then estimating the per unit impact of each driver on the relevant soil function. In most cases the effects are context dependent and the environmental effects may be indirect (e.g. through impacts on linked environmental assets such as water resources or biodiversity).

The second stage was one of economic valuation. Three methods were used: • Applying market prices where the effects are on marketed outputs (e.g. nutrient cycling impacts on crop production); • Using the methods of environmental economics to value environmental impacts (evidence from economic valuation studies); and • Cost-based approaches where the aim is to estimate the cost of compensating measures for the loss in soil function.

Of these the cost-based measures are less informative because they indicate nothing about the benefits or losses to public welfare. However, in many cases a lack of direct valuation evidence necessitated a cost-based approach. This can be use to provide a lower bound estimate of public benefit.

Page 2 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Market estimates of value are relevant to two of the soil services- nutrient cycling for crop production, and supporting construction.

Soil is rarely traded or used in isolation from other inputs. Normally it is closely bound up with a location, a climate and a set of institutional and legal parameters.

Public and private valuations will differ because of government intervention through taxation and subsidy and due to the externalities of farming.

Hedonic models have been used to explain the spatial variation in land prices in terms of variation in the attributes of land. These models provide implicit prices (marginal valuations) for the attributes, which typically include land quality. Again, the prices are those that reflect private benefits from land and these are not necessarily the same as social benefits.

Studies show that differences in soil attributes that affect production are reflected in land prices. However, with market failure the implicit prices for soil attributes are not much of a guide for policy decisions. They will overestimate soil value to the extent of production-linked subsidy. In so far that use of a soil necessarily leads to external costs from pollution its implicit value will also be over-estimated.

From a national welfare perspective, the marginal value of a change in soil services will be measured by the social gain or loss. The social value of changes in farm output has proved difficult to measure in a situation in which there is pervasive farm support linked to farming activities (as under the CAP prior to de-coupling) and where farm externalities are sizeable (Saunders, 1996). However, under the de-coupled context of the SPS, farmland prices will provide a reasonable basis for the marginal social value of farmland before taking into account externalities.

Valuation of non-market (environmental) outputs - Soils provide many services that are of value to the public, but for which there are few or no developed markets, although the market for carbon is beginning to develop. The support for water infiltration, biodiversity and landscape, carbon sequestration and archaeological protection are the clearest examples. However, economic valuation data relating to soil protection are sparse.

Cost-based approaches - Cost based approaches can be used to provide a lower bound estimate of society’s WTP for a policy where an existing policy exists. Cost based approaches are also useful where the aim is to select a least cost mechanism from a set of options that can deliver on an existing policy (e.g. to meet the requirements of the WFD or climate change polices). They are less useful for policy assessment since this requires an estimate of policy benefits.

The report has focused on methods using existing datasets and an indication of the types of methods that would be appropriate for deriving new non-market benefit values. However, it is not possible to define a ‘best’ method in the abstract. It would be necessary to fix a budget, an objective and a context (policy and location) in order to do this. One development of this study that we would recommend as a further project would be to take a number of policy issues relating to soil services and work up the corresponding technical datasets together with an outline valuation study that would produce benefit estimates. These could then be used within a policy cost- benefit framework.

Page 3 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

1.2.2 Analysis of Soil Functions Carbon storage and sequestration - Soil has an important role in storing carbon. The quantities of sequestered carbon in soils in England are broadly in equilibrium after a long period of decline. However, soil carbon levels can be increased especially through changes in the management of arable land and through changes in land use. At a social value of £70 per tonne the value of sequestered SOC is significant, with a present value of up to £1,380 per ha. Accounting for other changes in carbon sequestration (GHG and energy effects) associated with the change in use of the soil increase this to a maximum of £3,160 per ha where arable land is converted to woodland. Drainage of peat in both the uplands and lowlands releases carbon and the value of the losses are substantial. Other changes have relatively minor effects.

The environmental function that soil performs as a store for carbon that may otherwise contribute to elevated atmospheric levels of carbon dioxide is one of the most important to place a value upon. In the UK, there have been concerns expressed recently that there has been a perceived decline in SOC levels in both arable soils and those under grassland. The estimated losses to grassland are neither so large nor readily attributable to an identifiable management pressure, but because grasslands comprise about 70 % of agricultural land further loss must be arrested to secure a valuable storage pool for carbon.

Water storage and flow mediation - Several drivers can change the ability of soils to hold water. In locations subject to risk of flooding these effects can be important. It is not possible to generalise about the change in risk or cost that follows from a change in soil use, because impacts are location specific and require individual assessment. The major effect is through paving over and construction, and to a lesser extent, afforestation. Impacts may be transferred. Consideration of whole catchment effects is therefore essential where any downstream part of a catchment is subject to flood risk. The effects of forestry are less clear. Where flood risk is an issue site-specific assessment are needed and MAFF (2001c) indicated the appropriate cost-benefit procedures that should be used.

Nutrient cycling - For nutrient cycling, nitrogen is different from P and K because organisms in the soil or in symbiotic association with crop roots naturally fix N, whereas P and K vary due to weathering of the soil particles, deposition from the atmosphere or additions by farmers. The value of the marginal loss in soil function can be estimated by applying fertiliser replacement value to N and market prices to the lost output for P and K. Where major reduction in the P or K index or in pH occur the losses are substantial for crops sensitive to the loss in soil service. It emphasises the importance of maintaining the nutrient cycling capacity of the soil through adequate applications of maintenance P and K, and through liming. Policy intervention to correct any loss of soil function would only normally be appropriate where there is evidence of market failure. It is not clear where such failure might occur so long as producers are clearly informed about the consequences of reduced nutrient cycling capacity.

Supporting construction - The impacts of soil deterioration and deviation from BMPs on soil as a base for construction appear to be generally small, mainly affecting site accessibility. Since the use of land for construction is a market activity the case for policy intervention to protect soils is weak. The only case for intervention would rest on any expected negative effects on the environmental services provided by soils rather than their value for construction.

Page 4 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Natural attenuation of pollution/contamination - The valuation of changes in a soil’s natural attenuation of pollution/contamination will depend on which pollutant is affected. The main driver for reducing the ability of soil to attenuate pollution on contaminated land is the contamination itself. The soil is able to attenuate pollutants through absorption on to the soil exchange complex and by degradation processes. However, once the threshold for absorption, rate of degradation or soil organism tolerance is exceeded, the attenuation function is effectively lost. The soil also has a value in its ability to bioremediate, in other words use soil physical, chemical and biological processes to degrade contaminants (particularly hydrocarbons) to less toxic compounds. The main driver affecting this function will be climate change. Changes in leaching of heavy metals, N and P and movement of pathogens may have implications for water quality in associated catchments. This will be the case where there are existing risks in relation to the WFD, where drinking water abstraction takes place or where eutrophication is an environmental issue. We were only able to quantify the possible marginal costs of a loss of attenuation capacity for N and P using policy costs under the WFD. These varied substantially depending on the extent of the change in attenuation capacity but could be high where BMP was not adhered to. Changes could also occur in the ability of soils to buffer acid deposition, but it was not possible to quantify these effects, although remedial action through liming is a straightforward measure at least on accessible land.

Support for ecological habitat and biodiversity - Valuing the biodiversity and associated landscapes supported by soils is immensely complex. A number of studies have used the methods of environmental valuation to quantify the benefits provided by sites high in biodiversity. These include both use and non-use values of the biodiversity. Use values are determined by the number of visitors and the benefits they derive from visits, whereas non-use benefits are mainly determined by the perceived uniqueness or rarity of a site or species and the degree to which it is threatened.

Archaeological site/artefact protection (heritage) - The best indicator of potential impacts in the archaeological protection function of soil is the extent of changes in land use and, in particular, ploughing and cultivation depth, and areas protected in an archaeological plan. There is a lack of evidence on the value of the benefits or disbenefits from soil management impacts on the archaeological heritage. Benefits (or disbenefits) from changes in management that affect the archaeological interest will mainly be non-use because such unprotected sites are typically on private land not accessible to the public.

1.2.3 Appropriate Methodologies for Measuring Benefits The economics literature mainly concentrates on explaining variation in the market price of land in terms of its underlying attributes, for example the changes in capital values caused by CAP reform. Although hedonic pricing studies do provide marginal values for land ‘quality’ attributes there are few in the UK and they are neither especially reliable nor useful for the present study.

Cost-based valuations can generally be used if there is adequate technical knowledge. However, they provide a lower bound estimate of public benefit and a very limited guide for policy development. That is unless soil policy measures form a subset of a wider set of options for an existing policy when cost-based valuations can be used for selecting least-cost mechanisms. Direct estimates of benefit that relate to soil-based measures are sparse, but can provide a wider context for the benefits from soil functions (e.g. with biodiversity as an output).

Page 5 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Where the aim is to generate new information on environmental benefits from soils, non-market valuation methods need not be used (see reviews by Garrod and Willis, 1999; Pearce et al ., 2006). In the context of soil services only stated preference approaches (contingent valuation and choice modelling) are relevant. Within choice modelling, choice experiments are likely to be preferred because they are welfare consistent and give implicit prices in the attributes of choice. In the soil context they could be used to price benefits from individual soil services.

Page 6 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 1.1 summarises our view on the most appropriate methods for determining the social benefits from changes to soil services and functions. The table also indicates areas where further research would enable improvements to be made in the estimation of benefits from changes to soil functions. Where new research is proposed for the valuation of non-marketed soil services, only stated preference methods are appropriate. The choice of method (CV or choice experiments) would depend on the precise objective and context.

Table 1.1: Appropriate methods for valuing the benefits from soil-related policy measures

Appropriate methodology for Data requirements to facilitate Function assessing benefits from soil –related estimation of benefits from soil- measures using existing data related measures Carbon storage and Estimates of the social cost of None recommended sequestration carbon provide a good basis for analysis. Water storage and A cost-based approach (damage Benefit assessment is highly flow mediation prevented; effects on water site/catchment specific. Stated treatment costs) coupled with preference methods are public risk assessment is appropriate.. appropriate. Nutrient cycling Market effects are best analysed Marginal changes in microbial using available market price and nutrient cycling due to diffuse cost information. contamination Supporting A cost-based approach is the None recommended construction appropriate methodology for assessing changes in a soil’s ability to support construction. Natural attenuation In the absence of direct A case can be made for the of pollution/ measures of public benefit, cost- evaluation of the benefits from contamination based approaches are measures to protect the pollution recommended. Publc benefit attenuation capacity of soils. A applies to contaminated land for meaningful policy context would journeys to and from landfill. need to be defined and the Bioremediation and natural methods of environmental attenutation also reduces topsoil valuation applied. Stated use. preference methods are appropriate. Information on the extent of contaminated soils for re-use following remediation would be valuable. Archaeological At present, only cost-based This is an under-researched area site/artefact methods can be used because of which would benefit from protection a lack of relevant valuation economic valuation of the public research. Social costs provide a benefits from soil lower bound estimate of the protection.Stated preference public’s WTP for such protection. methods are appropriate. Support for Valuation data are not sufficiently More economic valuation is ecological habitat detailed to provide a basis of needed to quantify the benefits and biodiversity assessing specific soil-related from protecting the biodiversity measures. The social cost of functions of soils. Stated existing policy measures preference methods are provides a lower bound estimate appropriate. of public benefit.

Page 7 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

1.2.4 Information Gaps and Further Research The main requirements for technical and economic information to improve the evidence base are: • Linking technical measures of changes in water quality to benefit estimates. New benefit measures specified in technical and economic terms may be needed. • Improved catchment water use models must be used to derive the marginal change in water storage and flood peak flows for land use changes in catchments. Also required is associated cost information for specific catchment contexts and estimates of the benefit from reduced flood risk. • A more direct valuation of soil type for crop production should be investigated that accounts for fertility and ease of working, which currently are not reflected in land values. • Information on the marginal change in soil functions across the range of soil types. Even for biomass production, where trials have been carried out on a number of soil types, information is not available for the impact of reducing soil reserves on all soil types. • Specific figures for the loss of pollutants from soil, as a result of various pressures. Figures provided by Defra project ES0203 (The cost-effectiveness of integrated diffuse pollution mitigation measures) limit marginal losses of P, N and FIOs to specific farm models, two soil types and a single climate. There are not sufficient data to make robust predictions for a variety of land use-soil-climate scenarios, which hinders extrapolation to the national level. Information is particularly lacking for the marginal change in the ability of soil to attenuate release of pesticides and FIOs from land. • More work is needed to improve the information base relating to potentially important drivers of afforestation and its impact on flood risks (afforestation and construction). • There is a lack of linkage between water quality indicators (e.g. N, P, faecal indicators) and estimates of public benefit from changes in these indicators. Economic valuation of soil-related measures is required. More stated preference research is needed to identify the public’s WTP for specific measures that improve water quality. • Benefit estimates for soil and land management policies that protect the archaeological, pollution attenuation and biodiversity capacity of soils (see Table 1.1)

• One development of this study that we would recommend as a further project would be to take a number of policy issues relating to soil services and work up the corresponding technical datasets together with an outline valuation study that would produce benefit estimates. This would take forward the practical application of methods to generate new benefit estimates for soil services.

1.2.5 Policy Implications The main policy implications relate to the environmental functions of soils. Where markets satisfactorily take account of soil values (nutrient cycling in agricultural production, support for construction) the case for intervention to protect soils is much weaker. The information available supports the case for soil protection to minimise losses of sequestered carbon and take account of carbon effects in other land-based environmental polices (e.g. afforestation). Impacts on flood risk especially from

Page 8 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development paving over appear important but impacts are catchment-specific. It proved difficult to quantify the marginal costs of losses in attenuation of pollution. However, there was a case for intervention to facilitate BMP in cases of major deviation from BMP. There was evidence of public benefit from the soil’s function in supporting biodiversity and this is likely to be most important in areas of high ecological value.

There is no universal best method that can be applied routinely to produce new benefit valuation data. This is a developing area and further work is needed to design specific studies. In this, numerous issues need to be considered including the soil(s) in question, the scale of the study (local, national), and whether the aim is to value changes in soil services and/or aggregate benefits from specific soil policy measures.

Page 9 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

2. Introduction This project was commissioned as a first step in the development of an appropriate methodology to estimate an economic value of soil functions. The aim was to produce a literature review and methodology as a suggested guide to valuation.

The research was required to consider (at least) the following soil services: • Carbon storage and sequestration (environmental interaction) – climate change • Water infiltration (environmental interaction) – flood defence, sewerage system, roadside swales • Nutrient cycling (food and fibre) – crop production • Supporting construction (platform for construction) – engineering solutions required to cope with particular soil types/situations (this may value one soil type over another) • Natural attenuation of pollution/contamination (environmental interaction) – contaminated soil remediation • Archaeological site/artefact protection (heritage)– rescue of damaged/disturbed archaeology

In addition to the above, we have added the following function: • Support for ecological habitat and biodiversity

It is recognised that not all of these functions have been studied to a level to enable an appropriate valuation to be carried out, but the main aims of the project are to review the literature and develop a methodology. This methodology will be based on the literature review and broader relevant knowledge of environmental valuation. The emphasis on the valuations is in terms of marginal change through marginal change in soil functions in order to direct policy decisions. This will help inform policy-making decisions on limited public funds.

Monetary estimates have been sought where possible, together with an appraisal of the functions not reflected in the monetary estimates. The methodology suggests how these values can be integrated to provide an estimate of the total value of soil functions. Attention has been given to avoid double counting of values where this is a potential problem.

This research was required to estimate an economic value for the identified soil functions and to suggest a methodology for developing a total soil valuation. The researchers set out to develop and, where considered necessary to achieve the best overall objective, extend the specific objectives.

In all cases, the baseline valuation was taken as Best Management Practice (BMP). This term covers a whole range of land and enterprise management practices that in the UK are represented by the three guides provided by Defra as the Code of Good Agricultural Practice for Soil, Water and Air.

Deviations from BMP should be expressed as a change in the marginal value of the soil function and other issues. Such deviations from BMP could be the result of pressures on soil management arising out of social or policy drivers of change that effect the way society views and values land use. The matrix shown in Table 2.1 below is an example of the types of drivers of change and resulting pressures on management that effects soil functions.

Page 10 Table 2.1: Matrix of examples of Drivers of change (x) and resulting pressures on soil that are relevant to the economic valuation of the soil functions These relate to the six functions ascribed to soil by the First Soil Action Plan for England (Defra, 2004). W = water storage, C = carbon storage & sequestration, & P = natural attenuation of diffuse pollution.

Soil functions Category of Drivers of Change Pressure on soil/land Nutrient Support for Archaeological Supporting Drivers W C P cycling Biodiversity site protection construction

Ecological Climate change Increased winter rainfall X X X X X Droughtiness in summer X X X X X Temperature increase X X X X Policy Water Frame-work Extensification X X X X X Directive Anti - air pollution Reduced aerial deposition of XXX measures N, S & acidity Provision of amenity areas Affore-station X X X X Habitats Directive Change of vegetation/usage X X GAEC and CSF Set-aside margins X X X X X Reduced tillage methods X X X X Extensification of livestock X X X X X Bio-renewable energy Change to biomass crops X X X X X policy Social Population change Urban fringe development X X X X X Farm economic pressure Change in manage-ment X X X X X X Creation of amenity areas Affore-station X X X X X X Protection of archaeo- Archaeological management XXX logical heritage plans

Page 11 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

The literature review is combined within the chapters rather than being taken as a separate task

Chapter 3 begins by discussing the various methods of valuation that are appropriate to the different soil functions and why they are suitable. Generally, this is because some have a market value and others do not, hence other approaches are needed.

The following chapters look at each soil function in turn. Whilst this report brings together a wide range of functions that are very different in character, we have tried to use a common format, looking at the technical characteristics, the drivers of change and valuation method.

Chapter 11 concludes by bringing together the methodology and looks at the evidence on the marginal valuation of soil functions. This is followed by a section on information gaps and further research.

Page 12 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

3. Marginal Valuation of Soil Functions and Services

3.1 Valuing Soil Services The value of soil is determined by the stream of benefits it provides to society. Benefits are reflected in people’s willingness to pay (WTP) as indicated by real transactions in competitive markets or through evidence on preferences from which values can be derived.

Private market transactions largely determine the value of soil as an input to food and fibre production, and historically much of the interest in soil has centred in its value for agriculture. In the same way, the market determines the value of soil as a base for housing and construction, where the development value of land is important. The value of a site for development lies mainly in its location and not in the characteristics of the soil. The soil may alter the basic value if it is problematic for development, particularly if this is linked to its hydrology, but these can be engineered out if the overall value of development is sufficient. The value of land as indicated in market exchanges may not represent its value to society if there are external costs or benefits (e.g. through the support of biodiversity), or if government intervention affects land prices (e.g. through farm subsidies). In such cases private values would need to be adjusted to account for these effects.

However, some soil services, and especially the support of landscapes and biodiversity and the protection of archaeological interest, are not adequately valued in markets. Here, the public directly benefits from the supply of services that are often freely available as an element of ‘nature’. The value of such soil services, as indicated by the welfare gain to the public, is more difficult to determine because of the absence of market prices. It is necessary to resort to the techniques of environmental valuation (Garrod and Willis, 1999; Pearce et al ., 2006) in order to estimate the magnitude of the benefit.

The majority of valuation research carried out to date on soil has concentrated on the costs of soil erosion (Ribaudo & Young, 1989; Ribaudo, 1989; King and Sinden 1988;Dragovich, 1990; Whitby and Adger, 1996) the value of which gives some indication of the value of good quality soil. Other hints as to the value of soil are available through studies attempting to value global ecosystem services (Costanza, 1997, Nunes, 2001). Very little research has been carried out to date that attempts to value the many ecosystem services provided by soil specifically, such as nutrient, water and carbon cycling, (Bateman, 2000). Given that more carbon dioxide is stored below the ground than in the atmosphere and above ground biomass combined, the value of carbon sequestration benefits of soil could be of great importance for future policy.

Because there are few data on the non-market benefits from soils a different approach to valuation is often used. This uses the cost of treatment, mitigation or damage control as a lower bound indicator of the gain or loss from an action. Hence, for example, if carbon sequestration in soil reduces the damage from greenhouse gas emissions, the value of enhancing or reducing this service from soil could be estimated from the compensating increase or decrease in the cost of other carbon emission control programmes adopted by government. Another example could be the value of coastal wetland in terms of the cost of a sea wall and other defences to perform the same protective function.

In summary there are thus three possible methods that may be used in order to derive social values for soil and land:

Page 13 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

• Market-based approaches using values from market transactions (usually of land); • Valuation of non-market (environmental) outputs; and • Cost-based approaches.

These are discussed in more detail below.

3.2 Market-based Approaches Market estimates of value are relevant to two of the soil services – nutrient cycling for crop production, and supporting construction (see Chapters 5 and 6). This reflects the fact that agriculture, forestry and development are mainly private sector activities in which land is an important input. For these services from soil we can apply market prices to value the private benefits. However, where environmental benefits or disbenefits are significant, and not valued by markets, these would provide an additional element to the social value (+/-).

3.2.1 Land Valuation - Background Soil is rarely traded or used in isolation from other inputs. Normally it is closely bound up with a location, a climate and a set of institutional and legal parameters. It is this bundle of inputs called ‘land’ that is normally traded rather than the single input ‘soil’.

The present value (PV) model is the basic starting point for the private valuation of land (Lumby, 1988). This is a temporal investment model that explains present value in terms of the future stream of income from the asset:

n i V= ∑ Ri/(1+r) i=1

Where V is the present land price, R i is the net income at end year i, r is the real discount rate, and n the time horizon. This type of model has been widely used to explain changes in land prices over time using time-series data (e.g. Traill, 1979; Lloyd et al ., 1991). Land prices are dominated by the incomes land generates in the private sector, and where land is bought with the intention of letting to a tenant, R i is the rent net of landowner costs. A simplified model generally used with time series analysis is:

V=R*/r

Where R* is the constant perpetual stream of real rent and r is the constant real discount rate.

However, the PV models do not usually perform well especially when current incomes and interest rates are used as the basis for future expectations. Numerous modifications have been proposed in order to account for the effects of factors such as risk, price support policies, tax, development potential and institutional characteristics (owner–occupation or tenancy)(e.g. Currie, 1981).

3.2.2 Intervention Governments intervene widely in the land market in order to achieve a variety of objectives in their agricultural, rural and environmental policies. Output and income- supporting instruments may compete with agri-environmental instruments as governments seek to maximise the public value of the total output from land.

Page 14 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Public and private valuations of land may thus differ because of government intervention through taxation and subsidy and due to the externalities of farming (Barnard et al ., 1997; Pretty et al ., 2000). A number of studies have sought to determine the effect of price support on land values. Traill (1979) estimated that a 1% increase in CAP support prices was capitalised into a 10% increase in land prices. Harvey (1991) estimated that roughly half of support provided to the agriculture sector was capitalised into land prices and rents. If this support were removed he argued that prices would fall by 45%, a reflection of the difference between private and public valuations.

Just and Miranowski (1993) estimated that 15% of the land value was explained in terms of government support payments in the U.S. Weersink et al. (1999) re- formulated the PV model to decompose land prices into two elements – farm output and subsidy. He demonstrated that these are assessed differently by purchasers in the Ontario land market. Government subsidies were discounted at a lower rate than farm production, which implies that they were treated as a less risky source of income. Protection of soils valued for agriculture.

Soils of high quality for agriculture have some degree of protection from development under PPS7, but not absolute protection. It is clear, however, that under the SPS, considerable quantities of land will effectively come out of productive use in agriculture. This will occur progressively over time as full adaptation to the SPS occurs. This land will be maintained to the GAEC standard but produce only environmental outputs. Much of the soil on this land will be relatively poor quality in the uplands, but substantial areas of arable land are also likely to be in fallow because it is not profitable to farm for reasons of low productivity, high cost, low prices or poor location. These aspects have been discussed in recent studies for Defra (e.g. Renwick et al ., 2003; GFA-Race, 2003; Hodge et al . (2006).

Whilst the change to the SPS may not greatly change the social value of the output from soil, the low net private output from much of the UK land area will become more apparent. The case for protecting high quality agricultural land from development use has been one primarily based on the security of domestic food supply. There may be a case for re-assessing the cost to society of providing this element of security given the increasing demands from development and environmental uses which have the potential to provide greater social value. (See for example Section 3.3).

3.2.3 Hedonic Pricing Models of Soil Hedonic models have been used to understand the spatial variation in land prices in terms of variation in the attributes of land. These models provide implicit prices (marginal valuations) for the attributes, which typically include soil quality. Again, the prices are those that reflect private benefits from land and these are not necessarily the same as social benefits.

The usual model follows that of Lancaster (1966) as developed by Palmquist (1989). It considers land as a differentiated factor of production. The price of land is taken to be a function of the attributes of land as follows:

P=f(Z 1, Z 2, Z 3…………….Z n)

Where P is the price per ha of a parcel of land, and Z i represents the level of an attribute of land or an associated attribute such as climate or tenancy status. This approach is usually applied to land rather than directly to soil because market price data are available on land but not for soil. The dataset for estimation typically comprises the characteristics of a range of soils in different locations at a given date.

Page 15 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Variation in soil attributes between sites is captured through the specification of Z i. One merit of this model is that it both gives an absolute price for any levels of Z i but also defines the marginal price of Z i. The coefficients can be interpreted as the shadow (implicit) value in the land market of an investment to change a land attribute (e.g. measures to improve soil drainage or increase soil fertility).

Maddison (2000) applied the hedonic model to agricultural land prices in England and Wales. He used price per unit area as the (continuous) dependent variable and regressed this on a set of variables describing the area, size of dwelling house, climate, population density, etc. Soil quality was indicated by the official classification, which divides soils into 7 grades.

Table 3.1: Implicit prices of soils in England and Wales (from Maddison, 2000)

Land class variable Implicit price (£ per acre) ALC grade 1 (relative to Grade 7) +142.2 ALC grade 2 (relative to Grade 7) +375.4 ALC grade 3 (relative to Grade 7) +237.0 ALC grade 4 (relative to Grade 7) +12.2 ALC grade 5 (relative to Grade 7) -341.8 ALC grade 6 (relative to Grade 7) +351.8

Note: ALC is Agricultural Land Classification: Grade 1 is the highest quality for agriculture and grade 7 the lowest.

Although he was able to explain 62% of the variation in land prices, the land quality coefficients do not appear reliable indicators of soil quality (Table 3.1). Grade 6 land has a higher value than grade 1. Numerous reasons may be proposed as to why the model does not give very credible soil values. ALC takes into account soil properties (mainly physical), but also slope, climate and flood risk. Climate is the primary determinant of the ALC grade rather than soil quality. In addition, the set of land prices used reflected not only agricultural values but also residential and development values for which land grade would be less relevant. High-grade land is protected from development by the planning system and this will reduce its market value in a development context. It may also be the case in Maddison’s study that attribute variables were correlated (e.g. climate attributes with soil grade) which would lead to less precise estimates of the shadow prices for soil.

In a Scottish study, Roberts et al . (2002) regressed land prices on size of parcel, land capability class, land cover, environmental designation, and indicators of accessibility, peripherality, population density and occupational class structure. The model was limited by a lack of data on tenancy status and farmhouse/farm building characteristics. Nevertheless, the soil capacity variables were highly significant with higher soil quality related to higher land price. However, as with the Maddison study, soil characteristics were not evaluated in any detail.

Hedonic pricing has been used to price land attributes in the USA and Canada. Research has focussed on soil erosion, drainage and soil conservation in part to determine the extent to which the land market takes into account future impacts on soil productivity (Ervin and Mills, 1985; Miranowski and Hammes, 1984; Palmquist and Danielson, 1989; Lussier et al ., 2001). These studies used a variety of measures of soil quality including pH, topsoil depth, potential erosivity, wetness and soil quality

Page 16 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

indices. Implicit prices are estimated for the characteristics. Miranowski and Hammes (1984) estimate the marginal value of 1 inch of topsoil at $12-31 per acre and a 1 ton per acre reduction in erosion potential at $5.6 per acre. Palmquist and Danielson (1989) found that where a soil was wet enough to require drainage its price was reduced by 25%. As might be expected, soils with a high erosion potential also had lower values. Lussier et al . (2001), for example, using a sample of Quebec land sales estimated shadow prices that were significant for land drainage and pH, but not for other chemical attributes, organic matter or water stable aggregates.

Faux and Perry (1999) used land sale price data to estimate an implicit price for irrigation water in Oregon. They not only demonstrated soil prices varying from $248- $2,918 per acre depending on soil class but that the implicit price of water varied with land class since water was more valuable on higher quality soils. A large number of such studies that determine farmers’ willingness to pay for water have been reviewed by Conradie and Hoag (2004).

These studies show that differences in soil attributes that affect production are reflected in land prices. However, with market failure the implicit prices for soil attributes are not much of a guide for policy decisions. They will overestimate soil value to the extent of production-linked subsidy. In so far that use of a soil necessarily leads to external costs from pollution its implicit value will also be over-estimated. These aspects are discussed further below.

3.2.4 Social Value of Changes in Soil Services From a national welfare perspective, the marginal value of a change in soil services will be measured by the social gain or loss. The social value of changes in farm output has proved difficult to measure in a situation in which there is pervasive farm support linked to farming activities (as under the CAP prior to de-coupling) and where farm externalities are sizeable (Saunders, 1996). However, under the de-coupled context of the SPS, farmland prices will provide a reasonable basis for the marginal social value of farmland before taking into account externalities.

For example, a discrete measure that improves, say, soil water percolation and increases crop output will have a ‘marginal’ value to landowners as follows:

pi*∆q i –c

Where p i and q i are the price and quantity of output i and c is the cost to the farmer of applying the measure. ∆q i is the change in the quantity of output. The marginal social value of a soil measure that increases crop output may be lower than the private marginal value if there is an external cost (E) from increased pollution:

pi*∆ iqi –c- E This attributes the entire external cost to the soil measure. Alternatively, the marginal social value will be increased when there is an external benefit.

From a policy perspective, any estimation of costs and benefits from a soil protection or enhancement measure introduced via an agent (e.g. farmer, developer) would also need to take into account the transaction costs incurred.

3.3 Valuation of Non-market (environmental) Outputs Soils provide many services that are of value to the public, but for which there are few or no developed markets. A number of techniques have been developed to value

Page 17 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

non-market benefits or disbenefits (Garrod and Willis, 1999; Pearce et al ., 2006). These are based on the study of preferences, which are either revealed through actions or stated under quasi-real contexts. In the context of soil services only stated preference approaches are relevant. Pearce et al . classify these into two groups - contingent valuation and choice modelling. Within the choice modelling group, choice experiments are likely to be preferred because they are welfare consistent and give implicit prices in the attributes of choice. In the soil context they could be used to price benefits from individual soil services. The reader is referred to Pearce et al . (2006) for a detailed review of these methods.

Where there is no associated market output (e.g. land managed solely for its contribution to biodiversity), non-market values for the output provide a context for assessing the valuation of marginal changes to soil. However, specific economic valuation of the benefits from soil measures is more precise and informative. Where there are also market outputs (e.g. from agriculture) the environmental values would need to be added to the change in value of the marketed outputs from the soil measures to provide a composite valuation (see 3.2.4).

Soils support a number of environmental functions, including water infiltration, biodiversity and landscape, carbon sequestration and archaeological protection. The extent of valuation research on these topics is rather limited. Few studies relate directly to soils and where they do the context for valuation is usually in terms of the ecosystem services or outputs supported by soil. In addition, it may not be possible to quantify the impact of soil as an input because the output (e.g. biodiversity) is dependent on other inputs apart from soil (e.g. climate, human activity).

One notable study that did attempt to directly value ecosystem services and functions is that of Costanza et al . (1997). They used a classification that included climate regulation, water regulation, erosion control and sediment retention, soil formation, nutrient cycling, waste treatment, and food production. Based on estimates of consumer and producer surplus (the welfare gains above costs), the dominant service in value terms was nutrient cycling, which contributed 51% of the total value of the global ecosystem flows. This related mainly to estuaries, algal beds and tropical forests. Soil formation and water regulation are relatively unimportant at 0.15% and 3.3% of the total. It is difficult to identify the particular role of soil in these services although it clearly contributes to a number of regulatory and other functions.

3.4 Cost-based Approaches Cost-based approaches for valuing soils include the costs of lost output, soil protection and soil remediation. They are less informative than benefit-based measures of value but can give insights where estimates of benefits (see 3.2, 3.3 above) are not available. Cost-based approaches are relevant in the context of the Water Framework Directive (WFD) where any change in pollution levels in water bodies due to changes in the services provided by soils may have implications for policy intervention. Thus, an increase in pollution resulting from a reduction in the ability of soil to attenuate pollution may imply additional costs because of the increased risk to the water body.

When society has made a decision to implement a policy with a given cost, despite not having full information about the benefits that will be forthcoming, it can be argued that the implicit benefit to society is at least the cost incurred. On this basis, cost- based approaches can provide a minimum estimate of society’s WTP for a policy. It might be noted, however, that governments do not always engage in policies where (ex-post) benefits exceed costs. There are examples of a possible failure of some

Page 18 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

environmental policies to pass the cost-benefit test (Pearce, 2005a. 2005b). It is also the case that payments made as voluntary incentives or as compensation for regulatory measures are transfer payments that do not necessarily measure the true social costs incurred. Under flat-rate incentive schemes, for example, all but the marginal recipient is paid in excess of compliance costs. However, these exchequer expenditures are often used as a proxy for social cost in cost-benefit analysis of policy measures in the land sector (e.g. CJC Consulting, 2003).

In this report we suggest that, where direct benefit measures are not available, the change in the cost of delivering on existing framework policy measures such as the WFD be used as a minimum estimate of society’s WTP for a soil measure.

3.5 Marginal Valuation of Soil Services: Conclusions A marginal valuation of a change in the functions or services provided by soils can, in principle, be assessed in one of two ways: • Hedonic pricing methods which derive implicit changes in value in relation to soil characteristics; or • Direct estimation of the impact of a change in soil services by combining technical assessment with existing valuation data or new valuation estimates.

Of these, the scope for using hedonic pricing is limited by the quality and scope of the cross-sectional datasets that available. As indicated in Section 3.2.3 they have largely been used to explain variation in land prices and the effect on land prices of government intervention in agriculture. With the currently available UK datasets, the research by Maddison (2000) would seem to represent the limit on what can be achieved.

The most useful approach is that of direct estimation. This involves (i) a technical element which links the change in soil property to the corresponding change in soil function (e.g. ability to support construction; impact on crop yield), and (ii) a valuation element which applies benefit values derived using the methods described in this chapter. Where it is the intention to estimate new values for environmental services from soil the choice of valuation method will depend on the precise objectives and context for the research.

3.6 Total Marginal Valuation of Soil Services The concept of ‘total soil value’ is not easily interpreted because soil is a joint input (with climate-related inputs, labour and capital) in the delivery of marketed and environmental outputs (see 3.2.1). It is best considered as a marginal concept reflecting the total costs or benefits from a change in soil services brought about by policy or changes in external factors such as climate.

In cases where more than one soil service is affected it is possible to derive a total marginal valuation of the change. Where independence can be assumed, this is the sum of the benefits of the changes in individual services. These could be a mix of market and non-market benefits depending on the outputs affected by the measure. There will be no double counting so long as the services (and their valuations) are separately defined and estimated. This approach is only possible where the technical impacts of the measure on soils services are well defined and there are relevant benefit values for each individual service that changes as a consequence of the policy measure.

Page 19 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

However, the additive approach is quite demanding because of the need for individual values for each soil service. It would be quite a challenging research project to estimate separate environmental benefits although choice experimental methods do provide such implicit values. In practice it would be simpler but less informative to directly estimate the total benefits delivered by the policy measure – for example using contingent valuation.

Page 20 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

4. Carbon Storage and Sequestration

4.1 Technical Review 4.1.1 The Soil Function of Storing Carbon The estimated global soil organic carbon (SOC) pool, 1550 Pg C (1 Peta g = 10 15 g), is approximately twice that of the atmospheric pool (770 Pg C) (Batjes 1998) and 2.5 times that of the biotic pool (610 Pg C). As such, it is in a position of controlling influence over carbon flux dynamics. Therefore the environmental function that soil performs as a store for carbon that may otherwise contribute to elevated atmospheric levels of carbon dioxide is one of the most important to place a value upon. In the UK, there have been concerns expressed recently that there has been a perceived decline in SOC levels in both arable soils (King et al, 2005) and those under grassland (Bellamy et al., 2005). If these are indicative of a general decline in the 562 Tg C (1 Tera g = 10 12 g) calculated to be in the topsoil of the 66,000 km2 of arable land in England and Wales (Smith et al. 2000a), then perhaps 56 Tg C (i.e. 10% of the topsoil pool) is theoretically recoverable by re-sequestration to soil. The estimated losses to grassland are neither so large nor readily attributable to an identifiable management pressure, but because grasslands comprise about 70 % of agricultural land further loss must be arrested to secure a valuable storage pool for carbon.

4.1.2 Factors that Affect the Soil’s Ability to Sequester and Store Carbon Past modelling exercises have sought to predict the effect of various land use and agricultural management changes (designed to mitigate greenhouse gas emissions) on soil carbon content (Smith et al. 1997, 2000a & b; Rickman et al. 2002; King et al., 2004). These changes can be considered to be the drivers of change that are likely to be positive in their impact on carbon storage by soils (they are designed to be so), though many of them can be reversed and have historically been seen as having negative effects on soil carbon storage (King et al., 2005). Basically, any change that increases the organic inputs to the soil system (manures or crop residues), whilst maintaining similar production levels will lead to a new equilibrium being achieved in the soil, thus additions of manure, straw, compost and sewage sludge will tend to have this effect. Although increases in inputs will have some benefits, as this will lead to increased turnover of C, the net benefit in terms of C is small. So also will changes in land use from arable to grassland and woodland, partly because these tend to maintain higher organic returns to the soil, but also because in these cases the output of soil organic carbon is also lowered. This occurs because the soil is no longer cultivated, where cultivation tends to increase the respiratory loss of soil organic matter (the chief means by which it is lost) by opening up fresh reserves to aerobic microbial degradation. Thus changes which reduce the intensity of cultivation also tend to result in increased soil carbon storage.

Other major negative impacts on soil carbon storage are any land use change that seals soil from the natural carbon cycle (construction etc .), which acts to reduce inputs to the soil as above. In addition changes from permanent vegetation cover such as forest or permanent grassland, to tilled land will also tend to increase the oxidation of stored soil carbon as discussed above for cultivation. Similarly, any change to the management of peat or organic soils that results in their drainage will also lead to increased oxidation and loss of the organic matter. A recent example of the latter is the siting of wind farms in remote upland areas where the soils are either blanket peats or have a peaty surface horizon. Peatland is the largest and most important store of soil carbon in the UK, and is particularly vulnerable to damage from increased anthropogenic activity.

Page 21 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Some of the management changes suggested to mitigate carbon dioxide loss run the risk of altering the emission rates of non-CO 2 greenhouse gases such as nitrous oxide, which have been shown to have considerable potential to attenuate any carbon saving in their effect (Smith et al. 2000c; King & Bullard 2001). This should also be borne in mind when considering the economic value placed upon any change in management to sequester extra carbon to soil. Similarly, most of the changes will also incur an alteration in the pattern of energy use, which must also be incorporated into assessments of their overall impact and value.

4.1.3 Drivers that May Lead to a Positive Change in Soil Carbon Storage The land use and agricultural management drivers of a potentially positive change in soil carbon storage can be summarised as those in Table 4.1 which has been taken from King et al. (2004). These represent changes from both arable agriculture and managed grassland which are considered to be pursued in the near to mid-term future. The socio-economic drivers of these changes can be policy driven, such as those that may arise from the requirements to meet water quality standards for the Water Framework Directive (e.g. moves from fertilised cereal production back to extensive grass or woodland), or purely economic within the farm budget, such as moves to reduced tillage systems which incur lower labour and fuel costs. It is not considered likely that the needs for carbon sequestration alone will constitute a likely policy driver of change, but they may be a backdrop concern that is taken into consideration when considering measures to promote changes for other reasons.

Table 4.1 shows for the positive land use and management changes as the range of carbon that may be sequestered to soil (SOC) for a range of soil types. The amounts of carbon saved from fossil fuel emissions (DE and IE); the carbon equivalent of consequent emissions of other greenhouse gases (GHG); and the total carbon equivalent sequestered or saved (TSC).

Page 22 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 4.1: The final carbon sequestration and saving rates applied to each unit area of land undergoing change, by scenario. Units are kg ha -1 yr -1 CO 2 – C.

Change in land use or SOC DE + IE GHG TSC management

Arable to permanent 552 to 828 425 327 to 609 1304 to 1862 woodland

Arable to willow energy crop 552 to 828 304 327 to 609 1183 to 1741

Arable to Miscanthus energy 490 to 734 275 327 to 609 1092 to 1618 crop

Conventional to zero tillage 145 to 235 22 -181 to -84 -14 to 173

Conventional to reduced 40 16 0 56 tillage

Addition of straw residues 532 to 717 0 -61 to -17 471 to 700

Application of additional 610 44 -3 651 sewage sludge

Addition of livestock manure 50 to 208 13 to 25 8 to 25 71 to 258 to arable land rather than grassland

Set-aside field margins on 490 to 734 440 25 to 46 955 to 1220 arable land

Extensification of converting 479 136 0 to 172 615 to 787 break crops to grass in rotation

Extensification with outdoor 479 136 0 to 2 615 to 617 pig breeding on grass in rotation.

Conversion to stockless 479 238 7 to 13 724 to 730 organic management

Conversion to organic 479 296 7 to 10 782 to 785 management with livestock

Grassland to permanent 0 1963 2354 4317 woodland

Grassland to willow energy 0 1842 2354 4196 crop

Grassland to Miscanthus 0 1813 2354 4167 energy crop

Change to clover based 0 196 -33 163 pastures

Conversion of conventional 0 1749 533 2282 to organic dairy management system

Page 23 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

4.1.4 Drivers that May Lead to a Negative Change in Soil Carbon Storage The land use and agricultural management drivers of a potentially negative change in soil carbon storage can be summarised as those in Table 4.2 which has been taken from King et al. (2005), with major changes of drained peatlands and urbanisation added in. These represent mainly historic changes from extensive grassland or semi- natural vegetation to arable agriculture and managed grassland. It is difficult to envisage what socio-economic drivers of change may lead to these being taken up in even the mid-term future, but it is possible that local changes in land use force some of them.

In addition to the changes below, the increased installation of permanent under- drainage systems between 1973 and the 1990s, would also have accounted for a decline in SOC but this was not quantifiable. Such widespread drainage has now ceased, and is unlikely to recommence. The drainage of organic soils however, still continues locally for various reasons, and has more dramatic consequences for the carbon balance.

Table 4.2: Major changes in arable management practice and cropping patterns in England since 1940, which have resulted in a decline in soil organic carbon (SOC) within the top 30 cm of soil Rates of change apply over the time intervals indicated.

Estimated rate of change in SOC (kg ha -1 yr -1 C) Management change Time-frame Decrease

0 – 5 years 5 –20 years 20 – 40 years 40 – 60 years

From permanent 1940 – 2000 -1557 -1045 -108 -108 grass and natural vegetation to arable

Decrease in area of 1965 – 2000 -525 -6.5 oilseed rape

Decrease in area of 1940 – 2000 -750 -9.3 potatoes

Fluctuation in the 1940 – 2000 -1557 -1045 -108 -108 proportion of arable as temporary grass

Drainage of upland 1960-2006 -1500 -1500 -1500 -1500 peat

Intensive 1960-2006 -12600 -12600 -12600 agricultural use of lowland fen peat

Urbanisation (total 1947-1990 -3000 loss) NB. Negative values indicate that negative savings are made, i.e. gaseous emissions of greater CO 2 equivalence are made.

There have been few measurements for the impact of drainage on peat reported in terms of carbon loss. One recent study in New Zealand quotes an average annual loss from 3.7 t ha -1 a -1 with a range of 2.5 – 5.0 t ha -1 a -1 (Schipper & Mcleod, 2002) for

Page 24 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

a drained peat bog converted to dairy pasture. Temperatures, and therefore oxidation rates, are liable to be higher in New Zealand than the UK, but the above loss rate compared favourably with the average for temperate areas reported in a review by Armentano & Menges (1986). Therefore an estimate of the lower end of the range, say 2.5 t ha -1 a -1 would not be inappropriate for UK conditions. However, this is at the higher end of a range quoted for Scandinavian peat in Cannell et al . (1999), 0.2 – 2.9 t ha -1 a -1 , and more than twice the estimate for Scottish peat in the same paper of 1.3 t ha -1 a -1 . The latter estimate was considered a possible underestimate, so a value of 1.5 t ha -1 a -1 as a UK average has been adopted in Table 4.2. Another form of loss common from upland peat areas is that of erosive loss. However this cannot be quantified on any typical areal basis, as it depends entirely on the weather, grazing pressure and human impact forces that cause the loss.

The long-term loss of carbon from drained lowland fen peat soils used in intensive agriculture and horticulture also continues unabated. Cannell et al (1999) also contains an estimated annual loss of carbon from organic lowland soils in agricultural use equivalent to 12.6 t ha -1 a -1 . This huge rate of loss is not of course sustainable and these soils will eventually end up as mineral soils with an SOC content of 2-3 %, but will take about 80 years to do so. It is feasible that other surrounding areas of organic soils could be taken into agricultural use in response to the economic drivers of yield reduction as current soils lose their profitable high fertility. However, there is probably little scope for this as most are already under cultivation, but the above figure has been inserted into Table 4.2 for lowland fen peat.

The remaining driver of negative change in soil organic carbon contents is that of soil sealing and loss upon construction. This is very difficult to quantify accurately, because it is common practice to remove the topsoil and re-use it elsewhere which would save considerable amounts of carbon that would otherwise be lost. However, in their estimates Cannell et al (1999) worked on the assumption that soils converted to urban areas lost all of their carbon, whereas those in suburban areas lost only half, from a presumed previous grass cover. The estimates for the loss of an average 13000 – 19000 ha a -1 of land to urbanisation in the UK, between 1947 and 1990, was 2.9 t ha -1 a-1 , and for 20000 ha of land lost in Northern Ireland between 1970 and 1990, it was 3.5 t ha -1 a -1 . The average loss value to soil sealing has therefore been set at a compromise of 3t C ha -1 in Table 4.2. This however, is a one-off value for total loss and does not continue year on year after sealing like the estimates for losses from drainage or vegetation conversion.

4.2 Valuation of Carbon Storage and Sequestration 4.2.1 Public Benefits from Marginal Changes to Carbon Storage and Sequestration It is now well established that greenhouse gas emissions damage the global environment. Economics research has attempted to quantify the cost to global society of these emissions and hence the benefits from emission reduction. Carbon storage and sequestration contributes to a reduction in net emissions. The benefit from sequestration can thus be inferred from the damage cost of carbon (and other greenhouse gas) emissions. Damage costs have not been derived directly from the public’s willingness to accept damage, or willingness to pay for mitigation. Rather, estimates have been made from indirect modelling of the cost of global impacts (Helm, 2005). The benefits from carbon storage and sequestration are therefore included under the cost-based estimate below.

Page 25 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

4.2.2 Cost-based Approaches Carbon Storage and Sequestration Policy Background Under the Kyoto Protocol, and subsequent agreement amongst EU countries, the UK has a commitment to reduce greenhouse gas emissions by 12.5% below the base year 1 level, on average over the first commitment period, 2008-2012. There is also a manifesto commitment aiming to reduce CO 2 net emissions by 20% by 2010. In addition, following the publication of the Energy White Paper in February 2003, the UK has a longer-term goal to put the UK on a path to reduce carbon dioxide emissions by 60% by 2050 with real progress by 2020.

Carbon sequestration and release is accounted for in the UK Greenhouse Gas Inventory which reports greenhouse gas emissions by source and removals by sinks (NAEI, 2005). The 2004 review of the Climate Change Programme (UK Government, 2004) indicated that the UK was comfortably on track to meet the Kyoto commitment, but without additional measures would fall short of the domestic 20% target. The UK Climate Change Programme is currently being reviewed to determine the action needed to put the UK back on track to meet its domestic 20% target.

Value of Sequestered Carbon: Damage Estimates The social cost of carbon is the monetary value of worldwide damage from the anthropogenic emission of carbon dioxide into the atmosphere. The models used to estimate the social cost of carbon relate emissions to atmospheric changes; atmospheric change to temperature change; and temperature change to damage. This damage includes sea level rise, floods and storm events, the impact of climate change on agricultural production, and effects on population health and disease. Given the uncertainly involved in the impacts it is not surprising that estimates of the social cost of increased carbon emissions vary widely depending on the assumptions used, including the discount rate. Marginal damage costs depend strongly on the coin of discount rate because they reflect the additional future damage from small changes in current emission (Tol, 2005). The models also vary in the weight given to extreme and catastrophic events.

There is also disagreement on how to deal with spatial variation in the impacts of global warming, and especially how impacts on people with different incomes are treated. Results are sensitive to the way in which equity weighting (the weights given to people of different income levels) is applied. These issues are discussed at length in Pearce (2005).

A study by Clarkson and Deyes (2002) for H. M. Treasury estimated the social cost of carbon at £70/tC (with a range of £35 to £140/tC) and this is the value that Defra and other government departments currently use in appraising policies that lead to changes in carbon emissions. It is suggested that these estimates should be increased by £1/tC per year in real terms to reflect the increasing costs of climate change over time.

Nevertheless, this value is considerably higher than estimates made by most other economists. The meta analysis undertaken by Tol (2002a, b) of 103 estimates of marginal damage costs of carbon dioxide emissions; producing a mode of $2/tC, a median of $14/tC and a mean of $93/tC, because of the lognormal distribution of the cost estimates. Studies with lower discount rates and equity weighting had higher

1 1990 is the base year for emissions of CO 2, CH 4, and N 20 and 1995 is the base year for emissions of HFCs, PFCs and SF 6.

Page 26 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

estimates and much greater uncertainties, whilst peer reviewed studies had lower estimates and smaller uncertainties. Tol concludes that climate change impacts are very uncertain, but that using standard assumptions about discounting and aggregation, a marginal damage cost of carbon dioxide emissions of $15 /tC seems justified and that marginal damage costs are unlikely to exceed $50/tC (i.e. £27/tC). Mendelsohn (2005) concludes that the social cost of carbon will rise over time but there is every reason to expect that it will remain below $10 per ton for the next 30 years. Pearce has questioned the Clarkson and Deyes (2002) estimate and concludes that it is too high, preferring a marginal social cost of carbon of £4-27/tC. Brainard et al . (2003) used a carbon cost of £6.7/tC in estimating the sequestration benefits from forestry.

An alternative cost-based approach to estimating the marginal cost of carbon is to assess the marginal cost to government of current measures to control emissions. Pearce (2005) derives a range of values from £16-31/tC depending on the measures and quotes a Defra estimate of the cost of controlling carbon emission as £45/tC. According to Pearce the evidence from policy intervention to control emission is that the £70/tC cost is not currently applied in a consistent way. Defra is currently undertaking a review of values of the social cost of carbon. There is a rationale for adopting the same value for carbon in appraising policies in all sectors of government.

4.3 Benefits from Carbon Sequestration in Soils The right hand columns of Table 4.3 give the social values of sequestered carbon when priced at Defra’s preferred value of £70 /tC (see above). In relation to the changes in SOC alone, the value of the additional carbon fixed by changes in the use of arable land or its management vary from £2.8 to £48.3 per he per year. The highest values are where soil is converted from arable cropping to woodland or biomass production. A permanent change in land use of this type is calculated to add £1,380 2 per ha to the value of soil as a carbon sink.

2 Present value of the annual sequestration derived by discounting at 3.5% per year over an infinite horizon.

Page 27 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 4.3: Values of carbon sequestration for different scenarios that increase sequestration (£ per ha per year)

Value of total Current value of soil sequestered carbon Change in land use or management organic carbon (at (at £70 per t) £70 per t) (£/ha/year) (£/ha/year)

Arable to permanent woodland 48.3 110.8

Arable to willow energy crop 48.3 102.3

Arable to Miscanthus energy crop 42.8 94.9

Conventional to zero tillage 13.3 5.6

Conventional to reduced tillage 2.8 3.9

Addition of straw residues 43.8 41.0

Application of additional sewage sludge 42.7 45.6 Addition of livestock manure to arable land rather 7.5 11.6 than grassland

Set-aside field margins on arable land 42.8 76.2 Extensification of converting break crops to grass 33.5 49.1 in rotation Extensification with outdoor pig breeding on grass 33.5 49.1 in rotation

Conversion to stockless organic management 33.5 50.9

Conversion to organic management with livestock 33.5 54.9

Grassland to permanent woodland 0.0 302.3

Grassland to willow energy crop 0.0 293.7

Grassland to Miscanthus energy crop 0.0 291.7

Change to clover based pastures 0.0 11.4 Conversion of conventional to organic dairy management system 0.0 159.7

Note values calculated from mid-ranges of Table 4.1.

Equivalent values calculated for total sequestered carbon (Table 4.3) vary from £3.9- £302.3 per ha per year. Where carbon is sequestered as a result of a change in land use or management of the soil, the total saving in carbon varies from £3.9 to £110.8 per ha per year. The highest values are associated with conversion of arable farming to woodland or biomass. The highest value of £110.8 equates to a present value 3 of £3,160 for the permanent change in land use.

Table 4.2 gives the losses in soil organic carbon that has occurred due to a number of changes in soils management and use. Table 4.4 values the future loss that would occur if such changes in management were to start now.

3 The present value is the sum of the future stream of effects discounted at 3.5%.. It is the present sum that is equivalent to the future stream. See H. M. Treasury (2006) for a fuller description.

Page 28 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 4.4: Value of management changes resulting in losses of soil C (£ per ha)

Time horizon used Present value of Equivalent loss per Management change in the valuation loss at £70/tC year over horizon (years) (£) (£) From permanent grass and natural vegetation 60 -1,282 -51.4

Decrease in area of oilseed rape 20 -170 -11.9

Decrease in area of potatoes 20 -243 -17.1 Fluctuation in the proportion of 60 -1,282 -51.4 arable as temporary grass

Drainage of upland peat 60 -2,619 -105.0 Intensive agricultural use of 40 -18,835 -882.0 lowland fen peat

Urbanisation (total loss) 5 -948 -210.0

Note: Present values at 3.5% were accumulated over the different time horizons for changes in SOC as indicated in Table 4.2. Thus for a change from permanent grass and natural vegetation the changes in SOC per year over 60 years were costed at £70/tC and discounted at 3.5% per year. This gives a present value of £-1,282 per ha. The annual stream equivalent to this present value at 3.5% is £-51.4 per ha per year over 60 years.

The present values of the losses are priced at £70 per t C over the relevant horizon and discounted at 3.5% (Table 4.4). For example, if drainage of previously undrained upland peat takes place, the loss of stored carbon is valued at £2,619 per ha (£105 per ha per year). The major losses occur with drainage and intensive use of fen peat and drainage of upland peat. The value of these losses exceeds the agricultural or forestry value of the soils. These losses would be increased if the monetary value of C increased over the time horizon.

4.4 Conclusions Soil has an important role in storing carbon. The quantities of sequestered carbon in soils in England are broadly in equilibrium after a long period of decline. However, soil carbon levels can be increased especially through changes in the management of arable land and through changes in land use. At a social value of £70 per tonne, the value of sequestered SOC is substantial, with a present value of up to £1,380 per ha where arable land is converted to woodland. The value of the change in total sequestered carbon (including GHG and energy effects) is maximal at £3,160 per ha, again for conversion to woodland. Very large losses of carbon value are possible especially with drainage or intensive use of peatland. Conversion from permanent grass also has a high negative marginal value. The magnitude of these effects indicates a case for soil protection in some cases although a complete analysis would need to include the opportunity cost of protection in terms of any lost output.

With respect to appropriate methodologies for valuing carbon storage, there is a sizeable body of evidence on the social cost of carbon. Although estimates show wide divergence, government has adopted a value of £70 per tonne for internal policy assessment. Combining such social cost data with technical assessments of impacts on carbon storage provides the best basis for assessing the cost-benefit of policy measures that relate to soil’s ability to store or release carbon.

Page 29 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Policy measures may have other opportunity costs in terms of output foregone. The social cost of such effects would also need to be taken into account to form an aggregate analysis. This is not problematic for marketed outputs but the feasibility of including changes in environmental outputs would depend on the availability of relevant economic valuations.

Page 30 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

5. Water Storage and Flow Mediation

5.1 Technical Review The First Soil Action Plan for England (Defra, 2004) recognised the role that soils play in receiving incident rainfall and holding the water within the soil matrix before finally delivering the water to stream and river outflows of a catchment. The soil’s ability to provide such a reservoir for water storage is adversely affected by any action that may either impede infiltration of water into the surface, or expedite its passage through the soil (installed drainage systems). The soil’s “water holding capacity” will be affected by subtle changes in soil structure and soil organic matter content, but significant marginal change only occurs as a result of serious surface compaction (or capping), or sealing such as that caused by urbanisation. Such practice would be detrimental to storage and flow mediation, whereas other wide scale changes to land use such as afforestation would have more beneficial changes due to a loss in rainfall to the soil surface by canopy interception (and evaporation). In certain catchments this reduction in water supply may however, itself be a problem lower down the catchment.

The main drivers identified in this process are as follows: • Construction – building over the land surface, restoration of soils following development – building, road building, pipeline laying, quarrying and opencast mining involving moving and replacing soil, traffic damage during development. • Management that changes capping and compaction. This can affect intensively grazed grassland as well as arable land and varies according to season. In wet years, autumn cultivations can result in sub-surface compaction and soil erosion. Grazing wet grassland creates surface compaction that reduces water infiltration. In both cases, the condition should be temporary and in the case of arable crops will usually be made good when the next crop is established. In grassland, routine surface treatments are available on most farms and it is in the farmer’s interests to maintain good water infiltration to enable high volumes of grass growth. • Drainage – changes that improve or cause deterioration in drainage. Clearly, there are substitution issues when upland peats suffer from oxidative SOC loss due to drainage and in these cases run-off can be increased as the peat loses water holding capacity. However, drainage of mineral soils generally increases infiltration and reduces the immediate surface run-off. • Afforestation. An increase in afforestation within a catchment will impact the water storage properties of soil and flow mediation through the landscape in several ways. Initially in upland areas with a high rainfall the installation of open- ditch drainage schemes may temporarily increase the run-off from the upper catchment until the trees grow sufficiently to close canopy. For most of the life of the forest, the main impact is to reduce the incident precipitation on the soil surface, typically by about 30% (Cape et al., 1991) for both coniferous and deciduous woodland. This is due to evaporation of precipitation from the tree canopy before it contributes to throughfall to the soil surface. The overall impact in a catchment is to enable the catchment to absorb more storm events before flooding occurs lower down the catchment. A danger occurs however, if coniferous forests are clear-felled, rather than selectively over a number of years, in which case the canopy cover is removed and a large increase in run-off will occur after a couple of seasons when the soil water holding capacity has recharged.

Page 31 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

• Tillage to grassland – improved structure and reduced runoff, or other agricultural land use change, principally grassland to tillage. The latter is less likely to occur under the current situation with the emphasis of subsidies on sustainable management of land including farmland in terms of BMPs and Cross Compliance and the low commodity prices currently available for crops. • Irrigation. This is carried out in the growing season clearly at times of soil moisture deficit. There are occasions, however, when irrigation can be required even though it may be raining, because the expected amount of rainfall is unknown. At such times as well as others, this can lead to soil erosion when soil moisture levels reach capacity. • Climate Change. This may impact water storage in two ways, the first being the straightforward increase in winter rainfall, and therefore also the frequency of extreme storm events. The second, is that as the predicted summer rainfall decrease in some regions and the northerly movement of the limit of some arable crop species progresses, then hitherto grassland areas may become viable for arable cropping and be drained and tilled accordingly.

5.2 Water Storage Potential (WSP) The main public benefit or disadvantage of changes in water holding capacity is in relation to the risk and consequences of flooding. Authorities deal with the risk of flooding by planning and engineering solutions and defences to cope with a certain frequency and size of event. At different points in a catchment, it may be appropriate either to plan for the size of an extreme event that happens only once in a hundred years (habitation), whilst at others to deal with the size of a more frequently experienced event such as that which occurs once in only ten years (moderately valuable horticultural crops). However, at other points anything above the annual flood risk can be allowed to flood (grassland areas)(Bailey et al., 1980). The size and frequency of the storm events that lead to flooding are determined by the climate, but there are two key features that are determined by the catchment that have to be dealt with. These are the amount of water that finds its way into rivers from an event, and also the interval between the event and peak flow in the river.

There are well-accepted methods available for modelling the mean annual flood (QBAR) of catchments for both small agricultural catchments (Marshall and Bayliss, 1994) and those catchments with a much higher proportion of urban land (Hall et al., 1993). The proportion of urban land in a catchment is highly influential as approximately 30% of urban areas are impervious and 70% of the incident rainfall is shed as run-off (Marshall and Bayliss, 1994). Catchment in the UK range from something like 70% urban at the upper extreme to less than 2.5 % for most rural locations.

The equation used to estimate flood risk in the report of Marshall & Bayliss (1994) recognises six influential components; namely, the size of the catchment, proportion of urban land and forest, the winter rainfall acceptance potential of the soils in the catchment and the length of streams in the catchment as well as the overall annual rainfall, that has to be coped with. A model that ADAS runs for small rural catchments based on this equation, enables the details of many of these features, as well as several additional ones (such as the proportion of ploughed and grass land in the catchment) to be changed in turn. The model generates a differential change in the amounts of flood generation for different return periods. This has been done for three “typical” agricultural catchments in each of the three main rainfall regions of England and Wales (<650mm; 650-750mm; >750mm), to scale the potential change in flood generation for changes to those outlined in Table 5.1 below. The change of an

Page 32 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development increase in paved area is thought to be promoted by an increase in housing of peri- urban fringe areas, whilst those of increased afforestation and decreased ploughed area (tillage) (with a corresponding increase in extensive grassland likely) may arise from changes designed to meet WFD goals. Some changes may eventually arise due to climate change, such as a shorter window of machinery working days in the autumn and spring periods, due to increased winter rainfall, and this may lead to increased compaction risk in arable areas (and also high livestock density areas). This increase in compaction has been simulated in the model by increasing the area of restored land in the catchment, as compaction is the chief constraint on cropping restored soil profiles (King, 1988).

Table 5.1: The % change to the standard peak flow flood expected for the change to the proportion of the indicated land uses in a catchment (+10% of catchment area – for catchments of <2,000ha)

Change in +10% of catchment area Return period Compacted Paved area Afforested area* Ploughed area (year/year) (restored) area (+10%) (+10%) (-10%) (+10%)

1/1 +24% -3% * 0% 0%

1/10 +24% -3%* 0% 0%

1/25 +24% -3%* 0% 0%

1/100 +24% -3%* 0% 0%

*Afforestation was not part of the model used, and these values merely reflect a 30% reduction in rainfall as throughfall over 10% of the area.

The model used for Table 5.1 was primarily designed to operate for stream catchments at the field and farm scale (those tested were 160 – 1,100ha), rather than larger river catchments, and so is relatively unsophisticated. Certainly it is too small to incorporate forestry and this element has only been included at the simplest level of response in a reduction in effective rainfall. This will be inadequate to model the true response of larger blocks of forest in catchments, especially where lateral water flow may be affected. The response to different levels of rainfall proved to be the same in this model and so only generic proportions for all rainfall amounts are given in Table 5.1. Similarly, the difference that a 10% increase in compaction, or decrease in ploughing, makes at this scale is negligible, but may not be so in larger catchments where a more significant area of land is affected. Alternatively, in urban catchments where a higher proportion of the area is already sealed, a small change in compaction say, may have a far greater impact than the insignificant one from the above rural model (the surrounding land being unable to mask its effect).

More sophisticated models do exist and are operated by hydrologists for flood defence planning ( e.g. that described in Marshall and Bayliss, 1994), and these should be referred to for more accurate costing of changes. However, they should be done on a catchment by catchment basis, and Table 5.1 is included to emphasise the large impact that changes in the paved area of catchments has and that even small scale changes in vegetated land use can have an effect on the scale of flood events, irrespective of total rainfall amount.

Changes in water storage and flow mediation may also affect water quality since two main nutrient polluting agents, nitrogen and phosphorus, are partially controlled in

Page 33 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

their impact by the amount of water incident. The amount of water flowing through a catchment is of key importance in determining whether the water draining to groundwater breeches the 11 mg L -1 EU limit for N as well as the concentration of surplus N in the soil (a dilution effect). Similarly, the intensity of rainfall events will partially determine the degree of particulate detachment that gives rise to P run-off and soil features that increase the infiltration capacity of a soil will also mitigate P erosion problems on sandy soils. Water quality issues are more fully discussed in Chapter 7.

5.2.1 Valuation of Marginal Changes in WSP As indicated above, various drivers can affect WSP and the risk of flooding and its cost. In principle a change in the risk of flooding can be valued in a number of ways: • The public’s WTP to avert an increased risk. • The estimated damage cost of the change in probability of flooding. • The cost of mitigation to reduce flood risk.

There is a substantial literature dealing with the modelling and valuation of flood risk (e.g. Flood Hazard Research Centre, 1995; Penning-Rowsell and Fordham, 1994). Most economic studies focus on the cost-benefit of investment in measures to reduce losses from flooding rather than the value of changes in WSP. Benefits are typically assessed in terms of reduced losses to property but the analysis is complicated by the probabilistic nature of flooding events. MAFF (1999, 2001a, 2001b) produced a series of guidance notes on flood and coastal defence including methods of appraisal of options for flood defence. These include a detailed methodology for the economic appraisal of projects that reduce flood risk. Defra is currently developing a flood risk and management strategy (Defra (2005c).

The Environment Agency has produced Indicative Floodplain maps to show the areas of land at risk from river and coastal flooding. Whilst every case should be treated in relation to the specifics of the location MAFF (2001b) did distinguish between five types of land use at risk from flooding which cover a range of cost-benefit situations. The highest cost benefit is on Band A (developed urban land with a high risk of flooding). The lowest is Band E (low-grade agricultural land, often grass). This suggests that, in relation to changes in WSP, the main economic effects will be on Band A land where the principal driver of changes in WSP is large-scale construction. However, the MAFF land classification has to be used with caution. Catchment characteristics may be important since influential land use changes may be several miles upstream of the location of risk. It follows therefore that changes in the WSP of land higher in a catchment could have implications especially for urban areas lower down the catchment. Table 5.1 indicates that this is only expected to apply to construction (paved area) or possibly afforestation. Entec in its study on the environmental consequences of increased housing supply has noted that significant new housing development can increase flood risk in other down-gradient areas even when the development is sited remotely from a recognised flood plain (Entec, 2004). This reflects the blocking of flow paths for water and displacing areas into which floodwater would otherwise have been stored.

The particular case of forestry as a driver of changes in WSP has recently been assessed as part of the Defra review of forestry evidence (CJC Consulting, 2005). The Forestry Commission (Gregory et al , 2003) recognises that forests ’might afford some protection against localised flooding’ but that ‘overall…….the mix of management practices, species and tree ages found in large forests means that they have minimal effect on downstream flooding’. However, the FC argue that carefully

Page 34 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

designed and managed wet-woodland in low lying flood plain areas could slow down flood waters (Forestry Commission, 2000). The review concluded that the impact of forestry on flood risk is generally minor. More work is needed to improve the information base relating to forestry and flood risks.

Whilst reiterating that every case needs to be modelled and evaluated in the context of the specific physical and economic characteristics of the catchment, the major impact on flood risk of changes in WSP derives from paving over that reduces water percolation.

5.3 Conclusions Several drivers can change the ability of soils to hold water. In locations subject to risk of flooding these effects can be important. It is not possible to generalise about the change in risk or cost that follows from a change in soil use, because impacts are location specific and require individual assessment. Estimation of the cost implications (through changes in water supply, flood damage and flood risk) provides the best route for assessing the benefit of a change in soil’s ability to store water and mediate flows.

The major effect is through paving over and construction, and to a lesser extent, afforestation. Impacts may be transferred downstream. Consideration of whole catchment effects is therefore essential where any downstream part of a catchment is subject to flood risk. The effects of forestry are less clear. Where flood risk is an issue site-specific assessment are needed and MAFF (2001b) and Flood Hazard Research Centre (2005) indicate the appropriate cost-benefit procedures that should be used.

Page 35 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

6. Valuation of nutrient cycling (food and fibre) for crop production

6.1 Technical Review The soil has a value in its ability to cycle and store nutrients and supply these to a growing crop. Nitrogen (N) is fixed from the atmosphere by soil bacteria and is supplied from the soil through the break down and mineralisation of organic matter. Nutrients are supplied to the soil through atmospheric deposition, plant residues and the spreading of organic manure, artificial fertiliser, and other materials to land. These nutrients can be supplied to crops thanks to the ability of the soil to store nutrients and cycle them back into forms that are available to the crop from year to year.

The first section in this chapter investigates the soil’s ability to cycle N from the atmosphere and to supply N from organic reserves. The following sections investigate drivers that reduce the soil’s ability to provide nutrients for crop production through storage and cycling into plant available forms. These drivers include the running down of soil P and K reserves, and reductions in pH that can reduce soil microbial and structural health for crop production.

6.1.1 N Fixation Nitrogen is assimilated from the atmosphere by naturally occurring soil bacteria and other micro-organisms. There are two types of micro-organism that carry out this function: • Those that fix N in association with plants and supply N to the plant (known as symbiotic bacteria) • Those that fix N independently of plants (known as free-living N fixing organisms) Plants that support N-fixing bacteria in their root nodules are known as legumes, and are able to fix large quantities of N from the atmosphere, as a result of this symbiotic relationship. There are many natural symbiotic relationships. Of these, associations involving legumes and Rhizobium bacteria are by far the most important agriculturally. The amount of N that can be fixed by legume-rhizobia systems can range from 60 to 500 kg N/ha/yr. (Defra project OF0316). Lucerne grown as silage can fix up to 500 kg N/ha/yr., while a more typical figure for a White clover/grass mix is 100-200 kg N/ha/yr. There are a number of management systems that influence the amount of N fixed. However, for a particular legume species there is usually a close relationship between dry matter yield and the quantity of N fixed. A number of models have been developed to estimate the amount of N fixed by legumes in particular circumstances. Watson et al. (2002) summarised the factors that models tend to include for predicting N-fixation. Not all models take account of all the factors, but the main ones include: • Yield of legume and grass • Yield of grass-only reference crop (or some other indicator of soil N supply) • % legume in mix • Years after establishment of grass/legume mix • N content of legumes plus grass • Correction for fixed-N in stubble and roots

Page 36 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Simple yield-based models, based on crop type and average yield data, could be used to estimate the amount of N fixed in soils nationally. More complex models could be used where data is available. Appendix 2 of Defra project SP0523 suggests that a grassland sward, actively managed to contain 30-40% White clover, typically fixes about 200 kg N/ha/yr. The value of this N fixation in fertiliser replacement alone (based on current Ammonium nitrate fertiliser prices of £150 per tonne) is £86/ha. This fixed N replaces the need for fertiliser N and reduces the emission of greenhouse gases associated with fertiliser production. Fixing 200 kg of N eliminates the need to emit 200 x 0.98 = 196 kg ha -1 a - 1 -1 CO 2–C from fertiliser production (assuming 0.98 kg CO 2-C kg N (based on figures in Bullard & Metcalfe, 2001). In addition, Defra report SP0523 suggests that fixation by clover in permanent grassland swards can reduce N fertiliser inputs from 300 to -1 -1 -1 100 kg ha , and save the equivalent of -33 kg ha a CO 2-C via reduced N 2O fluxes. In contrast to symbiotic bacteria, free-living N fixing micro-organisms are thought to contribute much lower amounts of N to crops. Estimates of annual rates of fixation by free-living organisms in agricultural soils are in the range of 1-5 kg N/ha/yr. (Loomis and Connor 1996). Thus, it may be possible to maintain production with the correct balance of species in a grass sward, such that no yield differences will be produced. The valuation of N cycling to crop production will then be in terms of fertiliser replacement values. This is not so for most arable crops. In leguminous crops such as peas and beans, no artificial N is provided in most cases and they leave residual N useable by the following crop. Oilseed rape and potato crops also leave residual N in the soil. Common practice is to assess the ability of a previous crop to leave N residues, which will have a value in terms of their equivalent artificial N fertiliser.

6.1.2 Soil Nitrogen Supply Nitrogen in soil organic matter is unavailable to higher plants. Processes that transform these materials into mineral forms (nitrate and ammonium) that are available to plants are therefore very important parts of the nitrogen cycle. Soil organic matter can be viewed as a reserve of N that can be made available to plants through processes of mineralisation and nitrification. In reality soil organic matter is in a constant state of flux. N is incorporated into soil micro-organisms through a process called immobilisation and organic matter is depleted through processes of decay and mineralisation. Nevertheless, the processes of mineralisation and nitrification, carried out by two separate populations of soil micro-organism, are the key steps in supplying N to plants. The amount of N supplied to plants by mineralisation and nitrification depends largely on the amount of organic matter in the soil. Organic soils with around 10% organic matter content can release 70-100 kg N/ha, while peaty soils can release 150-200 kg N/ha. However, for the majority of mineral agricultural soils with low to average organic matter content (3% or less) the amount of mineralisable N is considered to be small and not practically significant (MAFF, 2000a). The greatest change in soil N supply will occur as organic matter is oxidised from peaty soils. When soil organic matter content falls from 20% to 10%, the soil changes from being a peaty soil to an organic soil and the annual soil N supply can be halved (MAFF, 2000a). If climate change gives rise to warmer, drier summers and the drains on productive peaty soils are maintained, the soil N supply function will continue to decline with the organic matter content.

Page 37 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

6.2 Supporting Crop Production The marginal value of soil for biomass (food and fibre) production can be expressed as the value of lost produce when farmers or land managers deviate from best management practice (BMP). Such deviations from BMP can change soil properties in the long and short term and can reduce the ability of the soil to recycle nutrients and support crop production. Two examples of long term change are exchangeable phosphorus levels and organic matter content, while an example of short term change is surface compaction. The examples of long term change in particular do not merely concern the ability of the farmer to make use of soil nutrients. They change the quantity and form of soil nutrients that are present for the farmer to manage, and therefore significantly impinge on the ability of the soil itself to support crop production. The nutrient cycling of P and K is different from that of N, since none will be fixed from the atmosphere by soil micro-organisms and there are no gaseous losses and reductions in reserves will reduce the natural ability of the soil in a number of ways described below. The following paragraphs represent the situation for arable agriculture. The deviations from best practice covered in this section are: • Allowing soil P and K reserves to run down below maintenance levels • Allowing soil pH to fall below the optimum level of 6.5 for arable land • Causing surface compaction • Causing sub-surface compaction • Allowing soil metal limits to exceed the Sludge (Use) in Agriculture 1989 limits

The following sections present the marginal value that is lost if soils are allowed to deteriorate. Values are given for the amount of lost production in t/ha, but also in the case of P, K and lime, the value of the lost soil reserves.

Another potential driver for change in the crop production function is climate change. However, previous studies have shown that no great change in the ability of soil to produce biomass is anticipated under current climate change scenarios. There may be geographical shifts in the capacity of soil to grow certain crops, but the overall biomass production potential is not expected to increase to any great extent. For example, the amount of maize grain produced is likely to increase due to rising temperatures and an extension in the geographical range for maize growing (Defra 2005, CC0272). More irrigation may be required on sensitive crops due to lower summer rainfall quantities. However, no great regional shift in the capacity of soils to support crop growth is expected, rather, the production of one crop type may be substituted by another.

6.2.1 Effect of Decreasing Soil Olsen P Status Although nutrient balances for livestock farming sectors are still positive, there has been a gradual decline in the overall P balance for cereals, oilseed rape, potatoes and sugar beet since 1974 (Johnston and Dawson, 2005). Since 1995 the balance for P has been negative on many fields where organic manures are not applied. This change in the overall P balance has been reflected in a small decline in the proportion of soils in the higher P Index categories, although any decline in the amount of extractable P will take many years to occur (Withers et al. 1994). However, in the following sections we investigate the possible impact on arable crop yields should extractable P and K reserves be allowed to run down.

Page 38 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Over the years, several studies have drawn attention to how difficult it is, even at a single site, to be precise about how cereals respond to changes in the level of soil extractable P (Johnston et al. 1985; Arnold and Shepherd, 1990). However, it is now generally accepted that there are ‘critical’ levels of soil P below which optimum yields are not likely to be achieved (Johnston and Dawson, 2005). In general, there will be no loss of yield due to a lack of plant-available P, provided there is sufficient P in readily available and less readily available pools. However, if Olsen P is allowed to drop below 10 mg/kg (i.e. into ADAS index 0), typical yield reductions of 25-75% are possible across a range of crops. For this study, we have used data provided by Johnston (2001) and Johnston et al. (2001a), where the following yield reductions were observed with a drop in Olsen P from 20 mg/kg to 5 mg/kg: • Potatoes 25% reduction • Sugar beet 25% reduction • Cereals 25-50% reduction

These assumptions were derived from experiments such as those carried out on sandy clay loam soil at Saxmundham, Suffolk (Johnston et al. 1985). Potatoes, sugar beet, spring barley and winter wheat were grown in rotation during 1969-76 on soils with Olsen P values ranging from 48 mg/kg to 7 mg/kg. Table 6.1 presents the results.

Table 6.1: Yields of arable crops grown on sandy clay loam soil at four levels of Olsen P (Johnston et al. 2001b)

Olsen P, mg/kg Crop 48 30 17 7 Yields, t/ha

Potatoes, tubers 43.7 40.1 38.3 30.6

Sugar beet, sugar 6.57 6.53 5.97 4.25

Barley, grain 5.18 5.01 4.81 4.47

Wheat, grain 6.42 6.65 6.06 5.00

Table 6.2 presents evidence of the reduction in the yield of winter wheat achieved by the addition of a given rate of N fertiliser, as soil Olsen P is reduced. At the optimum N fertiliser rate (160 kg N/ha) yields were reduced by 22% from 10.12 t/ha at 30 mg/kg Olsen P to 7.85 t/ha at 5 mg/kg Olsen P.

Table 6.2: Yields of winter wheat given four rates of N at four levels of Olsen P on the sandy clay loam at Saxmundham, 1981-82

N applied, kg/ha Olsen P, mg/kg 80 120 160 200 Yield of grain, t/ha

30 9.32 9.64 10.12 10.25

19 9.37 9.83 10.25 10.30

10 8.46 9.14 9.10 9.34

5 7.75 7.88 7.85 8.08

Page 39 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 6.3 presents the reduced yields for sugar beet, wheat and barley when nutrient reserves were allowed to drop from ADAS Index 2 to ADAS Index 1 or 0.

Table 6.3 Reductions in yield due to a drop in soil P reserves

Average P Index 1 P Index 0 Crop yield Lost yield Lost yield Lost yield Lost yield (t/ha) (%) (t) (%) (t)

Sugar beet, sugar 9.3 10-20% 1.0-2.0 30-75% 3.0-7.0

Wheat, grain 7.8 10-12% 0.8-1.0 20-25% 1.5-2.0

Barley, grain 6.1 5% 0.30 10-65% 0.6-4.0

In addition to lost yield, the reserves of major nutrients have a value in themselves. For large areas of the country these nutrient reserves have been built up over many years. In recent years, many farmers have reduced P and K fertiliser applications and allowed nutrient reserves to run down (Goodlass and Welch 2005, Johnston et al. 2001b), possibly driven by the need to reduce their costs of production. This can result in a gradual reduction in Olsen P levels (Johnston 2001). The rate of decline depends on many factors including the initial Olsen P value, and the size of the less readily available pool of P. However, experiments have shown that when fertiliser P is not applied under normal farming systems, Olsen P levels are halved in approximately 8 to 10 years (Johnston et al. 2001b).

If Olsen P levels are allowed to run down, large amounts of P inputs over several years would subsequently be required to increase reserves to maintenance levels. The exact amounts are not well defined. However, as a broad guide, to increase the soil P reserve by 10 mg/litre (About one Index level at Olsen P Index 1) would require around 400-600 kg/ha of phosphate (MAFF, 2000a).

For a cereal crop, the loss in value associated with a reduction from soil P Index 2 to soil P index 0 would equate to 1.8 t wheat/ha/yr. and the requirement to apply around 400-600 kg/ha of water soluble phosphate applied over a number of years.

6.2.2 Effect of Decreasing Soil Extractable K Status For potassium, the pattern in terms of reduced recent use and crop response to K reserves is similar to P. The area of cereals receiving no K at all has increased significantly in recent years. Indeed, 40% of the winter wheat area sown in 2002 received no K fertiliser (PDA, 2005). According to the Potash Development Association, soil K deficiency can lead to: • Lower yield • Less efficient N response • Reduced 1000 grain and specific weights • Weaker straw • Increased disease susceptibility

The literature indicates that low K reserves (Index 0) can result in a yield reduction of up to 75%. The PDA (2005) suggest that reducing potassium reserves from K Index 2 to K Index 0 can reduce cereal yields by 20-60%. Table 6.4 presents the reduced

Page 40 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development yields for potatoes, winter wheat and spring barley where K reserves were allowed to drop from Index 2 to Index 1 or 0 (PDA, 2005).

Table 6.4: Reductions in yield due to a drop in soil K reserves from Index 2

K Index 1 K Index 0 Average Crop Lost yield Lost yield Lost yield yield (t/ha) Lost yield (t) (%) (%) (t)

Potatoes 40.0 5-30% 2.0 – 12.0 20-40% 8.0 – 16.0

Wheat grain 7.8 5-50% 0.4 – 4.0 50-60% 4.0 – 4.7

Barley grain 6.1 5-50% 0.3 – 3.0 30-50% 1.8-3.0

Experiments on sandy loam soils at Woburn support these figures. Yields of four different crops declined as exchangeable K (Kex) decreased, as shown in Table 6.5.

Table 6.5: Effect of exchangeable K levels on the yields of four arable crops on sandy loam soils at Woburn (Williams, 1973)

Kex , mg/kg Kex , mg/kg Kex , mg/kg % yield reduction Crop 311 131 36 Index 3 to Yields, t/ha Index 0

Potatoes, tubers 44.3 25.2 10.1 77

Sugar beet, sugar 7.32 5.36 2.80 62

Barley, grain 4.37 4.07 2.82 35

Oats, grain 5.04 4.49 4.62 8

If K Indices are allowed to run down, large amounts of fertiliser K inputs over several years will be required to increase reserves to maintenance levels (i.e. Index 2). The exact amounts are not well defined. However, as a broad guide, to increase the soil K reserves by 50 mg/litre (about one Index level at K Index 1) would require around 300- 500 kg/ha of K 2O (MAFF, 2000a). If soil P and K reserves were allowed to run down simultaneously, the effect on yield would be compounded. At P and K Indices of 0, it is fairly safe to assume that yields would be reduced by 50-75%. The marginal loss in soil value can be calculated on this basis.

A number of specific datasets could be used to assess the spatial extent of soils that are prone to losing nutrient reserves of extractable P and K: • The Representative Soil Sampling Scheme (RSSS) provides an indication of current trends in soil nutrient status • The British Survey of Fertiliser Practice presents recent trends in fertiliser use. Combined with RSSS data it may be possible to estimate changes in extractable nutrients for particular regions and sectors

Page 41 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

• LandIS datasets can be used to assess the spatial extent of non-potassium releasing clays (i.e. Carboniferous clays). The soils formed from these older deposits will be most susceptible to the running down of extractable K.

The marginal change in soil value can be assessed in terms of reduced yields and reduced amounts of extractable P and K. The value will depend on the economic value of the produce and fertilisers concerned.

6.2.3 Effect of Reduced pH Optimum arable crop yields are not possible when the soil pH falls far below 6.0-6.5. This is partly due to the direct effects of the increased hydrogen ion concentration, but mainly due to indirect effects. Principal amongst these is the effect of soil pH on the concentration of aluminium in the soil solution (Russell 1973).

The value of the soil can therefore be estimated in terms of the amount of produce (in kg/ha) that is lost if lime is not applied on non-calcareous soils to maintain optimum pH levels. Table 6.6 presents the reduced yields for potatoes, winter wheat and winter barley when the soil pH is allowed to fall to various levels, based on a number of long-term liming experiments (Johnston 1975, Bolton 1977, Johnston and Whinam 1980). This information is based on established pH related ‘critical’ thresholds for crop yields. For example, one would expect a 90% reduction in winter wheat yield if the soil pH were to drop from 6.5 to 4.0. Based on average yields, this would represent a drop in yield from 7.8 t/ha to 0.8 t/ha.

Table 6.6: The effect of reduced pH on the yield of three arable crops (Johnston 1975, Bolton 1977, Johnston and Whinam 1980)

Average yield Lost yield (t/ha) Crop (t/ha) pH 4.5 pH 5.0 pH 5.5 pH 6.0

Potatoes 40.0 3.0-7.0 1.5-3.0 - -

Winter wheat 7.8 6.0-7.0 3.0-4.0 0.0 0.0

Winter barley 6.1 5.5 2.5-3.0 1.2-1.8 0.2-0.3

A long-term liming experiment at Rothamsted (Johnston and Whinham 1980) showed that reducing the soil pH from 6.5 to 5.5 resulted in a 2 t/ha loss in yield for winter barley. Increasing the soil pH from low levels had the inverse effect. An Agricultural Lime Association (ALA) trial (http://www.aglime.org.uk/technical15.htm) showed that raising the soil pH from 5.1 to 6.7 increased crop yields in the four subsequent years. In year 1, sugar beet yields were increased by 9.2 t/ha; in year 2, spring barely yields were increased by 0.7 t/ha; in year 3, sugar beet yields were increased by 3.6 t/ha; and in year 4, spring wheat yields were increased by 0.2t/ha. The total value of the increased crop yields resulting from the application of 11 t/ha of screened limestone during the winter of 1986-87 was £418.00. The lime application cost £99/ha.

The value of a soil’s lime reserves can be related to the pH level of non-calcareous soils. Soil pH levels have been increased over the years to pH 6.5 for the majority of arable land and 6.0 for improved grassland. The value of the lime in these soils can be estimated in terms of the amount of lime that would be required to raise the pH from a ‘typical’ unlimed level of 4.0-4.5 to the appropriate target level. If the pH were allowed to fall well below optimum levels it would take several years to restore the soil to optimum conditions. This is due to the fact that lime generally takes a number of

Page 42 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

months to neutralise acidity. Naturally acidic clay soils have a greater buffering capacity than naturally-acidic sandy soils, and hence more lime is required to increase their soil pH.

To raise the pH of a clay soil from 6.0 to 6.5 requires about 3 t CaO/ha or 6 t ground limestone/ha (MAFF, 2000a). By comparison, to raise the pH of a sandy loam soil to the same extent only requires about 2.5 t CaO/ha or 5 t ground limestone/ha.

The rate at which soils lose lime depends on four main factors: • Initial pH • Use of nitrogen fertilisers • Annual drainage volumes (a function of soil type and annual rainfall) • Annual crop off take

Gasser (1973) noted that the higher the soil pH the greater the lime loss rate. Losses also depended on soil type, amount of through drainage and fertiliser use. At a soil pH of 5.0 the range of losses was from 79 to 157 kg CaCO 3/ha/yr, and at pH 7.0, the range was 314-628 CaCO 3/ha/yr. Chambers and Garwood (1998) found larger CaCO 3 losses, because of greater fertiliser N use in recent years. Losses of lime ranged from 40 to 1270 CaCO 3/ha/yr, and the relationship between initial soil pH and CaCO 3 loss was:

Annual CaCO 3 loss (kg/ha) = (pH value x 405) –1790

A greater degree of the variation in CaCO 3 losses was explained if N fertiliser use was incorporated into the equation. Most of the nitrogen from ammonium-based fertilisers, - including urea (which has an initial alkaline reaction), is oxidised to nitrate (NO 3 ). Nitrification is an acidifying reaction, so the addition of N fertilisers can result in soil acidification.

+ - + NH 4 + 2O 2 → NO 3 + H 2O + 2H To neutralise the acidifying effect of one kilogram of N from ammonium nitrate fertiliser takes around 1.8 kg ground limestone (Skinner et al. 1992). The acidifying effect of ammonium sulphate is even greater. The amount of ground limestone required to neutralise 1 kilogram of N as ammonium sulphate can be as high as 5.4 kg.

When the effects of the various factors are combined, it is possible to calculate that grassland receiving 200 kg ammonium nitrate/ha, with leaching and crop removal of 400 kg lime/ha, will lose over 0.75 t lime/ha/yr. (Skinner et al. 1992).

To conclude, the total marginal change in soil value where lime is not applied on non- calcareous soils (which represent c.70% of agricultural land in England and Wales) is made up of the value of lost yield (due to reduced pH) and the value of lost lime in soil drainage, crop removal and the use of N fertilisers year-on-year.

6.2.4 Surface Compaction Severe surface compaction can occur when: • Heavy machinery is used with high ground pressure tyres at harvest • Machinery travels on wet soils at harvest • Wet soils are worked at drilling time

Page 43 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

The impact on yields is a function of the surface area that is compacted and the degree of compaction (Li et al 2006). Compaction inhibits seedling emergence, reduces aeration, reduces drainage and impedes roots. In this report, we express the degree of compaction in terms of its impact on yields.

Heavy compaction can give rise to yield reductions of 20% from the single pass of a loaded tractor (Ball and Ritchie 1999) to 50% where heavy machinery passes a number of times (Canarache et al. 1984) (Table 6.7). We have therefore assumed that various degrees of compaction can give rise to yield reductions that vary from 10 to 50%, depending on the increase in bulk density and the proportion of the surface area compacted.

Table 6.7: Reductions in yield resulting from various degrees of soil compaction

Yield loss, t/ha Average Crop yield (t/ha) Mild Moderate Severe (10% yield loss) (30% yield loss) (50% yield loss)

Wheat 7.8 0.8 2.3 4.0

Barley 6.1 0.6 2.0 3.0

Some studies have attempted to model the impacts of surface compaction and this is something that could be developed further for different soil types. For example, Canarache et al. (1984) found that for three of the four locations studies, grain yield of maize linearly decreased by 13 kg ha −1 (or 0.18% of the control plot yield) for each 1 kg m −3 increase in bulk density. Also, Kirkegaard et al. (1993) suggest that predicting yield losses resulting from compaction requires a modelling approach that incorporates the effects of compaction on root growth and crop water use on different soil types.

6.2.5 Sub-surface Compaction As is the case for surface compaction, sub-surface compaction is also produced by heavy machinery, high ground pressure tyres and working the soil when wet. Also, ploughing the soil at the same depth over many years can cause compaction as a “plough pan”. This can be alleviated by the use of tines, which should be set to an appropriate depth, about 2.5 cm below the base of the compacted layer.

In terms of limiting yields, sub-surface compaction is not thought to have a large effect in most situations. Soane et al. (1987) found that yield benefits associated with sub- soiling soils, with identifiable sub-surface compaction, were limited to: • Sandy soils • With spring crops • In years of moderate to severe drought

Even in years with a spring drought, spring crops that were showing signs of moisture stress early in the crop cycle recovered to yield normally when rain arrived later in the year. Compaction in heavier soils is normally alleviated through natural cracking in the spring and early summer, as clay aggregates shrink on drying. In addition, winter crops have longer to establish a viable root system and can exploit cracks that are present in the autumn/early winter and open up in the spring and early summer.

Page 44 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Severe sub-surface compaction on loamy or heavier soils is normally isolated to headlands and fields where field machinery use has caused severe damage at harvest through uncontrolled and heavy traffic.

A typical yield loss for sugar beet where subsurface compaction has occurred on a light-textured soil in a year of moderate to severe drought would be around 10% (5.3 t/ha). With an average sugar content of 17%, this equates to a sugar yield reduction of 0.9 tonnes/ha.

One can get a good picture of the extent of the soils that are vulnerable to yield reductions caused by sub-surface compaction by using combined land use and soil data sets. The NATMAP1000 percentage soil series map could be combined with the 1km 2 spring crop land use map to provide an estimate of the area of land and by association, the amount of crop that could potentially be lost as a result of deviation from BMP (related to sub-surface compaction).

6.2.6 Heavy Metals Microbial populations play an important role in recycling and providing nutrients to crops. Research has shown that heavy metals in the soil can impact on the ability of soil bacteria to perform this function and on the ability of crops to produce optimum yields (Vigerust and Selmer-Olsen 1986). This section investigates the impact of heavy metals on microbial activity, crop health and ultimately crop yields.

Since the mid-nineteenth century there have been many trials investigating the impact of soil heavy metal concentrations on crop yields (Vigerust and Selmer-Olsen 1986). These studies have shown that impacts on crop yields are primarily related to heavy metal concentrations in soil solutions rather than in the soil itself, and that applications of sewage sludge increase heavy metal concentrations in the crop in the decreasing order Zn>Cd>Ni>Cu>Pb=Hg=Cr (Smith 1996). A number of study methods have been used, including pot trials, long-term cumulative field experiments, the use of heavy applications of sludge on agricultural soils and the “spiking” of soils with heavy metals.

Davis and Carlton-Smith (1984) grew rye-grass in a sandy loam soil (pH 7.0) in large plant growth containers and used sludges of controlled metal content to achieve a range of heavy metal concentrations in the soil. They concluded that the heavy metal concentration thresholds for a 10% reduction in yield were 319 mg kg -1 for Zn, 105 mg kg -1 for Cu and 221 mg kg -1 for Ni. Similarly, at Gleadthorpe, Bhogal et al. (2003) applied sewage sludges enriched with salts of zinc (Zn), copper (Cu) and nickel (Ni) to a loamy sand (a ‘worst case scenario’ situation) in 1982 and additionally naturally contaminated, but high level Zn and Cu sludge cakes in 1986. They showed that yields of both cereals and legumes were up to 3 t/ha lower than on the non-sludge control when total topsoil Zn and Cu concentrations exceeded 200 and 120 mg/kg, respectively. However, these yield reductions only occurred where topsoil ammonium nitrate extractable metal levels also exceeded 40 mg/kg Zn and 0.9 mg/kg Cu.

In contrast, long-term field experiments (McGrath 1984, Carlton-Smith and Stark 1987) showed no crop yield reduction where soil metal concentrations are in the range 635-2200 mg kg -1 for Zn, 239-850 mg kg -1 for Cu, and 42-510 mg kg -1 for Ni. However, Bhogal et al. (2003) make the point that in 1984 white clover was grown for the first time and yields were decreased by up to 60% on the sludge plots compared with the farmyard manure plots, despite the sludge applications having ceased more than 20 years previously (McGrath et al., 1987). White clover is known to rely on a relationship with the soil bacteria, Rhizobia to fix nitrogen from the atmosphere, while previous studies have shown that this ability is reduced when soil concentrations of

Page 45 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

heavy metals (particularly zinc) exceed certain levels (Davis and Carlton-Smith 1984, Vigerust and Selmer-Olsen 1986). The yield reductions mentioned by Bhogal et al. (2003) are therefore most likely to be due to the inability of Rhizobia to fix nitrogen in these metal contaminated soils.

In summary, as long as soil heavy metal concentrations are not allowed to exceed the limits defined in the Sludge (Use) in Agriculture Regulations, yield reductions are unlikely to occur for most crops grown in the UK.

To gain a better understanding of how metals behave in agricultural soils and how they may impact on crop yields and long-term soil fertility, it will be necessary to use long-term experimental sites where heavy metals have accumulated through repeated additions. Such studies are being undertaken in a joint Defra/UKWIR/EA/WAG funded programme of work at seven dedicated field sites in England and Wales (Gibbs et al., 2006).

6.3 Valuation of Nutrient Cycling Since the direct effects of changes in nutrient cycling due to deviations from BMP are on crop yields valuations can be made by applying market prices to the output lost. There may also be indirect effects on water quality of reduced phosphate use. These impacts of changes in phosphate losses were considered in Chapter 8.

As indicated above (Tables 6.3, 6.4) the effect of allowing P and K reserves to fall from maintenance applications is characterised by an increasing marginal loss in crop output as reserves fall. Crop response to fertiliser tends to be flat-topped, and small reductions from BMP will hence have minor impacts on net output. Table 6.8 gives the effect of a more major fall to Index level 1 for P and K. With P the major effect is on the value of sugar beet and wheat output. With K the losses are highly variable both within and between crops depending on situation. Potatoes are particularly susceptible to a loss in output if K levels are reduced.

Table 6.8: Reductions in output value due to a drop in soil P or K reserves (£ per ha)

P K Average Value of Crop yield (t per Price of Lost yield lost output Lost yield Value of ha) output (£ (t) at P (£ per ha) (t) at K lost output per t) Index 1 Index 1 (£ per ha)

Potatoes 40.0 80 2.0 – 12.0 160-960

Sugar beet, 9.3 177 1.0-2.0 177-354 sugar

Wheat grain 7.8 65 0.8-1.0 52-65 0.4 – 4.0 26-260

Barley grain 6.1 63 0.30 18.9 0.3 – 3.0 18.9-189

Note: 2006 prices based on Farm Management Pocketbook 36 th edition. Sugar beet at 17.5% sugar.

The value of the marginal loss in soil function at non-optimal pH can also be assessed by applying market prices to lost output. As indicated in the technical review, the lime required to maintain soil function varies with drainage volumes, crop off take and the use of N fertilisers. As an example of the effects we use the data on yield reductions

Page 46 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

from Table 6.6. The marginal loss as pH drops from 6.5 to 5.0 is £120-3004 per ha for potatoes, £195-260 per ha for wheat and £69-76 per ha for barley. These are substantial and emphasise the importance of maintaining the nutrient cycling capacity of the soil through adequate liming.

The economic effects of soil compaction depend on a soil’s vulnerability to compaction. Applying market prices to the reduction in yield can again be use to estimate the loss in output value. In the example given in section 5.1.4, the loss in sugar beet yield from sub-surface compaction (0.9 t sugar per ha) equates to a loss in value of £159 per ha.

With respect to the impact of heavy metal contamination on crop yields the technical review noted that the threshold levels for yield reductions depended highly on the experimental methods used. However, the review concluded that marginal impacts on output were expected to be zero so long as the relevant sludge disposal regulations were followed.

6.4 Conclusions Losses in crop yield can be expected as a result of deviations from agricultural BMPs (see section 8.1.6). Allowing soil P and K reserves to run down below critical thresholds can reduce crop yields by 25-30% for winter wheat, 15-60% for winter barley and 15-30% for grass. For P, the appropriate critical value is at least 20 mg/kg Olsen P, while for exchangeable K it is not possible to give a value as it depends on the amount of K in the fixed (accumulated in clays from fertiliser and manure residues) and lattice (original K in clay minerals) pools. Reduction in P reserves can have major effects on crop revenue through yield reductions, especially for sugar beet and wheat. With K the revenue losses are more variable depending on crop and location. Potatoes are particularly susceptible to a loss in output if K levels are reduced.

Allowing soil pH to fall below critical thresholds can reduce yields of potatoes by 15%, and winter cereals by up to 90%. Winter barley is the most sensitive cereal, with yield reductions evident below a soil pH of 5.8. The economic effects are substantial and emphasise the importance of maintaining an appropriate pH to optimise soil functions.

The amount of yield reduction caused by surface compaction depends on the degree and extent of the soil compaction, but can range from 10 to 50%. Yield reductions caused by compaction below the ploughed layer are restricted to spring crops on light soils in years of moderate to severe drought, for example sugar beet yields could be reduced by around 10% or 5.3 t/ha.

Potentially toxic elements will only reduce crop yields if concentrations are allowed to exceed the Sludge (Use) in Agriculture 1989 limits by several hundred milligrams per kg of soil. As long as the limit levels are not exceeded, we would not expect any significant reductions in yield. Crop loss or failure is therefore only likely to be restricted to flood plain areas within catchments that have a long history of heavy metal mining.

Climate change is not expected to give rise to any great shift in the ability of soils to support vegetative growth.

The value of a marginal change in soil function can be estimated by applying market prices to the lost output. Where major reductions in the P or K index or in pH occur

4 Crop prices are those given in Table 6.8.

Page 47 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development the losses are substantial for crops sensitive to the loss in soil service. It emphasises the importance of maintaining the nutrient cycling capacity of the soil through adequate applications of maintenance P and K, and through liming. Policy intervention to correct any loss of soil function would only normally be appropriate where there is evidence of market failure. It is not clear where such failure might occur so long as producers are clearly informed about the consequences of reduced nutrient cycling capacity.

Page 48 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

7. Supporting Construction – Accessibility and Soil Handleability in Relation to Particular Soil Types/Situations

7.1 Technical Review Construction and development can have several impacts on soil, from physical damage due to poor soil handling to contamination that reduces the ability of soil to perform primary functions. Construction also results in ‘soil sealing’, that is, covering of soil with impermeable materials such as buildings, or the compaction of the soil surface thereby reducing permeability.

For the purposes of this project we consider supporting construction to mean the soil’s function in enabling land restoration and the construction of pipelines, roads and buildings. From a construction point of view, the engineering properties of the lower subsoil are important in terms of bearing capacity, shrink-swell, and plastic limits. However, buildings and roads are not usually constructed directly on the soil itself. Indeed, it is common practice for construction sites to have topsoil and subsoil removed, at least in part, at the commencement of the site development process (Defra 2005a).

The soil influences access to construction sites and the machinery and time required to remove, handle, store and replace soils (Defra 2005b, MAFF 2000b) by the rapidity with which it will dry and its ability to support these activities. In the case of pipeline construction, soil quality is particularly important, as it influences a number of factors: • When the site itself can be accessed • The strength of access trenches • How successful the land restoration will be • Net change in land quality post-restoration

Access to Construction Sites Strong, well-structured and free-draining soils can be accessed for longer periods than soils that are in a poor condition. Arable land does not allow as much traffic as land in permanent grass when construction operations start. In the spring, soils with good structure dry more quickly than poorly draining soils. In the autumn and winter, soils with natural or induced slowly permeable layers do not drain well and give rise to saturated layers either within the soil or at the surface. This can restrict the start date for working the soil and reduce the number of machinery work days (MWDs) overall. Machinery work days have been calculated using the method developed by Smith (1977) that takes into account the median period during which soils are wet (i.e. at field capacity) and the nature of the soil itself.

The drivers that can either reduce or increase the number of MWDs include: • Drainage • Failure to fully implement BMP • Changes in land use including arable to permanent grass • Climate change

Each of these drivers will influence the soil water balance and the number of days that the soil is at or below field capacity, thereby enabling soil handling to take place.

Page 49 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Effective drainage will increase the number of MWDs, while drainage deterioration has the opposite effect. Failure to fully implement BMP can damage soils, particularly where compaction is concerned. This can restrict drainage, delay the start of operations and reduce the number of MWDs. Bare soils evapo-transpire less water than those with a vegetative cover. Permanent set-aside or extensive grassland will therefore marginally dry out the soil and increase the number of MWDs. Finally, UKCIP 02 climate change scenarios suggest that in the future we can expect summers that are hotter and drier and winters that are warmer and wetter (Hulme et al. 2002). How this effects the number of MWDs will depend on how rainfall is distributed through the season. If early autumn becomes drier we could see an increase in the number of MWDs due to a delay in the return of soils to field capacity. However, if winters and early spring become wetter, we could see fewer MWDs due to soils drying out later.

Deviation from BMPs will increase soil compaction and reduce permeability. For the purposes of this study we have considered this to be analogous to a change from Clifton series to Hallsworth series where mean annual rainfall is 650mm. The Clifton series contains imperfectly drained soils with a loamy texture, whereas the Hallsworth series contains poorly drained soils dominated by clay and a slowly permeable layer near the surface. A shift from Clifton to Hallsworth series is analogous to increased compaction, as the compaction of a Clifton series soil will reduce permeability and infiltration rates and would result in it behaving more like a soil of the Hallsworth series in terms of its ability to accept rainfall. Table 7.1 indicates that where Clifton soils are degraded or compacted, autumn MWDs in an average (1959-78) year could be reduced by 16 days and Spring MWDs could be reduced by 11 days. The number of MWDs that are affected is totally dependent on what soil series is used as the baseline. For example, if the sandy loam Quorndon soil series (well drained, but susceptible to compaction when wet) was used as the baseline situation, annual autumn and spring MWDs (in a “Normal” year) could be reduced by 57 and 26 days respectively.

Table 7.1: Changes in autumn and spring MWDs due to the given drivers

Drivers Autumn MWDs (Sep-Nov) Spring MWDs (Mar-Apr)

Drainage deterioration* -46 -14

Deviation from BMPs** -16 -11

Climate change*** -41 -14 * changes in the number of MWDs that could occur as a result of drains being allowed to deteriorate, when compared with soils where drains are maintained. ** - changes that could occur as a result of increased soil compaction and reduced permeability, when compared with well structured moderately well-drained soil. *** changes in the number of MWDs that could occur as a result of wetter winters.

The changes in MWDs that could occur as a result of climate change are derived from comparing the MWDs for one soil association (in this case Dunkeswick) under two different annual rainfall scenarios. These two scenarios are taken from the Soil Survey of England and Wales Bulletin No. 10 (Jarvis et al. 1984), which uses the 1959 to 1978 meteorological dataset to derive mean “normal” years for monthly rainfall and one in four year “wet” years. Using this data, the difference in the number of autumn MWDs between a “normal” year and a “wet” year is 41 days (51 in a “normal” year and 10 in a “wet” year), while the difference in the number of spring MWDs is 14 (i.e.

Page 50 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

14 in a normal year and 0 in a wet year). This gives us an indication of the scale of change that could occur under various climate change scenarios.

Of course, if spring is drier and earlier under the UKCIP 02 climate change scenarios, this would increase the number of spring MWDs. However, uncertainty is high, and all we have attempted to demonstrate in this report is the scale of change, in terms of the number of MWDs, that could occur from relatively small changes (e.g. the one in four 1959-78 wet year becoming the mean year) in volumes and patterns of annual rainfall.

Where soil removal and handling best practices are not adhered to, this would not result in cost to the construction engineers in terms of lost time. However, it could result in costs in the form of compensation to the farmer for a reduction in land quality and reduced yields. In the urban setting, the same compensation issue does not arise, and the cost of dealing with various soil types is a very small and insignificant proportion of the overall construction costs. However, on unsealed urban sites, where planning restrictions are enforced and contractors are obliged to adopt best practice, the start of construction is limited by the same ground condition issues that arise in rural areas.

7.2 Valuation of Supporting Construction 7.2.1 Public Benefits from Marginal Changes to Supporting Construction Marginal changes in the “handleability” of the soil will not give rise to any public benefits. Any public benefits will be unaffected, as contractors will have to achieve the same standards during the construction phase and for restoration projects. The costs are those assumed by the contractor in the extra time and energy required to strip, store and replace soil to current planning standards. In rural areas, if codes of practice are not adhered to, the costs reside in the compensation payable to farmers for lost production due to soil structural damage.

7.3 Economic Analysis For the majority of construction work any variation in a soil’s ability to support development will have a very minor impact on land values. This reflects the substantial premium that development land typically has over any other uses. This premium is determined by local demand for domestic or commercial building within the constraints imposed by the planning system. The technical evidence is that, for the most construction activity, soil condition has little effect on development costs as indicated in land prices. The exception is where access for construction is a significant cost element, and this may apply, for example, to pipelines and windfarm construction.

However, since construction is a market activity any variation in soil value for construction should be accounted for in land prices and rents. The main variation in soil-related cost will reflect differences in soil type and rainfall. Impacts of changes in a soil’s ability to support construction will be minor. In terms of the construction, it would be the cost of lost days for having plant and men doing nothing, e.g. the cost per day of scrapers and labour, which could be £30/hr for a digger and £35 for a bulldozer, with labour at £15/hr.

7.4 Conclusion Construction and development can result in physical damage, contamination and soil sealing. During development, problems of soil damage are generally avoided by

Page 51 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development employing engineering solutions and the removal and storage of soil to be restored following construction. Failure to implement BMPs will result in increased costs to the developer, which acts as a deterrent. In terms of restoration, reductions in the soil function will occur when it is not carried out successfully and further costs will be seen in compensation paid to farmers due to crop loss. For other future land uses there is no deterrent and little cost to the developer (only society).

The impacts of soil deterioration and deviation from BMPs on soil as a base for construction appear to be generally small, mainly affecting site accessibility. Since the use of land for construction is a market activity the case for policy intervention to protect soils is weak. The only case for intervention would rest on any expected negative effects on the environmental services provided by soils rather than their value for construction.

There is significant intervention in the market for development land through the planning system. In principle, this trades off the economic benefits from development against its social and environmental costs. It is not clear that further policy intervention to protect soils is appropriate so long as purchasers are able to fully inform themselves about soil condition. However, were there to be a case for intervention, then a cost-based approach is the appropriate methodology for assessing changes in a soil’s ability to support construction. This reflects the fact that any effects would be accounted for in market transactions for land and the buyer’s WTP.

Page 52 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

8. Natural Attenuation of Pollution/Contamination and Contaminated Soil Remediation

8.1 Technical Review Soils in their natural state can attenuate a great variety of potential pollutants. However, to perform this function, soils need to be maintained in certain conditions that depend on: • The right proportions of air, water and solids (i.e. the right structure) • Adequate soil organic matter • A viable microbial biomass • A cation exchange capacity for buffering

Soils can attenuate pollution in a number of ways. These include: • Retaining pollutants in the soil through sorption and exchange processes • Degradation by chemical and biological processes • Immobilisation • Dilution

However, there are distinct differences in the type of pollution and the attenuation function that soils perform in rural and urban areas. On greenfield sites, pollutants are mainly limited to the common agricultural diffuse pollutants, such as nitrate, phosphorus and pesticides. Additions of heavy metals and organic compounds also occur through atmospheric deposition and the land spreading of fertilisers and wastes. Also, soils attenuate pollution from point sources, such as disused mine workings. Heavy metal losses to controlled waters can be high, but would be higher without the presence of soils. In urban areas, some sites are subject to severe contamination by heavy metals, hydrocarbons, sulphates and other pollutants, mainly from industrial activities. In many cases, soils and made ground is so severely contaminated that extraction and landfill is the only option. However, where soil materials are sufficiently friable and the concentration of heavy metals is not excessive, hydrocarbons can be degraded through the use of bioremediation, and the soil can be re-used on site. The second section in this chapter considers the ability of soil to attenuate industrial contaminants in brownfield situations. Consideration is given to the potential for using Monitored Natural Attenuation (MNA) on brownfield sites, and the extent to which this can be performed by the soil. The third section concentrates on the attenuation of diffuse water pollution from agriculture. Diffuse pollution has received increased attention in recent years with the continued implementation of the Water Framework Directive, and given that agriculture occupies 75% of the land in England Wales, we have devoted a large section to this subject. 8.2 Natural Attenuation of Industrial Pollutants The main driver for reducing soil contamination to acceptable levels is the contaminated land regime, which is set out in Part IIA of the Environmental Protection Act 1990. The Landfill Directive is also a driver in that it aims to reduce the amount of contaminated soil materials that are sent to landfill. A number of different bioremediation technologies, including heat treatment and composting with green

Page 53 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

waste, have been developed to help clean up contaminated sites and meet landfill reduction targets. Without the ability of soil microbes to degrade hydrocarbons to less toxic daughter compounds or metabolites, the option of using some contaminated materials on reclamation sites would not be available. Data on the quantity of materials that are bioremediated would provide some indication of the amount of topsoil resource that is preserved, as well as the cost savings, relative to landfilling and importing of topsoil, that this service provides. The burying of contaminated materials also has additional costs associated with the potential loss of pollutants to groundwater, where landfills are not secure. Bioremediation on site degrades the pollutant, thereby eliminating the possibility of environmental impact. It enables use of previously contaminated soil materials, either as topsoil/subsoil equivalent or as engineering fill material, and also saves costs by avoiding the need for landfill. Another possible strategy on contaminated sites is Monitored Natural Attenuation (MNA). This relies on the ability of soil to degrade and dilute a range of different hydrocarbons combined with monitoring to check that no contaminant plume is migrating to a sensitive receptor. As a result, expensive treatment or landfilling strategies can be avoided. MNA, however, is mainly associated with attenuation in drift and geological materials. Studies to date have concentrated on situations where the contaminant plume has migrated through the soil and has invaded the underlying drift and parent rock (Jones et al., 2001). It is the degradation processes occurring within the geological materials (e.g. chalk, sandstone, sand and gravel) that are monitored rather than in the soil itself. Contaminated soil materials to date have been dealt with through bioremediation, or removal from site. Nevertheless, natural attenuation performs a role in less contaminated soils, where concentrations of heavy metals and organic compounds are below the old ICRCL (United Kingdom Interdepartmental Committee for the Redevelopment of Contaminated Land) or the more recent Soil Guideline Value (SGV) levels. If soils did not have a capacity to absorb or degrade a certain concentration of these materials, the guideline limit levels would have to be set much lower on many sites. So, through its ability to attenuate industrial pollution to a certain degree, the soil performs a role in allowing the use of certain sites without the need for remediation, while on more heavily contaminated sites, the ability to bioremediate avoids the need for costly “dig and dump” solutions.

8.2.1 Climate and Biodegradation Climate change will also have an effect on the biodegradation, chemical degradation, sorption, immobilisation, dilution, and dispersion of contaminants. Reduced rainfall volumes will tend to concentrate soluble pollutants, while increased rainfall intensity may increase the transport of pollutants sorped onto soil particles. Reduced rainfall volumes, will also give rise to increased soil moisture deficits and possible changes in microbial activity, both in terms of the microbial species composition and the rate of activity, resulting in a possible reduction in biological degradation. Comparing biodegradation in temperate and Mediterranean climates could give some indication of how the biological dimension of natural attenuation may be affected by drier summers and wetter winters. 8.3 Natural Attenuation of Diffuse Pollution The soil has a role to play as part of certain diffuse pollution mitigation measures that range from land use change to the establishment of a cover crop. Pollution is mitigated through a measure’s effect on the source of pollution, but also the

Page 54 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development mobilisation and transportation of pollutants (Chadwick et al. 2006). Where marginal change in soil properties is concerned, the source dimension can change as a result of increased soil reserves of a potential pollutant, whether it be phosphorus, organic matter (nitrate), or heavy metals. However, the largest effect that marginal change in soil physico-chemical and biological properties will have on the pollution chain is on the mobilisation and transport dimensions.

A number of drivers can impact on the ability of a soil to perform the attenuation function. These include: • Land use change, e.g. from grassland or long-term set-aside to tillage land • Changes in pH • Saturation with heavy metals (and P) • Afforestation • Construction • Deviation from best management practices (BMPs) on agricultural land

The impact of activities linked to the last three drivers relates to weather patterns in any given year and to climate change as a whole. More damage and greater reduction in the ability of the soil to attenuate pollution will occur in wet years. For example, forestry and construction activities will result in soil disturbance. If these activities take place following or during rainfall events, when soils are wet, compaction will occur resulting in a reduced ability of the soil to attenuate pollution by P and sediment. BMPs are particularly weather dependent, and the soil’s ability to attenuate pollution will depend on the planning of land activities according to the timing of rainfall and soil wetness. For all drivers climate change is likely to narrow the window of opportunity for BMP to be followed at crucial times of the year (spring and autumn).

This section focuses on the ability of soils to attenuate the impact of heavy metals, pesticides, phosphorus (P), sediment, pathogens, nitrate (N) and acid deposition. Sediment is included, as it is an important pollutant associated with P pollution and relates to the physical, chemical and biological state of the soil and its ability to resist the erosive power of rainfall. There are a number of soil properties that change over the long and short term that influence soil erodibility. A healthy, free-draining soil, with stable, organic-rich aggregates will be far more resistant (even as a bare soil) to the erosive forces of rainfall than an organic-poor, weakly structured, or compacted soil. There has already been concern raised (Bellamy et al., 2005) about the reduced organic matter levels in arable soils. These lowered SOC levels can give rise to greater soil erodibility and hence increased sediment delivery to water.

Table 8.1 summarises the drivers that affect a soil’s attenuation function and the units and scale of marginal change for each of the pollutants. Sections 8.1.1 to 8.1.6 provide some background on the first five drivers.

Page 55 Table 8.1: Drivers of marginal change in a soil’s ability to attenuate pollution from heavy metals, phosphorus (P), sediment, pathogens, nitrate and acid deposition - and units and scale of marginal change.

Pollutants Drivers Heavy metals Phosphorus 2 Sediment FIOs 3 Nitrate Acid deposition

Deviation from BMP Metals leached or +0.3-2.2 kg Increase in loss +10-50% increase +1-45 kg N /ha/yr Increases if manures taken up in crop if TP 2/ha/yr. Across a (t/ha/yr) across a range of across a range of are left on the application rates too range of BMPs 1 Approximately 0.3- BMPs BMPs* surface or spread at high or pH declines. 2.2 t/ha/yr marginal times due to increased ammonia volatilisation. Extensive grassland Increased +2.2 kg total Increase in loss (up Reduction – size +15-20 kg N/ha from Reduction due to (with livestock) or opportunity for phosphorus to + 2.2 t/ha/yr) depends on land use extensive grassland decrease in number long-term set-aside spreading of organic (TP)/ha/yr of livestock (no livestock) to materials +35-40 kg N/ha from tillage land long-term set-aside Reduced pH Loss of marginal N/A N/A N/A Reduced N Reduces the ability of saving in cost of mineralisation, if the soil to buffer the landfill 4 or alternative extreme. effects of acid disposal route deposition (t/ha/year) Poor crop growth will reduce N uptake Saturation with heavy Loss of marginal P leaching N/A N/A Reduced N fixation N/A metals (and P) saving 4 and mineralisation. Construction Loss of marginal Loss potentially Loss potentially N/A Reduction in N Reduced natural saving 4 increased during increased during transformation buffering construction. construction. function Afforestation Reduced capacity to Analogous with Decrease in loss Reduction – size Reduced nitrate Reduced buffering fix heavy metals in arable to grass. So, - (t/ha/yr). Increase depends on previous leaching (kg/ha) – capacity in coniferous 2.2 kg TP/ha/yr. when clear felling or land use. size depends on coniferous plantations due to Increase during clear ground disturbance previous land use. plantations, due to declining pH. felling. declining pH. 1 The increase in losses of P and N in the table relate to the area of land where BMP is not carried out. 2 TP = Total Phosphorus. 3 Faecal indicator organisms. 4 this relates to the cost of landfilling biosolids where the option of spreading to land has been removed due to either construction or legislation combined with low soil pH or high soil heavy metal concentrations.

Page 56 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

8.3.1 Long-term Set-aside to Tillage Land (or grassland to tillage) This driver can move in both directions, although in the current economic climate it is more likely that tillage land will be converted to extensive grassland or long-term set aside. This change affects the ability of the soil to attenuate pollution in that soils that are bare for part of the year under tillage farming will have a permanent vegetative cover. This constant vegetative cover performs a number of roles. It: • Shields the soil surface from raindrop impact • Increases the stability of surface soil aggregates and resistance to the shear forces of run-off • Restricts compaction from cultivation equipment • Increases evapo-transpiration rates • Increases uptake of nutrients (particularly nitrate) in the autumn (compared with no crop cover or a late-drilled cereal crop).

A change from arable to extensive grassland will reduce nitrate leaching. The lack of annual soil cultivation reduces mineralisation rates, and along with the accumulation of N in soil organic matter reserves reduces the pool of soil nitrate that is available for leaching in the autumn. Also, the permanent vegetation cover results in higher evapo- transpiration rates through the autumn, such that drainage (and associated nitrate leaching) is delayed relative to the arable situation. As a result, starting from a baseline nitrate leaching situation of 40-50kg N/ha/yr., converting from arable to extensive grassland will typically reduce leaching by 15-20kg N/ha. If the change is from arable to permanent set-aside without livestock, the reduction in N leaching is nearer 35-40kg N/ha (i.e. > 90%). However, this includes a change in the quantity of the N source, as well as an increase in the ability of a soil to attenuate N leaching.

Arable reversion will also have an effect on the ability of the soil to attenuate pollution by P, sediment, FIOs and pesticides. This is due to the fact that losses of these pollutants are commonly associated with soil erosion and run-off. Grass affords protection from the forces of raindrop impact and surface run-off, and increases soil aggregate stability (Tisdall and Oades 1982). In addition under permanent set-aside or extensive grassland heavy machinery is largely absent, thereby reducing the degree of soil compaction. Furthermore, the increased evapo-transpiration rates relative to the arable context maintain the soil in a drier condition over the autumn period. The resulting delay in the return to field capacity combined with increased infiltration rates under grass (Holtan and Kirkpatrick 1950) result in fewer run-off and soil erosion events. The soil’s ability to attenuate pollution by sediment-associated pollutants is therefore increased.

The one exception is pollution by FIOs. If livestock were not present before the conversion, arable reversion to extensive grassland is likely to increase FIO pollution by around 10-30%. However, this does not relate to the soil’s ability to attenuate FIO pollution, but to the driver itself and the presence of FIOs in livestock excreta/manures.

8.3.2 Reductions in pH There is evidence that the frequency of lime applications has been reduced in recent years (Goodlass and Welch, 2005). Although Gardner and Garner (1957) stated that liming can reduce the tendency of light-textured soils to cap under heavy rainfall, small changes in pH are unlikely to have a great impact on pollution attenuation. However, a significant fall in pH to 5.5 or below may impact on a soil’s ability to

Page 57 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

attenuate heavy metal pollution and acid deposition. Yield effects due to reduced pH will also reduce the nutrient sink into the growing crop and increase losses, particularly of nitrate. If N fertiliser inputs are not adjusted to account for reduced yield, N losses could amount to several tens of kilograms per hectare.

The mobility of heavy metals and leaching risk increases with soil acidity. If soil pH falls to 5.5 or below, the ability to attenuate heavy metal pollution is impacted in two ways: • The leaching of heavy metals from the soil itself may increase • The ability to use the land for the “recycling” of biosolids is lost at pH 5 and below

The marginal loss of soil value relates to the loss of the soil as a location to recycle biosolids. If this soil function is lost, there will be direct cost implications that relate to the increased cost of alternative disposal routes. These include landfill and incineration. Soils with heavy metal concentrations that exceed the Sludge (Use) in Agriculture limit level (which are pH banded for Zn, Cu and Ni) may also have a reduced long-term economic land value.

Heavy metal losses in leachates are of concern to the Groundwater and Water Framework Directives. There may also be “public good” and “willingness to pay” issues associated with avoiding leaching of heavy metals to controlled waters.

Reductions in pH also reduce the ability of the soil to buffer acid deposition. Acid deposition, more commonly known as acid rain, results from man-made emissions of sulphur dioxide and nitrogen oxides through the burning of fossil fuels for energy and transportation. Low pH soils have a low base saturation percentage on the mineral and organic exchange complex. This low proportion of bases on the exchange complex and in solution reduces the capacity of the soil to counter the effects of acid deposition. The ability of a soil to attenuate acid deposition is defined by the “critical loads” concept. The “critical load” is the maximum amount of pollutant that a soil can tolerate without damage being caused to the related ecosystem. For example, the UK critical load value for lowland dry heaths is only 12 kg ha -1 yr -1 (http://critloads.ceh.ac.uk/). Soils with low pH values have lower “critical loads”, and the risk of further acidification is increased. Therefore, the following drivers that may result in lower soil pH can reduce soil critical loads and increase soil sensitivity to acid deposition: • Arable reversion to extensive grassland, without maintenance of lime applications • The conversion of improved or semi-improved grassland to extensive grassland (particularly on low base status soils), without maintenance of lime applications

The loss in value is most easily defined by reduction in public good and the willingness to pay for sustaining sensitive ecosystems. Soil acidification can also result in the release of aluminium into surface and groundwater systems, which has potential human health implications.

8.3.3 Heavy Metals and Organic Compounds Maximum allowable limits for a range of heavy metals, where biosolids are applied to land, are defined in the Sludge (Use) in Agriculture Regulations. These limits relate to the maximum acceptable concentration for the most sensitive receptor. For example, for zinc and copper, crop yields and biomass production were considered the most sensitive receptors at the time the Regulations were put in place. However, more recently, concerns over the effects of metal inputs on soil rhizobial populations have increased, with soil zinc limits reduced on a precautionary basis in the DoE Code of

Page 58 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Practice for Sludge Use in Agriculture. Humans were considered the most sensitive receptors for cadmium. The ability of the soil to act as a “reserve” for heavy metals from biosolids is maintained for as long as the Sludge Regulation limits are not exceeded. However, once the limits are exceeded it is no longer possible to use the soil’s biosolids recycling function due to the constraints on spreading imposed by legislation. In fact, the function that soil performs in avoiding transfer of pollutants to sensitive organisms and in preventing leaching of heavy metals is preserved beyond the limit level. However, permission to make use of this function is withdrawn as a result of the legislation If this soil function can no longer be used, there will be direct cost implications that relate to the relative increased cost of alternative disposal routes. Soils with heavy metal concentrations that exceed the Regulations limit level may also have a reduced economic land value.

Although the ability of the soil to accept heavy metals from sewage sludge is lost at the Sludge (Use) in Agriculture Regulation limit levels, the soil itself does not lose the ability to attenuate pollution by heavy metals from other sources. These other sources include fertilisers, manure, paper waste and green waste compost.

The ability of the soil to attenuate pollution by heavy metals is affected by soil organic matter content and pH. Heavy metals tend to be absorbed onto soil organic matter, provided the soil pH does not fall below 5.5-6.0. The mobility of many heavy metals and leaching risk also increases with soil acidity. The toxicity of heavy metals is largely dependent on the degree to which the metal is present as a free ion (the metal not bound to any other chemical species in the solution). Organic matter content and soil pH are used to estimate the free metal ion concentration in the soil, and hence the potential toxicity of metals to soil and aquatic organisms. Reduced soil pH and reduced soil organic matter content will therefore increase the leaching and toxicity of metals in soils and water.

It is interesting to note that many Sites of Special Scientific Interest have soils with high metal concentrations. The increased environmental stress created by high metal concentrations can enable rarer and more highly valued plants to out-compete more vigorous colonising species.

Heavy metals also impact on soil microbial communities that are responsible for nutrient cycling and the biodegradation of contaminants. Soils that have severe heavy metal contamination are therefore prone to transfer pollutants to controlled waters and cannot perform the nutrient cycling function of unaffected natural soils. Such soils may be found on flood plains downstream of disused mine workings, or on dedicated sites used for sewage sludge disposal. In addition to heavy metals, the spreading of sewage sludge to land adds potentially toxic organic compounds to the soil. These compounds, which are also added through atmospheric deposition, include organic halogen compounds, surfactants, detergents, polychlorinated biphenyls (PCBs), dioxins and furans. Although these compounds are potentially extremely toxic, the soil provides a barrier between these organic pollutants and plants grown for human and animal consumption. Most are insoluble and hydrophobic which results in low bioavailability and very little uptake by plants (Erhardt and Prüß, 2001). Organic pollutants, such as dioxins and PCBs are sometimes found in agricultural crops, but this is more the result of atmospheric deposition than absorption from soils. Higher levels of these compounds could be found in agricultural produce and could effect yields were it not for the ability of soil to degrade, dilute and provide a barrier to uptake. Some toxic detergents found in sewage sludge are mobile and bioavailable in soils, but transfer from soil to human consumers is still limited due to dilution and sorption

Page 59 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

processes. The main threat from toxic detergents is to aquatic ecosystems where changes to fish populations have occurred (Erhardt and Prüß, 2001). However, without the attenuation function of soils, the impact on aquatic ecosystems would be much higher.

8.3.4 Construction Construction for the purposes of this section means the disturbance of soil to create sealed surfaces, buildings or infrastructure, such as roads, tracks or pipelines. Construction and development can have several impacts on soil, from physical damage due to poor soil handling practices to contamination that reduces the ability of soil to perform primary functions. Construction and development also result in ‘soil sealing’. In other words, the covering of soil with impermeable materials such as buildings, or the compaction of the soil surface thereby reducing permeability. Construction can have an impact on a number of pollutants as outlined in the following paragraphs: • Heavy metals – the organic ‘waste’ recycling function is lost if soils are sealed through construction. Post-construction, the attenuation function of soil mineral and organic matter is lost, and there can be an increase in the transfer of urban metals to water. • Phosphorus and sediment – The ability of the soil to absorb rainfall will be reduced during the construction phase, potentially resulting in increased P and sediment loads. Post-construction, the source for P and sediment is lost. • Pathogens – Construction removes the source of the pollution. • Nitrate – Post-construction, the soil N cycling functions is reduced. In sealed soils, microbial activity slows down significantly. There is a reduction in N fixation and in N mineralisation, due to reduced soil aeration. • Pesticides – Construction reduces the soil attenuation function. Although practices are improving, high rates of pesticides can be used in urban areas, in gardens, parks and on “sealed surfaces”. The reduced infiltration capacity of most urban surfaces potentially increases the transfer of pesticide pollutants to water. The introduction of SUDS (Sustainable Urban Drainage Systems) reduces the impact of construction, but does not eliminate the pollution issue. • Acid deposition – The influence of construction on the ability of the soil to buffer acid rain is not clear. Förster (1998) found that the concentration of calcium in roof run-off was actually increased relative to rainwater. Calcium (base rich) compounds, leached from roof materials, and from other construction materials, may serve to buffer the effects of acid deposition in some circumstances. However, the overall effect on urban run-off may be neutral and heavy metal loads are more likely to be adverse to local ecosystems. The function of the soil itself to buffer acid deposition is reduced or even lost post-construction.

8.3.5 Afforestation • Heavy metals and acid deposition – On non-calcareous soils, the buffering capacity of the soil and its ability to fix heavy metals is reduced after afforestation because of a decline in pH. This generally occurs under afforestation possibly due to the combined effects of reduced dilution of N and S deposition (related to high evapotranspiration rates), the lack of liming, increased uptake of base cations, and lower decomposition rates under forestry (Ferrier et al. 1990). Catchments with low base status substrates are particularly susceptible.

Page 60 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

• P and sediment – During clear felling operations in productive forestry, P and sediment loads can be increased due to compaction and the associated reduced soil pollution attenuation function. However, while the land is forested the effect is analogous to arable reversion. One would therefore expect P loads to be reduced due to increased vegetative cover, litter cover and increased soil aggregate stability. • Pathogens – are largely eliminated unless livestock are introduced to the forest. • Nitrate – The ability of the soil to attenuate N pollution is increased by afforestation, due to the low level of N inputs and N cycling within the soil- vegetation system. The change is analogous to arable reversion to permanent set-aside. With a baseline loss of 40-50 kg N/ha/yr., a reduction in nitrate leaching of > 90% would be expected.

In Table 8.1, the assessed effect of afforestation on the ability of a soil to attenuate P pollution assumed that the land use change was from arable agriculture to forestry. The effect on the ability to attenuate transfer of FIOs assumed that the land use change was from extensive grassland to forestry. For all other pollutants, the given effect was qualitative, with the scale of effect dependent on the baseline context. In the case of heavy metals and acid deposition, the changes largely related to soil pH. In the case of nitrate, the changes related to the absence of cultivation, and increases in soil organic matter reserves following arable reversion.

8.3.6 Deviation from Best Management Practice (BMP) BMP assumes that the soil is maintained in good agricultural and environmental condition, with sufficient organic matter to maintain soil aggregate stability and that any compaction is alleviated in good time. The values given in Table 8.1 apply to the range of change values across the full range of mineral soil types for specific farm types and practices. For example, in the case of N, the range 10-45 kg NO 3-N/ha represents the ability of all soils to attenuate N pollution when a cover crop is in place on a specific arable farm defined within Defra project ES0203 (The cost-effectiveness of integrated diffuse pollution mitigation methods). The top of the range (45 kg NO 3- N/ha) represents the loss from sandy soils, while the bottom of the range (10 kg NO 3- N/ha) represents the loss from clay soils.

The baseline context for BMP relates not only to farming practices required for cross- compliance under the Single Farm Payment scheme, including the requirement to maintain the land in “good agricultural and environmental condition”, but also to the following management practices: • Establish cover crops in the autumn • On light soils, cultivate land for spring crops in spring rather than autumn • Cultivate compacted tillage soils • Restrict traffic to tramlines over winter • Loosen compacted soil layers in grassland fields • Maintain or enhance soil organic matter levels • Reduce field stocking rates when soils are wet • Move feed and water troughs at regular intervals

Table 8.1 provides figures for the marginal change in pollution by phosphorus, nitrate and pathogens (resulting from deviations from BMPs and land use change). These

Page 61 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

figures are taken from previous Defra studies (NT2511 – Cost curve of nitrate mitigation options ; PE0203 – Cost curve assessment of phosphorus mitigation options relevant to UK agriculture and ES0121 – COST-DP: cost effective diffuse pollution management). The BMPs include establishing cover crops, reducing compaction in tillage soils, and restricting livestock access to soils in marginal places and at marginal times.

The figures for nitrate loss refer to annual losses from the soil profile to either surface or groundwaters. Nitrate is soluble and moves through the soil profile via the soil matrix in free draining soils or via cracks in heavy drained clay soils. Therefore, land management practices that have the following effects will increase the risk of nitrate leaching: • Increase the downward flow of water (e.g. sub-soiling) • Increase the available pool of nitrate (e.g. cultivation, application of manures, or N fertiliser in excess of crop demand) • Reduce the uptake of nutrients (e.g. no cover crop, crop damage or soil structural damage)

The figures for P and sediment refer to annual losses from the soil surface and from the soil profile (as dissolved P and colloids) to water systems. On free draining soils, the loss of P is largely associated with soil erosion (loss of sediment across the soil surface) and on drained soils, the translocation of sediment largely occurs via drain flow and large continuous pores (by-pass flow). On soils with high reserves of phosphorus (P index 4 and above) dissolved P can also be lost in drainage water flows. The total amount of P and sediment lost each year is related to the amount continually lost in drainage water flows, and the amount lost from discrete erosion events and incidental events associated with manure or fertiliser spreading.

Increased amounts of P loss are therefore associated with: • Soil structural damage resulting in surface run-off (associated with poor cultivation practices, compaction from machinery, livestock poaching etc.) • Surface sealing (the formation of surface crusts due to raindrop impact, particularly on sandy and silty soils) • Low organic matter content (giving rise to soil aggregate instability and the formation of crusts) • Bare soils • Poorly timed manure and fertiliser applications (incidental losses)

8.3.7 Attenuation of Pollution from Pesticides Soils have a value in attenuating pollution from pesticides. However, the ability of a soil to perform this function depends on a number of factors that relate to the properties of the soil and pesticide themselves. These factors include: • Pesticide solubility in water • Pesticide speed of degradation in soil • The degree to which a pesticide binds to soil • Soil organic matter content • Soil clay content

Page 62 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Whether a pesticide constitutes a risk to controlled waters will depend on its behaviour in soil. For example, Clopyralid does not constitute a risk to groundwater, because it degrades moderately quickly in soil through microbial action. Soils are therefore least effective at reducing pollution from soluble pesticides that do not bind strongly to soil and degrade slowly. These pesticides include Isoproturon, Dicamba and Dichlorprop. The converse is that soils are most effective at preventing pollution from pesticides that bind strongly to soil. These include Diquat, Glufosinate-ammonium, and Paraquat.

The strength of binding and the amount of pesticide bound relate to soil organic matter and clay content. For example, Glufosinate-ammonium binds more strongly to clay soils and those with higher organic carbon content. However, for pesticides that degrade rapidly, the speed of degradation does not necessarily increase directly with soil organic matter content. One example is MCPA (4-chloro-2-methyl phenoxyacetic acid), the degradation of which is more rapid in soils with lower organic matter contents (<10%). It is therefore not possible to relate the soil’s ability to attenuate pesticide pollution to organic matter content in all cases.

Pesticide loss is often associated with heavy or prolonged rainfall after application. Pesticide pollution is therefore influenced by the same soil properties and rainfall events that affect P and sediment loss. Soluble pesticides are lost in run-off and drainage water flows, while those that bind strongly to soil particles are lost with sediment. Therefore, land management practices that degrade soil structure and reduce water infiltration rates will increase the risk of pesticide pollution. Deviation from BMPs is probably the most important driver in reducing the soil’s ability to attenuate pesticide pollution. This is particularly the case if the deviation involves serious soil compaction through: • Use of high ground pressure tyres • Not keeping to tramlines or traffic route ways for support machinery • Working on or cultivating wet soils • Over-cultivation of soils

To conclude, the soil properties that could be used to assess marginal change in the ability to attenuate pesticide pollution will vary according to the pesticide in question. For those pesticides that are associated with lost sediment, BMPs that reduce soil run-off and erosion will reduce losses. Uncompacted soils that have an open structure and give high infiltration rates will have a greater value in reducing pesticide pollution than crusted, capped or compacted soils. Soils with a vegetative cover will also have a natural tendency to attenuate sediment and run-off associated pollution more effectively than bare soils.

8.4 Economic Analysis Only in a limited number of cases has it been possible to quantify the physical impact of drivers on the ability of the soil to attenuate pollution. This severely limits the extent to which values can be ascribed to changes in this function of soils.

Impacts on phosphorus, nitrate and pathogen pollution are calculated above (Table 8.1) for three drivers (extensive grassland, afforestation and deviation from BMP). It is not possible to estimate the marginal valuation of changes in a soil’s ability to attenuate pathogens. This would require detailed analysis of specific pathogens and the effects of changes in pathogen levels in soil on animal and human health. Such

Page 63 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development an analysis was beyond the scope of the study. The economic analysis is therefore restricted to the valuation of changes in phosphorus and nitrate losses.

The benefit studies relating to water quality are quite limited in scope and do not provide a basis for valuing marginal changes in N, P or faecal indicators in water bodies. Hanley (1989) surveyed householders in East Anglia and used the contingent valuation (CV) method to estimate their WTP for water supplies containing a maximum of 50 mg/l nitrate. The mean bid was £12.97 per household (aggregated to £10.8m per year over the East Anglia population). This exceeded the actual cost of water treatment to reduce nitrate levels to 50 mg/l.

Green and Tunstall (1991) also used CV to estimate the benefits to the public from river water quality improvements. The study focused on recreational benefits – the enjoyment derived by recreational users from improved water quality. Across 12 sites the mean WTP was £1,203 per year. A number of studies value the benefits to anglers from water quality improvement (e.g. Davis and O’Neill, 1992, Spurgeon et al ., 2001). Georgiou et al . (2000) determined the WTP for specified levels of quality improvement in the river Tame and Eftec (2004) aggregate the WTP calculated by Georgiou et al . for a medium improvement in water quality using the figure of £5,560 per km per year. However, in none of the above studies is it possible to relate marginal changes in N, P or faecal indicators to changes in public benefits or disbenefits.

There is some evidence that the public are willing to pay to reduce eutrophication of inland waters. Bateman et al . (2006) estimate the willingness to pay (WTP) for the application of measures to prevent eutrophication in East Anglian rivers and lakes. The WTP per household was £38.5-£75.4 per year depending on the assumptions made about non-respondents. Aggregated to the region the benefits totalled £86- 170m per year which were substantially higher than the current annual cost of measures to reduce eutrophication in England estimated to be £54.8m per year by (Pretty et al ., 2001). Aggregating the lower bound estimate to the 21.73m households in England gives a WTP of £836m per year.

Pretty et al . (2000) and Eftec (2004) use the cost of drinking water treatment as one element in the cost to the public of agricultural pollution of surface and groundwater (Table 8.2).

Table 8.2: Estimates of the financial costs of water pollution (£m per year)

Eftec (2004) (UK) Pretty et al . (2000) (UK) % of total cost % of total cost attributable to Cost (£, m) attributable to Cost (£, m) agriculture agriculture

Nitrate removal by 70 15.6 80 16 water companies

Phosphate and 43 34.5 43 55 particle removal by water companies

Zoonoses removal by N/e 90 23 water companies

Page 64 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

However, these are total costs and do not identify the marginal cost associated with any change in the concentration of N, P or zoonoses. The major element in cost is that for phosphate and sediment removal.

An alternative approach to costing water pollution is to calculate the costs of controlling or reducing agricultural pollution at source. This is relevant to policy measures that may be developed in response to the requirements of the Water Framework Directive. The cost of a change in soil use or management that increases the output of one or more pollutants (e.g. management that diverged from BMP) can be calculated from the cost of additional compensating measures that would have to be introduced to avoid additional pollution.

For example, Cuttle et al . (2006), in their draft report, list 44 measures that may be applied on farms to reduce the emissions of nitrate, phosphate and faecal indicator organisms (FIOs). For nitrate alone, the cost to arable farmers of reducing losses is less than £1 per kg N per ha for small reductions. However, as would be expected, the marginal cost of reducing nitrate losses increases as the required loss is increased. To make reduction above 30 kg per ha the cost increases to £10 per kg N per ha (Table 8.3). Where only a reduction in P loss is required the cost is around £3 per kg P per ha. This increases substantially when major reductions in P loss are required. Where reductions in both N and P loss are required moderate levels of P reduction can be obtained at minimal additional cost. It is only where substantial levels of P emission reduction are needed that the costs increase dramatically.

Table 8.3: Cost of least cost measures to reduce N, P losses on arable farms (£ per ha per year) P loss reduction (kg per ha per year)

P loss reduction (kg per ha per year) N loss reduction (kg/ha/yr.) 1 2 5

0 3.2 6.5 98.0

10 6.2 9.45 98.0

20 14.1 17.2 100.8

30 54.7 57.5 111.3

40 157.9 160.1 197.5

In practice the costs of reducing agricultural impacts on water quality will vary with the target levels of pollution reduction that are set by government. These are likely to vary with the catchment in relation to the pressures from N and P pollution. Costs will also be higher than those in Table 8.3 because it may not be feasible to implement all of the measures investigated by Cuttle et al . (2006) within a policy framework due to difficulties in determining the level of compliance. In addition, there will be transaction costs associated with implementation.

8.4.1 Marginal Value of Changes in Pollution Attenuation The cost or benefit to society from a change in the soil’s attenuation of N or P will depend heavily on soil location. The recreation values of water resources are unaffected by changes to N and P. In terms of water company costs of water purification, there are no estimates available of the marginal changes in costs associated with any change in N or P concentration but these are expected to be quite small. Specifically, it will depend on what action is required in relation to N and P in

Page 65 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

the context of the Water Framework Directive and whether there are impacts on drinking water quality. Applying the compensating costs calculated by the method used to derive the data in Table 8.3 it is possible to estimate the cost on arable land of compensating measures to address a loss of attenuation capacity (Table 8.4).

Table 8.4: Marginal cost of changes in attenuation capacity for N and P

Marginal cost of Range assessed compensating Driver Phosphorus Nitrate (kg per ha per measures year) (£/ha/yr.) Deviation from +0.3-2.2 kg +1-45 kg +0.3 kg TP/+1 1.2-253 BMP TP/ha/yr. N/ha/yr across kg N to +2.2 kg Across a range a range of TP/ 45 kg N of BMPs* BMPs* Extensive grass +2.2 kg +15-20 kg N/ha +2.2 kg TP/ +15 11.9-17.9 (with livestock) TP/ha/yr. from extensive kg N to +2.2 kg or long-term set grassland TP/ +20 kg N aside (no +35-40 kg N/ha +2.2 kg TP/ +35 livestock) to from long-term kg N to +2.2 kg tillage 79.1-159 set aside TP/ +40 kg N Afforestation -2.2 kg Not estimated -2.2 kg TP -7.2 TP/ha/yr. Increase if clear felling.

The data in Table 8.4 emphasise the wide range in marginal costs that can occur in relation to deviation from BMP. Extreme mismanagement of soils with both impacts on N and P retention can have costs estimated at up to £253 per ha per year. However, in most cases they would be substantially lower. Conversion of extensive grassland or long-term set-aside to tillage is associated with marginal costs of up to £17.9 and £159 per ha per year respectively. Afforestation leads to a small increase in P attenuation capacity valued at around £7.2 per ha per year. These values would only apply where any change in attenuation had implications for meeting the requirements of the Water Framework Directive. In other situations the marginal value of the loss of function is likely to be very small or zero unless (i) drinking water is abstracted and additional treatment costs are incurred, or (ii) there are impacts on eutrophication of inland or coastal waters.

8.5 Summary and Conclusions 8.5.1 Technical It has not always been possible to quantify the impact of the various drivers mentioned above on the ability of the soil to attenuate pollution. Even where values are provided, they only relate to the baseline situation described in Defra project ES0203 (The cost-effectiveness of integrated diffuse pollution mitigation measures) for a number of defined model farms. For example, changing land use from arable to long-term set aside is likely to reduce N pollution by 35-40 kg N/ha/yr., P pollution by 2.2 kg TP/ha/yr. and FIO pollution by 10%. Values for the reduction in sediment loss can be derived from models such as PSYCHIC, while values for pesticide losses will be dealt with in Defra project ES0205. The ability of the soil to attenuate pesticide pollution will vary according to the behaviour of each particular pesticide in soil and

Page 66 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development water. For pesticides that bind tightly to soil, losses will be related to losses of P and sediment.

Changes in pH are likely to impact on the soil’s ability to attenuate heavy metal pollution and acid deposition. However, as noted by Smith (1996) heavy metals are usually only transferred to groundwater following large applications of sewage sludge that would only be permitted on restoration projects today. Even then, heavy metal concentrations have not been known to exceed the maximum guideline values set by the World Health Organisation (WHO) for drinking water (Grant and Olesen 1984).

In agricultural soils, additions of heavy metals to Sludge (Use) in Agriculture Regulation limit levels affects the soil’s ability to attenuate heavy metal pollution from sewage sludge, as the potential to accept biosolids would be lost and an alternative (perhaps more expensive) outlet would have to be found. The legal status of sludge incineration is currently under review and this may have implications for the future. However, increase of soil heavy metal concentrations to the 1989 Regulation limits is very unlikely to impact on losses to surface or ground waters (Smith, 1996). The soil would not lose its ability to attenuate heavy metal pollution for the majority of receptors, but the most sensitive receptor could be impacted. For example, the ability of White Clover to fix nitrogen from the atmosphere could be affected by soil zinc concentrations at the 1989 Regulation limit level.

The impacts of afforestation are analogous to a change from arable to permanent set aside, although as mentioned in section 4, there would be changes to the soil water budget (due to interception and reduction in throughfall) that could reduce soil drainage volumes. This could have an impact on return to field capacity (wetting up of soils and the start of drainage) and concentrations of soluble pollutants (e.g. nitrate) in leaching soil water. However, any change in drainage volumes due to forest interception could be offset by climate change (increased autumn rainfall). Under afforestation the impacts of acid deposition and heavy metals are likely to increase (reduced buffering capacity with declining pH on non-calcareous soil). P and sediment pollution may increase during clear fell operations. Such an increase would be analogous to a deviation from BMP in the agricultural context and severe compaction of soils. Pathogens are largely eliminated under forestry, but where livestock are present, the soil’s ability to attenuate FIOs is increased under established forest due to vegetative cover and lower run-off volumes. The reverse may be true during the process of afforestation if open drains are used for nursery stock.

Construction will increase run-off volumes due to surface sealing. However, in most cases the source of the pollution is removed and urban pollutants replace rural pollutants. The soil’s ability to attenuate urban pollutants is lost, except where sustainable drainage systems (SUDS) are used, and the pollutant problem is exported to another location. During the construction phase P and sediment pollution (and loss of other sediment-associated pollutants) is likely to increase due to disturbance and compaction (analogous to deviation from BMP).

Deviation from BMP will reduce the ability of the soil to attenuate the majority of pollutants. Relative to the baseline situation given in Defra project ES0203, P pollution may increase by 0.3-2.2 kg TP/ha/yr., N pollution by 1-45 kg N/ha/yr. and FIOs by 10- 50% (all ranges depending on the BMP in question). Sediment and pesticide losses are likely to increase in proportion to P losses, although soluble pesticides may also increase with N losses.

Page 67 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

8.5.2 Valuation The valuation of changes in a soil’s natural attenuation of pollution/contamination will depend on which pollutant is affected. Changes in leaching of heavy metals, N and P and movement of pathogens may have implications for water quality in associated catchments. This will be the case where there are existing risks in relation to the WFD, where drinking water abstraction takes place or where eutrophication is an environmental issue. We were only able to quantify the possible marginal costs of a loss of attenuation capacity for N and P using policy costs under the WFD. These varied substantially depending on the extent of the change in attenuation capacity but could be high where BMP was not adhered to.

Changes could also occur in the ability of soils to buffer acid deposition but it was not possible to quantify these effects, although remedial action through liming is a straightforward measure at least on accessible land.

We can therefore conclude that the direct estimates of benefit values are limited in scope and can only provide contextual information on the public’s WTP for the attenuation of pollution. The benefit data are not sufficiently specific to use a basis for policy development in relation to protecting soil services. A case can be made for more valuation studies in order to quantify the benefits or disbenefits to society from changes in the capability of soils to attenuate pollutants. The most direct would be to estimate the public’s WTP for measures that reduced pollution through the enhancement of soil capability. Alternatively the value of protecting an existing capability could be estimated. CV and/or choice experimental methods would be appropriate.

Cost-based approaches can be used where changes in soil management have implications for costs either to the private sector (e.g. water treatment costs) or to other public sector policy measures (e.g. costs of meeting the requirement of the WFD). In some cases the market will provide values which reflect remediation costs or losses in the output of contaminated soils. It should also be possible to estimate costs for changes in the attenuation capacity of soils in a WFD context when the research of Cuttle et al. (2006) is completed. Additional research on water company cost saving from attenuation measures would be needed as would research on the indirect impacts of such measures on the levels other pollutants.

Page 68 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

9. Archaeological and Landscape Heritage Protection

9.1 Technical Review Features that can be considered part of our cultural heritage that interact with the soil are of two kinds, namely those that are true archaeological /artefact features as suggested by the title of this section, and those which are anthropogenic landscape features. The former covers buried artefacts and remains of human activity and habitation as well as ancient monuments, whereas the latter covers features where the land form itself has been transformed and carries evidence of past human activity. Examples of the latter may be old “rig and furrow” field systems, evidence of old field boundaries and changed habitats as well as features such as earthworks which may also be described as ancient monuments.

Attempts to classify soils as to their features that may assist the location of archaeological remains have so far proved fruitless, as past human activity has occurred almost anywhere in the British Isles. What clues there may be to location of its evidence are geographic and topographic rather than soil features. With regard to the preservation of evidence of archaeological features it may be considered that soil features would be more influential in being either more or less beneficial. This is partly true in that certain soil features are indicative of archaeological features; such as micromorphological features that indicate historic soil disturbance (Courty et al ., 1989), buried pollen evidence, “dark earth” stratigraphy (Macphail & Courty, 1985), and some forms of phosphate analyses. However, these are derived features from the presence of the artefact, rather than inherent features that help preservation. Features that do assist preservation are those associated with extreme edaphic environments, such as the anoxia of peat deposits, or the oxidising conditions of extremely dry sites. Both of these extremes, assisted greatly by a cold climate, would help in preserving organic materials in that they are unfavourable to microbial decay. In addition the action of certain organic acids from peat deposits will also attenuate microbial activity. In other situations however, the soil chemistry may such that the corrosion of certain metal or ceramic artefacts is reduced. The most obvious preserving feature of soil however, is simply the protection from disturbance that burial affords.

9.2 Valuation of Archaeological Site/Artefact Protection Eftec (2005) undertook a review of the valuation of the historic environment. They found only nine relevant valuation studies in the UK and these largely related to canals, historic buildings such as cathedrals and impacts of road improvements on Stonehenge. Additional case studies were made but these covered much the same ground – a townscape, a park restoration, a cathedral and a castle. They concluded that it is difficult to generalise from a small number of studies but that generally positive values are attributed to the conservation or restoration of the heritage assets, implying that degradation is a disbenefit to the aggregate public. The public exhibits a WTP for restoration, and use values are higher than non-use value. This contrasts with the situation that pertains with biodiversity (see section 10). WTP was higher in higher income groups and amongst those making more trips to see the historic environment. In most UK studies there was a substantial proportion of people (30- 50%) who were not willing to pay anything for access or restoration of historic sites. Whilst some may be protest votes the implication is that interest in archaeology, as reflected in a positive WTP, is far from universal in the population.

Little of the Eftec (2005) review is relevant to the protection or restoration of land of archaeological interest without associated buildings. Since it is not usually possible to

Page 69 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development visit agricultural sites only non-use value is relevant and one may infer that that the WTP for the restoration or protection of agricultural sites will be lower than sites reviewed by Entec. The fact that the planning system provides for protection from development suggests that sites of archaeological interest are valued by the public. However, agriculture and forestry are largely outside the planning system.

We are aware of only one study that sought to estimate the WTP for policy measures to protect the archaeological heritage on farmland. Hanley et al . (1996) estimated the value at the margin from protecting the archaeological interest in the Breadalbane ESA at £10.00 per household per year. This provides evidence that the protection of archaeology is important to the local public although the estimates of environmental value derived in the study were high, and possibly unrealistic. Not much can be inferred about the archaeological heritage in SDAs from the recent Eftec (2006) study (see above). Whilst there was a WTP to maintain traditional farm practices, there was almost no WTP to improve field boundaries (including traditional stone walls). If anything, one might infer that the WTP to maintain archaeological features in the landscape is likely to be small.

Macmillan (2003) has reviewed the value of forests to protect archaeology. He states that there is considerable evidence to suggest the interest in, and the value placed on, archaeology is growing amongst the public. During the period of rapid expansion from 1930 to 1980, forestry was considered a threat to the archaeological heritage (Barclay, 1992). Although guidelines are not in place from either private or public forestry, concern was still voiced in a study of the management of English native woodlands that the archaeological heritage was inadequately protected (Crabtree et al ., 2002). However, there are very few valuation studies that relate to archaeology and he finds none that relate to archaeology in the UK other than that made as part of the evaluation of the ESA programme (see above). He attempts to transfer the ESA value estimates to the UK forest estate. On this extremely fragile basis he estimates the benefits of management in forests to protect archaeology as lying between £0 and £247 per ha.

We can conclude that there is a virtually complete lack of evidence on the value of the benefits or disbenefits from soil management impacts (drainage, compaction and decontamination) on the archaeological heritage. It is likely that the benefits will mainly be non-use benefits, since most sites potentially subject to damage will be on private land not accessible to the public. Drawing a parallel with the analysis of benefits from national parks (Bateman et al ., 1994) we can infer that where the archaeology has many substitutes across the country it is best treated as a local public good whose value will depend on its characteristics and the number of local occurrences. Benefits of soil protection will be greatest for archaeological interest that is rare in a national context, i.e. where this is of national interest with few or no substitutes. However, without economic valuation it is not possible to place a monetary value on the benefits from soil protection. Stated preference techniques would be appropriate for providing such valuations.

Despite the virtual absence of benefit estimates for archaeological protection, it is possible to use cost data to infer a lower bound estimate of the public’s WTP for the preservation of archaeology. Where land management payments are made to protect the archaeological interest in landscapes the cost to society provides such an estimate. This occurs, for example, under the Environmental Stewardship Scheme (Rural Development Service, 2005), where payments are made to take archaeological sites out of arable production.

Page 70 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

9.3 Factors that Affect Archaeological Site/Artefact Protection Drivers of change impacting on this soil function are associated with site disturbance or radical changes in the drainage patterns and soil chemistry at a site. Loveland & Thompson (2002) list the following three simple indicators of the potential for damage to archaeological sites (Table 12.1, page 158 in Loveland & Thompson, 2002), as well other more general indicators:

1. Change to the area of ploughed land; 2. Change in cultivation depth (net deepening relative to the original surface); 3. Change in area of land with an archaeological management plan (Loveland & Thompson, 2002).

The two chief pressures on soil from these drivers of change are therefore land use change itself, mainly urbanisation/sealing, and agricultural management changes, such increased ploughing activity and/or drainage installation.

The only realistic units of marginal change from any of the above drivers of change and pressure on soil are those of the area where the change applies. The change may be: 1. The preservation of larger areas of land that contain historic landscape features, or discovered ancient monuments; 2. Imposed changes in land use and management to preserve known archaeological sites; 3. Encouraged changes in agricultural management to avoid potential damage to either known or suspected archaeological remains.

The first example above would tend to apply to wider areas and valuation would have to be both cost based to accommodate economic loss from alternative land uses (i.e. yield loss of crops), as well as that of valuing the public benefit from the presence of the feature. These would be features that the general public may be willing to visit (possibly with payment) and generate economic activity, analogous to areas of certain attractive or special wildlife habitats.

The second example may be solely cost based in that one form of agriculture is favoured over a more profitable type (e.g. grazing versus tillage) but the area or type of monument does not convey much public benefit that can be realised.

The third example may also be cost based if lower yields are to be expected (but if the case for shallower depth tillage is taken then this is not necessarily the case) or additional marginal agricultural costs (extra labour or herbicide applications etc.) incurred. If the archaeological feature remains buried then no public benefit will be realised whilst this remains the case, and there may even be a public disbenefit economically in the form of a payment to the land manager to ensure the changed management. However, some public benefit may be realised, if limited disturbance exploration can continue (e.g. by Ground Penetration Radar) and scientific knowledge is increased during a period of preservation. This may indicate that in the future the site could have greater value than the management fee, if uncovered for public benefaction.

9.4 Conclusions The best indicator of potential impacts of changes in the archaeological protection function of soil is the extent of changes in land use and, in particular, ploughing and

Page 71 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development cultivation depth, and areas protected in an archaeological plan. There is a lack of evidence on the value of the benefits or disbenefits from soil management impacts on the archaeological heritage. Benefits (or disbenefits) from changes in management that affect the archaeological interest will mainly be non-use because such unprotected sites are typically on private land not accessible to the public. At present, only cost-based methods can be used to estimate the value of the benefits. The social cost of payments made under land management plans such as Environmental Stewardship can be interpreted as a lower bound estimate of the public’s WTP for such protection.

Page 72 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

10. Support for Ecological Habitat and Biodiversity

10.1 Technical Review Soil acts as a reservoir of biodiversity in two ways. The first is the diversity of organisms that dwell in the soil matrix itself and contribute to its development and characteristics, whilst the second is the associated diversity of organisms that are supported by the soil, vegetative communities and associated fauna, which are indicative of particular soil types.

Many of the gross biochemical functions and physical functions that communities of soil dwelling organisms carry out, are well known and have been subject to a great deal of study. Nevertheless, many of the organisms themselves and measures of their diversity and functions as communities are unknown, and only now are effective techniques being developed to study them. Many of the gross physical and biochemical processes mentioned underpin the other soil functions described in this report, such as providing fertility for biomass production, ameliorating the impact of pollutants in soil or mediating a beneficial soil structure for adequate water infiltration and storage.

The study and evaluation of the diversity of wider vegetative communities is already practised to a high level of sophistication, though it must be said with usually only a brief and undervalued reference to the soil characteristics. This is usually in reference to a greater geological/landscape feature, such as reference to such communities as “chalk grassland” for instance rather than specific soil features. As such whole habitats can be associated with specific soil features though these are usually identified and valued for specific conservation as whole entities. These may constitute SSSIs, Special Areas of Conservation (SACs) or Special Protection Areas (SPAs) at a local level or even AONBs at the whole landscape level.

The above cases however, only operate where the soil features (and other environmental pressures) have been allowed to dictate the composition of the vegetative community, but the same soil can be managed to support a totally artificial community such as occurs in production agriculture. In some cases this may be a desirable community to conserve in itself. However, as an historic landscape feature and part of our cultural heritage, in most cases the inherent function of the soil to be a reservoir of biodiversity has been “devalued” in both technical and economic senses of the term. In such cases, if the function is to be valued at all then the chief constraints on the soil being able to fulfil this function have to be identified, and the two most important for both above and below ground diversity, are almost universally agreed upon as: 1. Soil disturbance; and, 2. Raised and excess nutrient status.

Whether the specific situation is the evaluation of farmland biodiversity itself, or mitigating the impact of past or neighbouring agricultural activity on a SSSI, the two above factors are those which have to be dealt with. Soil disturbance limits the establishment of perennial plant species, and establishment of certain soil fauna, whilst also causing changes to soil chemistry and microbial function. A history of fertiliser applications raises both N and P levels to those of excess, which favours atypical opportunist species, which out-compete a wider range of more diverse species. Past lime applications at some sites may also have altered the natural soil reaction, and so changing the entire community of plants supported from the original.

Page 73 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Physical changes supported by agriculture, such as the installation of permanent under-drainage, may also have such a wholesale effect.

10.1.1 Factors that Affect Support for Ecological Habitat and Biodiversity The range of factors that have to be addressed for the effective management of sites to support whatever particular habitat is being conserved, are many and can be seen in those identified by CJC Consulting (2004) in their report to Defra, “Cost Effectiveness Study of Approaches for Delivery of PSA Target on SSSIs.” In this report the main factors that have to be accounted for in the management of SSSIs were identified. Management is necessary to improve the condition of SSSIs such that 95% of their area is in “Target Condition” (TC) by 2010 (Defra, 2003). Currently (actually 2003), about 57.5 % are in target condition and 42.5% in unfavourable condition (UC), which means that 88% of UC sites need active management to move them to TC sites to achieve the public service agreement for SSSIs (Defra, 2003). Thirteen areas of management were identified by CJC Consulting (2004) and those relating to soil functioning are: 1. Fertiliser use in agriculture.

A common cause of adverse condition in grasslands (Crofts & Jefferson, 1999), and takes many years of zero application and biomass removal to rectify.

2. In-appropriate agricultural management.

Original standards of management agreed upon for some ESAs etc. have proved ineffective and more stringent measures may need to be agreed upon under the new GAEC HLS schemes.

3. Inappropriate stock management.

Mainly inappropriate feeding regimes in upland areas that have led to localised overgrazing, vegetation trampling and soil erosion (Backshall et al., 2001).

4. Overgrazing.

The prime reason given for sites being in UC over 47 % of the area, and overwhelmingly in upland areas where 90-93 % of upland grassland suffer (Brown et al. , 2001). The result on soil is compaction and erosion, both of which impair the ability of the soil to carry out its former function of maintaining the original biodiversity.

5. Weed control – undergrazing.

A much smaller problem on only 5% of UC sites, where scrub and weed species are allowed to out compete the desired vegetation community. Only a soil problem in that it is often excess nutrients that exacerbate the situation, as all vegetation in the UK (except a small upland area) will progress to a woodland climax community if grazing animals are excluded.

6. Air pollution.

Acid deposition is a major problem in certain sensitive plant communities and critical loads are probably exceeded for over 70% of SSSIs, making it a widespread problem. Critical loads for nitrogen are exceeded on over 90% of SSSI sites (NEGTAP, 2001). Although, aerial deposition can have a direct effect

Page 74 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

through foliage, it is primarily a soil problem, reducing the designated biodiversity function by altering the range of supported vegetation.

7. Moor burning.

A necessary tool in the management of heather moors, which nevertheless can become a problem if allowed to run out of control. If practised on organic soils (upland peat), loss of peat by both oxidation and erosion can occur, possibly removing the soil cover entirely.

8. Freshwater drainage – ditch blocking and management.

Primarily affects wetland communities of upland blanket bogs, wet grasslands, fens and grazing marshes. Management is usually of corrective action to restore wet conditions to the soil profile.

9. Freshwater quality – diffuse pollution.

An effect of excess nutrients in a neighbouring soil profile impacting ecological conditions in a freshwater body. A direct cross reference to the function of environmental services mentioned in above sections, but which could carry specific costs if focussed on particular freshwater SSSI sites, or to address on- farm wetland areas in an HLS scheme.

All of the above problems are not specific to SSSI status and could be encountered in any land management scheme that has its primary objective to enhance or preserve the biodiversity function of the soil at a site. All incur some management cost, which provides a minimum estimate of the implied public benefit.

10.1.2 Drivers of Change that Impact the Soil Function of being a Reservoir of Biodiversity Drivers of change that affect the biodiversity supporting function of soil can be in any of the three categories of either ecological, policy driven or social population changes. Generally they will either lead to a change in land use or change in agricultural management practice and are often analogous to those mentioned under the cultural heritage function of soil and sometimes occur in tandem with it. The main ones considered are: 1. Climate change. The change encouraged by climate change is a fundamental change in the community supported, but may have the effect of reducing the value of current sites of a particular habitat, if this can no longer be supported. 2. Social changes may create amenity areas, mainly of woodland, and the species mixture supported can be directed by the soils on site. 3. WFD or other policy drivers may lead to a change in agricultural land to extensive grassland or woodland, and the vegetation will eventually regress towards the natural community supported by the sols of the area. 4. Specific agricultural policies can cause changes in agricultural management to favour the inclusion of more natural plant communities. The recent CAP and pursuit of GAEC status are such cases, and have led to changes in tillage patterns and fertilise usage. Similarly, changes in pesticide usage will impact species at field margins.

What units of marginal change that may arise form these drivers, are chiefly those of area of over which the change applies. However, there should also be some

Page 75 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

recognition given to the degree of change fostered and the social value that this engenders.

10.2 Valuation of Support for Ecological Habitat and Biodiversity 10.2.1 Public Benefits from Marginal Changes to Support for Ecological Habitat and Biodiversity Valuing the biodiversity and associated landscapes supported by soils is immensely complex. The environmental goods are very diverse, and benefits will vary with the characteristics of these goods, their relative scarcity and their location in relation to where people live.

A number of studies have used the methods of environmental valuation to quantify the benefits provided by sites high in biodiversity (e.g. Harley and Hanley, 1989; Willis, 1990). These include both use and non-use values of the biodiversity. Use values are determined by the number of visitors and the benefits they derive from visits, whereas non-use benefits are mainly determined by the perceived uniqueness or rarity of a site or species and the degree to which it is threatened. Some work has been done on using a ‘Total Economic Value’ system, which takes a sustainable management approach and cuts across a range of soil functions, but by necessity will exclude value for building (Edwards and Abivardi 1998, Guo et al 2001). Clearly, these values may experience marginal changes over time.

A recent study of SSSIs for Defra (CJC Consulting, 2004) concluded that the valuation data were too limited to allow any precise estimate of the overall benefit of SSSI provision. The aggregate use value of SSSIs was possibly in the range £370- £1,110m per year, although this use value cannot all be attributed to biodiversity since visitors enjoy a mix of attributes on SSSIs, which may include landscape and health benefits. There was evidence to suggest that the non-use value of SSSIs exceeds their use value, although no estimate could be made. Biodiversity on areas other than SSSIs, (including ESAs, CSS sites and other farmland and forests) is also valued by the public (e.g. Ecoscope Applied ecologists, 2003; Hanley et al, 2002; Garrod and Willis (1997). We can conclude that the total benefits from the appreciation of biodiversity are very large.

Very little research has been done to quantify the benefits from restoring or enhancing biodiversity, and CJC Consulting (2004) were unable to value the benefits from condition improvements on SSSIs. Similarly, no evidence is available from these valuation studies on the implicit price of the soil services that support biodiversity or the impact on biodiversity benefits of any change in soil services.

Biodiversity change may affect landscape values. These are difficult to quantify because of the extreme variety in landscape compositions and locations. Research on landscape values in the UK has primarily been focussed on the benefits from policy intervention to protect and enhance landscapes. Willis et al . (1993) valued the benefits from the South Downs and Somerset Levels and Moors ESAs (Table 10.1). The aggregate policy benefits were substantial.

Page 76 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 10.1: Benefits of ESA landscapes (£’000s per year)

User Benefits (Visitors Non-user benefits and residents) (general public)

South Downs 48,682 31,153

Somerset levels and Moors 10,758 41,789

Stewart et al. (1997) reviewed estimates of the public benefits from all the agri- environmental schemes in the UK and concluded that the ESA scheme is highly valued by the public for the environmental goods it produces. Whilst many of these goods are landscape features, the valuations relates to a mix of biodiversity and landscape benefits.

A number of studies have dealt specifically with forest landscapes. These vary from hedonic pricing research on the extent to which house prices are affected by the varying proportions of trees in the landscape, to the benefits from changing the structure of trees in the landscape. For example, Willis and Garrod (1992), using a hedonic price model found that a 1% increase in the relative proportion of broadleaved woodland in an area increased the expected selling price of a property by £42.81 (in 1988 prices). The same increase in the relative proportion of mature (mainly Sitka spruce) conifer reduced the expected selling price of a house by £141. A number of studies have shown that woodland in an urban setting has a major impact on house prices where they are in close proximity to houses. Powe et al. (1995) found that properties in an urban area increased in value by £3,441 (approximately 8%) within 500 metres of deciduous trees.

There is evidence to suggest that shape, structure and design of forests matter in landscape values (CJC Consulting, 2005). Entec and Hanley (1997) used a choice experiment (CE) approach in which people were presented with a number of different combinations of landscape attributes and associated price, and asked to select a preferred option or rank the options. It assessed the WTP for forest shape; felling method; species mix in autumn, and winter, and spring. In the most comprehensive study to date, Garrod (2003) has also estimated the WTP for different types of visible woodland from where people lived. Clear preferences were only found for broadleaved woodland in a peri-urban setting. Annual household WTPs varied between £199 and £265 per ha depending on the model used.

10.2.2 Cost-based Approaches Government and its agencies intervene to change land management in support of biodiversity. The main instrument is incentive payments made under Environmental Stewardship, the Environmental Sensitive Areas Scheme, and payments from English Nature and the Forestry Commission. Examples of management changes induced by incentive payments are restrictions on stocking rates and cutting dates, and payments to increase stocking in under-grazed habitats (CJC Consulting, 2004). With forest habitats, payments are made to encourage broadleaved planting and the management of woodlands for biodiversity (CJC Consulting, 2005).

Such payments can be interpreted as lower bound estimates of the social benefits derived. They may be substantially below the public’s WTP for such measures because the valuation evidence suggests that in many cases public benefits greatly exceed policy costs (e.g. Willis et al ., 1993; Hanley et al., 1996).

Page 77 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Only those land management measures that affect the ability of soils to support biodiversity are relevant for the present study. Hence, it would be necessary to determine which measures were primarily directed at changing soil conditions and use the social cost of those measures as a benefit estimate.

10.3 Conclusions It is clear from the literature that the public does value biodiversity and are willing to pay to support policies that enhance habitat and biodiversity. The major element in the valuation is non-use (existence and bequest) value, rather than use value. Changes in soil function will have a cost/benefit wherever the public values the biodiversity supported.

However, the valuation literature is not sufficiently detailed or extensive to allow the value of changes in soil functions on biodiversity to be estimated. All that can be deduced is that the case for protecting soil function will be greatest in the locations that policy associates with highest public value. These are in SACs and non-SAC, SSSIs (CJC Consulting, 2004). Policy already protects such sites under the Habitats Directive and legislation relating to SSSIs and measures are in progress to increase the SSSI area in favourable condition. Such measures address relevant soil issues.

Although economic valuation provides a broad underpinning for intervention to protect and enhance biodiversity, the valuation data are not sufficiently detailed to provide a basis of assessing specific soil-related measures. The social cost of existing policy measures provides a lower bound estimate of public benefit. However, from the set of land management measures used to support biodiversity it would be necessary to identify the cost of the subset of measures that was directed primarily at changing soil services. This should be possible but we are not aware that it has been done to date.

More valuation information would be valuable for policy development. The appropriate methods are those of contingent valuation and/or choice experimental approaches. The area is challenging because the non-use element tends to dominate biodiversity values. A sample frame and elicitation method that facilitates raising to the relevant population is thus required. Even so, estimating the benefits from specific soil-based policy measures should be feasible,

Page 78 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

11. Conclusions

11.1 Methodology Valuing a change in soil functions or services requires a two-stage approach. First the technical effects need to be quantified. This was done by identifying the main drivers of change in soil functions and then estimating the per unit impact of each driver on the relevant soil function. In most cases the effects are context dependent and the environmental effects may not be indirect (e.g. through impacts on linked environmental assets such as water resources or biodiversity).

The second stage is one of economic valuation. Three methods have been used: • Applying market prices where the effects are on marketed outputs (e.g. nutrient cycling impacts on crop production); • Benefits/costs to the public in terms of changes in public welfare, where impacts are on environmental goods (evidence from economic valuation studies); and • Cost-based approaches where the aim is to estimate the cost of compensating measures for the loss in soil function.

Of these, the cost-based measures are less informative, because they indicate nothing about the benefits or losses to public welfare. However, in many cases a lack of direct valuation evidence necessitates a cost-based approach. The case for obtaining new primary data is considered below (11.3).

11.2 Evidence on the Marginal Valuation of Soil Functions 11.2.1 Carbon Storage and Sequestration Soil has an important role in storing carbon. The quantities of sequestered carbon in soils in England are broadly in equilibrium after a long period of decline. However, soil carbon levels can be increased especially through changes in the management of arable land and through changes in land use. At a social value of £70 per tonne, the value of sequestered SOC is substantial, with a present value of up to £1,380 per ha (3,160 for total sequestered carbon) where arable land is converted to woodland. Drainage or intensive use of peatland, and land conversion from permanent grass are associated with high negative marginal values due to losses of stored carbon. The magnitude of these effects indicates a case for soil protection in some cases although a complete analysis would need to include the opportunity cost of protection in terms of any lost output.

11.2.2 Water Storage and Flow Mediation Several drivers can change the ability of soils to hold water. In locations subject to risk of flooding these effects can be important. It is not possible to generalise about the change in risk or cost that follows from a change in soil function because impacts are location specific and require individual assessment. The major effect is through paving over and construction, and to a lesser extent, afforestation. Impacts may be transferred downstream; consideration of whole catchment effects is therefore essential where any downstream part of a catchment is subject to flood risk. The effects of forestry are less clear. Where flood risk is an issue site-specific assessment are needed and MAFF (2001b) has indicated the appropriate cost-benefit procedures that should be used. One way forward would be to model increased flood risks in catchments where significant housing or infrastructure development is planned.

Page 79 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Alternative scenarios could be compared to calculate the value of different soils in the catchment for the water storage function.

11.2.3 Nutrient Cycling Reduction in the ability of soils to fix nitrogen from the atmosphere can result from an accumulation of heavy metals in the soil and from reductions in soil pH (Chaudri et al. 1993). It is hoped that legislation such as the Waste Framework Directive and the Sludge (Use in Agriculture) Regulations can prevent heavy metal accumulation to levels that would impact on the N fixing function of soils. However, this regulatory framework does not control the addition of pollutants from atmospheric deposition or from other land spread materials. As a result, it is important that pollutant concentrations and their effect on the soil microbial population continue to be monitored outside the statutory arrangements imposed by the regulations mentioned above. Soils also provide nitrogen to growing crops through the mineralisation of organic matter. This ability to cycle N can be reduced in two ways. Firstly, a reduction in the organic N pool can result from climate change and from agricultural systems where the soil is cultivated annually and does not receive regular additions of organic matter to compensate crop removal. If the organic matter content of a soil is reduced from 20% to 10%, the amount of N released per annum can be reduced by 110 kg/ha (MAFF, 2000a). Secondly, the ability of the soil microbial pool to mineralise organic matter can be adversely affected by a range of contaminants. More research is required to determine the marginal change in microbial nutrient cycling that results from increases in diffuse contamination.

Losses in crop yield can be expected as a result of deviations from agricultural BMPs. Allowing soil P and K reserves to run down below critical thresholds can reduce crop yields by 25-30% for winter wheat, 15-60% for winter barley and 15-30% for grass. Reduction in P reserves can have major effects on crop revenue through yield reductions, especially for sugar beet and wheat. With K the revenue losses are more variable depending on crop and location. Potatoes are particularly susceptible to a loss in output if K levels are reduced.

Allowing soil pH to fall below critical thresholds can reduce yields of potatoes by 15%, and winter cereals by up to 90%. Of the cereal crops, winter barley is the most sensitive with yield reductions evident below a soil pH of 5.8. The economic effects are substantial and emphases the importance of maintaining an appropriate pH to optimise soil functions.

The amount of yield reduction caused by surface compaction depends on the degree and extent of the soil compaction, but can range from 10 to 50%. Yield reductions caused by compaction below the ploughed layer are restricted to spring crops on light soils in years of moderate to severe drought.

Potentially toxic elements will only reduce crop yields if concentrations are allowed to exceed the Sludge (Use) in Agriculture 1989 limits by several hundred milligrams per kg of soil. As long as the limit levels are not exceeded, we would not expect any significant reductions in yield.

Climate change is not expected to give rise to any great shift in the ability of soils to support vegetative growth.

The value of the marginal loss in soil function can be estimated by applying market prices to the lost output. Where major reduction in the P or K index or in pH occur the

Page 80 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

losses are substantial for crops sensitive to the loss in soil service. It emphasises the importance of maintaining the nutrient cycling capacity of the soil through adequate applications of maintenance P and K, and through liming. Policy intervention to correct any loss of soil function would only normally be appropriate where there is evidence of market failure. It is not clear where such failure might occur so long as producers are clearly informed about the consequences of reduced nutrient cycling capacity. Hence, the value of marginal changes to this function will be informed from the market.

11.2.4 Supporting Construction For the majority of construction work any variation in a soil’s ability to support development will have a very minor impact on land values. This reflects the substantial premium that development land typically has over any other uses. This premium is determined by local demand for domestic or commercial building within the constraints imposed by the planning system. The technical evidence is that, for most construction activity, soil condition has little effect on development costs as indicated in land prices. The exception is where access for construction is a significant cost element, and this may apply, for example, to pipelines and wind farm construction.

However, since construction is a market activity any variation in soil value for construction should be accounted for in land prices and rents. The main variation in soil-related cost will reflect differences in soil type and rainfall. Impacts of changes in a soil’s ability to support construction will be minor. Given that construction is a market activity the policy case for protecting soil in this context is weak unless there are associated negative impacts on environmental services. Hence, the value of marginal changes to this function will be informed from the market.

11.2.5 Natural Attenuation of Pollution/Contamination A number of drivers can impact on the ability of soil to perform the attenuation function. In the urban context, severe industrial contamination can reduce this ability as the capacity of the soil to absorb and degrade pollutants is often exceeded. However, at certain levels of hydrocarbon contamination, natural attenuation processes or bioremediation can be used to avoid the need for landfilling with polluted materials. Where pollutants have accumulated in soil to levels below threshold limits as defined by soil guideline values, the attenuation function avoids the need for costly treatment altogether. The drivers that threaten the attenuation function in urban settings are contamination and urbanisation themselves. Climate change may also alter the soil’s capacity to degrade and dilute certain pollutants, particularly where the trend is towards drier conditions. In the rural setting, a number of drivers reduce the ability of soil to attenuate pollution from diffuse sources. These include: land use change from grassland or long-term set-aside to tillage; changes in pH; saturation with Heavy metals; afforestation; construction; and deviation from best management practices (BMPs) on agricultural land.

Changes in pH are likely to impact on the soil’s ability to attenuate heavy metal pollution and acid deposition. Saturation of the soil with Heavy metals affects the soil’s ability to attenuate heavy metal pollution. Saturation of the soil with Heavy metals to the 1989 Regulation limits is very unlikely to impact on losses to surface or ground waters.

Page 81 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

The impacts of afforestation are analogous to a change from arable to permanent set aside, although there would be changes to the soil water budget that could reduce soil drainage volumes. Under afforestation the impacts of acid deposition and heavy metals are likely to increase. Pathogens are largely eliminated under forestry, but where livestock are present, the soil’s ability to attenuate FIOs is increased under established forest due to vegetative cover and lower run-off volumes.

Construction will increase run-off volumes due to surface sealing. However, in most cases the source of the pollution is removed and urban pollutants replace rural pollutants. The soil’s ability to attenuate urban pollutants is lost, except where sustainable drainage systems (SUDS) are used, and the pollutant problem is exported to another location.

Deviation from BMP will reduce the ability of the soil to attenuate the majority of pollutants. Relative to the baseline situation given in Defra project ES0203, P pollution may increase by 0.3-2.2 kg TP/ha/yr., N pollution by 1-45 kg N/ha/yr. and FIOs by 10- 50% (all ranges depending on the BMP in question). Sediment and pesticide losses are likely to increase in proportion to P losses, although soluble pesticides may also increase with N losses.

The valuation of changes in a soil’s natural attenuation of pollution/contamination will depend on which pollutant is affected. Changes in leaching of heavy metals, N and P and movement of pathogens may have implications for water quality in associated catchments. This will be the case where there are existing risks in relation to the WFD, where drinking water abstraction takes place or where eutrophication is an environmental issue. We were only able to quantify the possible marginal costs of a loss of attenuation capacity for N and P using policy costs under the WFD. These varied substantially depending on the extent of the change in attenuation capacity but could be high where BMP was not adhered to. Hence, the value of marginal changes to this function will be informed from policy costs.

11.2.6 Archaeological Site/Artefact Protection The best indicator of potential impacts in the archaeological protection function of soil is the extent of changes in land use and, in particular, ploughing and cultivation depth, and areas protected or in an archaeological plan. There is a lack of evidence on the value of the benefits or disbenefits from soil management impacts on the archaeological heritage. Benefits (or disbenefits) from changes in management that affect the archaeological interest will mainly be non-use because such unprotected sites are typically on private land not accessible to the public.

11.2.7 Support for Ecological Habitat and Biodiversity The factors that should be considered with regard to changes of this function are soil disturbance and nutrient status, that is whether it is cultivated or has fertiliser (including lime) applied to the soil along with physical changes such as under- drainage systems.

It is clear from the literature that the public value biodiversity and are willing to pay to support policies that enhance habitat and biodiversity. The major element in the valuation is non-use (existence and bequest) value, rather than use value. Changes in soil function will have a cost/benefit wherever the public value the biodiversity supported.

However, the valuation literature is not sufficiently detailed or extensive to allow the value of changes in soil functions on biodiversity to be estimated. All that can be

Page 82 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

deduced is that the case for protecting soil function will be greatest in the locations that policy associates with highest public value. These are in SACs and non-SAC, SSSIs (CJC Consulting, 2004). Policy already protects such sites under the Habitats Directive and legislation relating to SSSIs and measures are in progress to increase the SSSI area in favourable condition. Such measures address relevant soil issues.

11.3 Appropriate Methodologies for Measuring Benefits The economics literature mainly concentrates on explaining variation in the market price of land in terms of its underlying attributes. Of particular interest has been the impact of farm support mechanisms on land prices and the extent to which CAP support has been capitalised into land. Although hedonic pricing studies do provide marginal values for soil attributes there are few in the UK and they are neither especially reliable nor useful for the present study.

Cost-based approaches to valuation can almost always be used assuming there is adequate technical knowledge. However, they provide, at best, a lower bound estimate of public benefit. They are thus very limited as a guide for policy development unless soil policy measures form a subset of a wider set of options for an existing policy (as with carbon emissions policy). In such cases cost approaches are valuable for selecting least-cost mechanisms. Direct estimates of benefit that relate to soil-based measures are sparse but can provide a wider context for the benefits from soil functions (e.g. with biodiversity as an output).

Table 11.1 summarises our view on the most appropriate methods for determining the social benefits from changes to soil services and functions using the currently available data. The table also indicates areas where further research would enable improvements to be made in the estimation of benefits from changes to soil functions, and the appropriate methods to use.

Page 83 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Table 11.1: Methods for valuing the benefits from soil-related policy measures

Appropriate methodology for assessing benefits Data requirements to facilitate estimation of Function from soil –related measures using existing data benefits from soil-related measures Carbon storage Estimates of the social cost of carbon provide a None recommended and good basis for analysis sequestration Water storage A cost-based approach (damage prevented; Benefit assessment is highly and flow effects on water treatment costs) coupled with site/catchment specific. There is a case for mediation public risk assessment is appropriate. further research to quantify benefits and risk effects in contexts where soil impacts are expected to be important. Stated preference methods are appropriate. Nutrient cycling Market effects are best analysed using available More research is required to determine the market price and cost information. Where there marginal change that results from are additional environmental effects, an aggregate increases in diffuse contamination analysis is required using the methods suggested in this table. Supporting A cost-based approach is the appropriate None recommended construction methodology for assessing changes in a soil’s ability to support construction. This reflects the fact that any effects would be accounted for in market transactions for land and the buyer’s WTP. Natural In the absence of direct measures of public A case can be made for the evaluation of attenuation of benefit, cost-based approaches are the benefits from measures to protect the pollution/ recommended. These can be used where pollution attenuation capacity of soils. contamination changes in soil management has implications for Stated preference methods are costs to the private sector or other public sector appropriate. A meaningful policy context policy measures. It should be possible to estimate would need to be defined and the methods such costs for changes in the attenuation capacity of environmental valuation applied (Garrod of soils at least for N and P when the research of and Willis, 1999). This would incorporate Cuttle et al. (2006) is completed. down stream impacts including eutrophication where appropriate. Where contaminated land is concerned, a cost- Additional research on water company cost based approach could also be used to assess the savings from attenuation measures is value of the soil’s ability to absorb or degrade needed. pollutants and the ability to bioremediate materials. However, there is also public benefit in Information on the quantity of contaminated reducing the number of journeys required to soils that can be re-used due to import topsoils (onto restoration sites) and send remediation would provide an indication of contaminated materials to landfill. Bioremediation the amount of soil resource, with its many and natural attenuation also reduces the need to functions and considerable value, that is further mine the nation’s topsoil resource. preserved as a result.It would also provide the volume that would otherwise have to be transported and landfilled. Archaeological At present, only cost-based methods can be used This is an under-researched area which site/artefact because of a lack of relevant valuation research. would benefit from economic valuation of protection The social cost of payments made under agri- the public benefits from soil protection. environmental land management programmes that Stated preference methods are relate to soils provide a lower bound estimate of appropriate. the public’s WTP for such protection. Support for Valuation data are not sufficiently detailed to More economic valuation is needed to ecological provide a basis of assessing specific soil-related quantify the benefits from protecting the habitat and measures. The social cost of existing policy biodiversity functions of soils. Stated biodiversity measures provides a lower bound estimate of preference methods are appropriate.Lower public benefit. Only the subset of land bound cost-based estimates of soil management measures that aim to change soil protection or enhancement measures could services is relevant. be derived from a more detailed analysis of the land-based policy measures currently in use to support biodiversity.

Page 84 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

11.4 Information Gaps Table 11.1 indicates the main gaps in the economic valuation data, and suggests the generic methods that could be used to fill these gaps. In all cases these are stated preference methods, and CV and/or choice experiments would be the appropriate tools. A specific design cannot be proposed without more detail on the policy context and information on the technical impacts that would define the changes in services and outputs from the soil(s) in question.

One development of this study that we would recommend as a further project is to take a number of policy issues relating to soil services and work up the corresponding technical datasets together with an outline valuation study that would produce benefit estimates. This would take forward the practical application of methods to generate new benefit estimates for soil services. It would also provide greater clarity on whether benefits from change to soil services can be extracted from policies that have wider impacts (e.g. through changes in land management). Although experimental approaches have the potential to achieve this, more research is needed to establish the practical feasibility of structuring realistic choice formats.

More generally, the main requirements for technical and economic information to improve the evidence base are: • Linking technical measures of changes in water quality to benefit estimates. New benefit measures specified in technical and economic terms may be needed. • Improved catchment water use models must be used to derive the marginal change in water storage and flood peak flows for land use changes in catchments. Also required is associated cost information for specific catchment contexts and estimates of the benefit from reduced flood risk • It is acknowledged that some soil types are more fertile and easier to farm than others, but this is not directly reflected in land values, which are a complex integration of many other features. A more direct valuation of soil type for crop production should be investigated. • Information on the marginal change in soil functions across the range of soil types. Even for biomass production, where trials have been carried out on a number of soil types, information is not available for the impact of reducing soil reserves on all soil types. • Specific figures for the loss of pollutants from soil, as a result of various pressures. Figures provided by Defra project ES0203 (The cost-effectiveness of integrated diffuse pollution mitigation measures) limit marginal losses of P, N and FIOs to specific farm models, two soil types and a single climate. There are not sufficient data to make robust predictions for a variety of land use-soil-climate scenarios, which hinders extrapolation to the national level. Information is particularly lacking for the marginal change in the ability of soil to attenuate release of pesticides and FIOs from land. • More work is needed to improve the information base relating to potentially important drivers of afforestation and its impact on flood risks (afforestation and construction). • There is a lack of linkage between water quality indicators (e.g. N, P, faecal indicators) and estimates of public benefit from changes in these indicators. Economic valuation of soil-related measures is required. .More stated preference research is needed to identify the public’s WTP for specific measures that improve water quality.

Page 85 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

• More research is required to determine the marginal change in microbial nutrient cycling that results from increases in diffuse contamination. • Benefit estimates for soil and land management policies that protect the archaeological, pollution attenuation and biodiversity capacity of soils (see Table 11.1).

11.5 Policy Implications Where a change in soil function only affects marketed output, for example crop yield, the case for policy intervention is weak unless there is evidence of market failure. Policy relating to soil functions should mainly focus on the role of soils in supporting environmental services including those provided by water resources and biodiversity.

Hence the main policy implications relate to the environmental functions of soils. Valuation information is incomplete and this limits the economic analysis that can be used in interpreting the impacts of changes in soil functions and services. The information available supports the case for soil protection to minimise losses of sequestered carbon and take account of carbon effects in other land-based environmental polices (e.g. afforestation). Impacts on flood risk especially from paving over appear important but impacts are catchment-specific. It proved difficult to quantify the marginal costs of losses in attenuation pollution. However, there was a case for intervention to facilitate BMP in cases of major deviation from BMP. There was evidence of public benefit from the soil’s function in supporting biodiversity and this is likely to be most important in areas of high ecological value.

With regard to obtaining new valuation data on soil services, we suggest (Table 11.1) the areas where improved valuation is likely to give the greatest pay-off for policy development, and the appropriate techniques to use. However, this is a developing area of research and there is no universal best method that can be applied routinely. Further work is needed to design specific studies. In this, numerous issues need to be considered including the soil(s) in question, the scale of the study (local, national), and whether the aim is to value changes in soil services and/or aggregate benefits from specific soil policy measures.

Page 86 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

12. References Armentano T. V. and Menges E. S. 1986. Patterns of change in the carbon balance of organic soil-wetlands of the temperate zone. Journal of Ecology , 74, 755-774.

Arnold, P.W. and Shepherd, M.A. 1990 Phosphorus and Potassium Requirements of Cereals. HGCA Review No. 16. Home Grown Cereals Authority, 1990.

Backshall, J., Maley, J. and Rebane M. (Eds.)2001. The Upland Management Handbook. English Nature, Peterborough, UK.

Bailey A. D., Dennis C. W., Harris G. L. and Horner M. W. 1980. Pipe size design for field drainage. ADAS, Land Drainage Service, Research and Development Report No. 5. Pub. MAFF, 1980, London. 53pp

Ball, B. C. and Ritchie, R.M. (1999) Soil and residue management effects on arable cropping conditions and nitrous oxide fluxes under controlled traffic in Scotland: 1. Soil and crop responses. Soil and Tillage Research, Volume 52, Issues 3-4, October 1999, 177-189. Barclay, G. J. 1992. Forestry and archaeology in Scotland. Scottish Forestry 46(1): 27-47.

Barnard, C., Whittaker, G., Westenbarger, D. and Ahearn, M. 1997. Evidence of capitalisation of direct government payments into US cropland values. American Journal of Agricultural Economics 79, 1642-1650.

Bateman, I., Willis, K. and Garrod, G. 1994. Consistency between contingent valuation estimates: a comparison of two studies of UK national parks. Regional Studies 28 454-474.

Bateman, I.J. and A. A. Lovett, “Estimating and Valuing the Carbon Sequestered in Softwood and Hardwood trees, timber products and forest soils in Wales” , Journal of Environmental Management 60, 301-323, 2000

Bateman, I. J., Day, B. H. Dupont, D., Georgiou, S. Matias, N.G.N. and Subramanian, L. 2006. Does phosphate treatment for prevention of eutrophication pass the benefit cost test? in Pearce, D. W. (eds.). Valuing the Environment in Developed Countries. Edward Elgar, Cheltenham.

Batjes N. H. 1998. Mitigation of atmospheric CO2 concentrations by increased carbon sequestration in the soil. Biology and Fertility of Soils, 27, 230-235.

Bellamy P. H., Loveland P. J., Bradley R. I. Lark R. M. and Kirk G. D. 2005. Carbon losses from all soils across England and Wales 1978-2003. Nature, 437, 245-248.

Bhogal, A., Nicholson, F. A., Chambers, B.J. and Shepherd, M. A. 2003 Effects of past sewage sludge additions on heavy metal availability in light textured soils: implications for crop yields and metal uptakes. Environmental Pollution 121, 413-423.

Bolton, J. 1977 Changes in soil pH and exchangeable calcium in two liming experiments on contrasting soils over 12 years. Journal of Agricultural Science, Cambridge 89, 81-86.

Brainard, J., Lovett, A. and Bateman, I. 2003. Carbon Sequestration Benefits of Woodland. Social & Environmental Benefits of Forestry Phase 2 , Report to the

Page 87 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Forestry Commission, Edinburgh. Centre for Research in Environmental Appraisal and Management, University of Newcastle upon Tyne.

Brown, A., Hall, R. and Townshend, D. (Eds.)2001 State of Nature. The upland challenge. English Nature, Peterborough, UK.

Bullard M.J.. and Metcalfe, P. 2001 Estimating the energy requirements and CO 2 emissions from the production of the perennial grasses Miscanthus , switchgrass and reed canary grass. Report to ETSU B/U1/00645/REP. Publ.. DTI URN 01/797. 93pp. Canarache, A., Colibas, I, Colibas, M,, Horobeanu, I., Patru, V., Simota, H., and Trandafirescu, T. 1984 Effect of induced compaction by wheel traffic on soil physical properties and yield of maize in Romania . Soil and Tillage Research, Volume 4, Issue 2, March 1984, 199-213. Cannell M. G. R., Milne R., Hargreaves K. J., Brown T. A. W., Cruickshank M. M., Bradley R. I., Spencer T., Hope D., Billett M. F., Adger W. N. and Subak S. 1999. National inventories of terrestrial carbon sources and sinks: the UK experience. Climate Change, 42, 505-530.

Cape J.N., Brown A. H. F., Robertson S. M. C., Howson G. and Paterson I. S. 1991. Interspecies comparisons of throughfall and stemflow at three sites in northern Britain. Forest Ecology and Management, 46, 165-177.

Carlton-Smith, C. H. and Stark, J.H. 1987 Sites with a History of Sludge Deposition. Interim Report of Field Trials: April 1983-March 1986 (SDA 9166 SLD). WRc Report No. 1047-M/2. WRc Medmenham, Marrow.

Chadwick, D. R., Chambers, B.J., Anthony, S., Granger, S., Haygarth, P.M., Harris, D. and Smith, K. 2006 A measure-centric approach to diffuse pollution modelling and cost-curve analysis of mitigation measures. In Proceedings of the SAC and SEPA Biennial Conference, Edinburgh, 5-6 April, 2006. Chaudri, A.M., McGrath, S. P., Giller, K.E., Rietz, E., and Sauerbeck, D. R. 1993 Enumeration of indigenous Rhizobium leguminosarum biovar Trifolii in soils previously treated with metal-contaminated sewage sludge. Soil Biology and Biochemistry 25, 3: 301-309.

Chambers, B.J. and Garwood, T.W.D. 1998 Lime loss rates from arable and grassland soils. Journal of Agricultural Science, Cambridge 131, 455-464.

CJC Consulting 2003. Economic analysis of forestry policy in England . Report for Defra/H. M. Treasury. http://statistics.defra.gov.uk/esg/evaluation/forestry/default.asp

CJC Consulting 2004 Cost Effectiveness Study of Approaches for Delivery of PSA Target Relating to SSSIs. Final Report for Defra, www.cjcconsulting.co.uk (accessed 31/03/06).CJC Consulting (2005) Review of Evidence for the Formulation of Forestry Policy in England. Report for Defra. http://www.defra.gov.uk/corporate/consult/forestry-strategy/review-evidence . pdf

CJC Consulting 2005 Review of Evidence for the Formulation of Forestry Policy in England . Report for Defra. http://www.defra.gov.uk/corporate/consult/forestry- strategy/review-evidence . pdf

Clarkson, R and Deyes, K. 2002. Estimating the Social Cost of Carbon Emissions . GES Working Paper 140. London: HM Treasury. http://www.hm- treasury.gov.uk/Documents/Taxation_Work_and_Welfare/Tax_and_the_Environment/ tax_environ_index.cfm

Page 88 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Conradie, B. I. and Hoag, D. L. 2004. A review of mathematical programming models of irrigation water values. Water SA 30, 287-292.

Costanza, R., R. d'Arge, R. de Groot, S. Farber, M. Grasso, B. Hannon, K. Limburg, S. Naeem, R.V. O'Neill, J. Paruelo, R.G. Raskin, P. Sutton, and M. van den Belt 1997. The Value of the World's Ecosystem Services and Natural Capital, Nature 387, 253- 260.

Courty, M. A., Goldberg, P. and Macphail, R. I. 1989 Soil Micromorphology in Archaeology. Pub. Cambridge University Press, Cambridge, UK.

Crabtree, J. R., Peterken, G. and MacDonald, D. 2002. Review of the Native Woodland Pilot Plan Scheme . Report for Forestry Commission England. CJC Consulting, Oxford.

Crofts, A. and Jefferson R. G. (Eds.) 1999. The lowland grassland management handbook, 2 nd edition. English nature and the Wildlife Trusts, Peterborough, UK.

Currie, J. M. 1981. The Economic Theory of Agricultural Land Tenure , Cambridge University Press.

Cuttle, P. M., P. M. Haygarth, D. R. Chadwick P. Newell-Price, D. Harris, M. A. Shepherd, B.J. Chambers and R. Humphrey 2006. An Inventory of Measures to Control Diffuse Water Pollution from Agriculture. User Manual . Draft Report for Defra.

Davis, J. and O’Neill, C. 1992. Discrete-choice valuation of recreational angling in Northern Ireland. Journal of Agricultural Economics 43 452-457.

Davis, R. D. and Carlton-Smith, C.H. 1984 An investigation into the phytotoxicity of zinc, copper and nickel using sewage sludge of controlled metal content. Environmental Pollution ( Series B) 8, 163-185.

Defra 2005 CC0272: Synopsis and review of relevant projects to assess mitigation options for nitrous oxide and methane to inform the Climate Change Programme

Defra 2003 Sites of Special Scientific Interest: Encouraging positive partnerships. Defra, London.

Defra a 2005 Soils in the Built Environment - A Strategy for the Construction Sector. Report produced for Defra by WSP Environmental.

Defra b 2005 Defra Guidance for Successful Reclamation of Mineral and Waste Sites. Published by the Department for Environment, Food and Rural Affairs. Defra 2006 OF0316 : The development of improved guidance on the use of fertility building crops in organic farming (CTE0204) Dragovich, D. 1990. `Does Soil Erosion Matter to People in Metropolitan Sydney?'. Australian Journal of Soil and Water Conservation, 3(1): 29-32.

Ecoscope Applied Ecologists 2003. Review of Agri-environment Schemes – Monitoring Information and R and D Results. Report for Defra.

Edwards P. J. and Avibardi C 1998. The Value of Boidiversity: Where Ecology and Economy Blend. Biological Conservation 83, 3 pp239 – 246 1998.

Eftec 2004. Framework for Environmental Accounts for Agriculture. Report for Defra, DARD, Scottish Executive and Welsh Assembly Government.

Page 89 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Eftec 2005 Valuation of the Historic Environment. Report for English Heritage, the Heritage Lottery Fund, the Department for Culture, Media and Sport and the Department for Transport.

Eftec 2006. Economic Valuation of Environmental Impacts in the Severely Disadvantaged Areas . Final Report for Defra.

Entec and Hanley, N. D. 1997. Valuing Landscape Improvements in British Forests. Report to the Forestry Commission. Entec UK, Leamington Spa, and Environmental Economics Research Group, Stirling University.

Entec 2004. Study into the Environmental Impacts of Increasing the Supply of Housing in the UK. Report for Defra. Erhardt, W., and Prüß, A. 2001 Organic Contaminants in Sewage Sludge for Agricultural Use. EU Project for European Commission Joint Research Centre, Institute for Environment and Sustainability, Soil and Waste Unit. 18 October 2001.

Ervin, D. E., and J. W. Mills 1985. Agricultural Land Markets and Soil Erosion: Policy Relevance and Conceptual Issues. Amer. J. Agr. Econ. 67, 938-42.

Faux, J. and Perry G. M. 1999. Estimating irrigation water value using hedonic price analysis: A case study in Malheur county, Oregon. Land Economics 75 (3) 440-452.

Ferrier, R. C., Miller, J.D., Walker, T. A. B. and Anderson, H. A. 1990 Hydrochemical changes associated with vegetation and soils. In The Surface Waters Acidification Programme (ed. B.J. Mason ), pp. 57-67. Cambridge University Press, London.

Flood Hazard Research Centre 2005 The Benefits of Flood and Coastal Risk Management: A Manual of Assessment Techniques . Middlesex University Press.

Forestry Commission 2000. Forests and Water Guidelines 3rd edition, Forestry Commission, Edinburgh.

Förster, J. 1998 The influence of location and season on the concentrations of macro- ions and organic trace pollutants in roof runoff. Water Science and Technology. Vol. 38, No 10 pp 83–90.

Gardner, H.W. and Garner, H.V. 1957 The Use of Lime in British Agriculture. E. & F. N. Spon, London, 216pp.

Garrod, G. D. 2003. Landscape Values of Forests . Social & Environmental Benefits of Forestry Phase 2, Report to the Forestry Commission, Edinburgh. Centre for Research in Environmental Appraisal and Management, University of Newcastle upon Tyne.

Garrod, G and Willis, K. 1997. The non-use benefits of enhancing forest biodiversity: a contingent ranking study. Ecological Economics 21, 45-61.

Garrod, G. and Willis, K. 1999. Economic Valuation of the Environment . Edward Elgar, Cheltenham.

Gasser, J.K.R. 1973 An assessment of the importance of some factors causing losses of lime from agricultural soils. Experimental Husbandry 25, 86-95.

Georgiou, S., Bateman, I., Cole, M. and Hadley, D. (2000). Continent Ranking and Valuation of River Water Quality Improvements: Testing for Scope, Sensitivity,

Page 90 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Ordering and Distance Decay Effects . CSERGE Working Paper 2000-18. University of East Anglia.

Grant, R. O. and Olesen, S. E. 1984 Sludge utilisation in spruce plantations on sandy soils. In Berglund, S., Davis, R. D. and L’Hermite, P. (eds.) Utilisation of Sewage Sludge on Land: Rates of Application and Long-term Effects of Metals . D. Reidel Publishing Company, Dordrecht, pp. 79-90.

GFA-Race 2003. The Potential Environmental Impacts of the CAP Reform Agreement . Report for Defra.

Gibbs, P.A., Chambers, B.J., Chaudri, A.M., McGrath, S. P., Carlton-Smith, C. H., Bacon, J. R., Campbell, C.D. and Aitken, M.N. 2006 Initial results from a long-term, multi-site field study of the effects on soil fertility and microbial activity of sludge cakes containing heavy metals. Soil Use and Management 22, 11-21.

Green, C. H. and Tunstall, S. M. 1991. The evaluation of river water quality improvements by the contingent valuation method. Applied Economics , 23 1135-1146.

Goodlass, G. and Welch, W. 2005 British Survey of Fertiliser Practice: Fertiliser Use on Farm Crops, 2004. ISBN 1 86190 128 3.

Guo Z, Xiao X, Gan Y, Zheng Y 2001 Ecological Economics 38 pp141 – 154

H. M. Treasury 2006. The Green Book: Appraisal and Evaluation in Central Government . HM Treasury, London. www.hm-treasury.gov.uk

Hall M. J., Hockin D. L. & Ellis J. B. 1993. Design of Flood Storage Reservoirs. Publ.., CIRIA/Butterworth-Heinmann.

Hanley, N. 1989. Problems in valuing environmental improvements resulting from agricultural policy changes: the case of nitrate pollution. In: Dubgard, A and Nielson, A. (eds.) Economic Aspects of Environmental regulations in Agriculture . Wissenschaftsverlag Vauk Kiel, Keil.

Hanley, N. D., Willis, K. .G., Powe, N. A. and Anderson, M. 2002. Valuing the Benefits of Biodiversity in Forests . Report to the Forestry Commission, Edinburgh. Centre for Research in Environmental Appraisal and Management, University of Newcastle.

Hanley, N., Macmillan, D.C. Wright, R., Bullock, C. H., Simpson, I., Parsisson, D., and Crabtree, J. R. 1996. Valuation of the Conservation Benefits of Environmentally Sensitive Areas . Report to SOAEFD Edinburgh.

Harley, D.C. and Hanley N. D. 1989. Economic Benefits of Nature Reserves: methods and results. Discussion Paper in Economics 89, Department of Economics, University of Stirling.

Harvey, D. 1991. The Agricultural Demand for Land: its Availability and Cost for Forestry . Countryside Change Unit, University of Newcastle.

Helm, D. 2005 (ed.). Climate-change Policy . OUP, Oxford.

Hodge, I., Reader, M., Revoredo, C., Crabtree, J. R., Tucker, G. and King, T. 2006. Project to Assess Future Options for Set-aside . Report to Defra. University of Cambridge.

Page 91 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Holtan, H. N. and Kirkpatrick, M. H. 1950. Rainfall, infiltration and hydraulics of flow in runoff computation. Transactions of the American Geophysical Union, 31. pp 771-779.

Hulme, M., Jenkins, G.J., Lu, X., Turnpenny, J. R., Mitchell, T. D., Jones, R.G., Lowe, J., Murphy, J.M., Hossell, D., Boorman, P., McDonald, R. and Hill, S. 2002 Climate Change Scenarios for the United Kingdom: the UKCIP02 Scientific report, Tyndall Centre for Climate Change Research, School of Environmental Sciences, University of East Anglia, Norwich, UK. 120pp.

Jarvis, R. A., Bendelow, V. C., Bradley, R.I., Carroll, D. M., Furness, R.R., Kilgour, I.N.L. and King, S.J. 1984 Soils and their Use in Northern England. Soil Survey of England and Wales No. 10. Harpenden.

Johnston, A. E. 1975 Experiments made on Stackyard Field, Woburn, 1876-1974. I. History of the field and details of the cropping and manuring in the Continuous Wheat and Barley experiments.

Johnston, A. E. 2001 Principles of crop nutrition for sustainable food production. Proceedings No. 459. The International Fertiliser Society, York, UK. 39pp.

Johnston, A. E., Lane, P.W., Mattingley, G.E.G., Poulton, P. R. and Hewitt, M. V. 1985 Effects of soil and fertiliser P on yields of potatoes, sugar beet, barley and winter wheat on a sandy clay loam at Saxmundham. Journal of Agricultural Science, Cambridge 106, 155-167.

Johnston, A. E., Poulton, P. R., Dawson, C. J. and Crawley, M.J. 2001a Inputs of Nutrients and Lime for the Maintenance of Fertility of Grassland Soils. Proceedings No 486, The International Fertiliser Society, York, UK. 40pp.

Johnston, A. E., Goulding, K.W.T, Poulton, P. R., and Chalmers, A.G. 2001b Reducing Fertilizer Inputs: Endangering Arable Soil Fertility? Proceedings No 487, The International Fertiliser Society, York, UK. 44pp.

Johnston, A. E., Dawson, C. J. 2005 Phosphorus in Agriculture and in Relation to Water Quality. Agricultural Industries Federation, 2005.

Johnston, A. E. and Whinham, W.N. 1980 The use of lime on agricultural soils. Proceedings No. 189. International Fertiliser Society, York, UK. pp 4-31. Jones, D., Lethbridge, G., McCarthy, P., & Thompson, S. 2001 Project SIReN: Phase 2a Conceptual Site Model & Groundwater Model. R&D Technical Report P2- 208/TR/2, Environment Agency, UK.

Just, R. E. and Miranowski, J. A. 1993. Understanding farmland price changes. American Journal of Agricultural Economics 75, 156-168 .

King J.A. 1988. Some physical features of the soil after opencast mining . Soil Use and Management, 4, 23-30.

King J. A. and Bullard M. J. 2001. Carbon sequestration options in the UK – a case study in Norfolk, with particular emphasis on the role of energy crops. Aspects of Biology 65, Biomass and energy crops II , 295-302.

King, J. A., Bradley, R. I., Harrison, R. and Carter, A. D. 2004. Carbon sequestration and saving potential associated with changes to the management of agricultural soils in England. Soil Use and Management 20, 394-402.

Page 92 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

King, J. A., Bradley, R. I. and Harrison, R. 2005. Current trends of soil organic carbon in English arable soils. Soil Use and Management 21, 189-195.

King, D. A. & Sinden, J. A. 1988. `Influence of Soil Conservation on Farm Land Values'. Land Economics, 64(3): 242-255.

Kirkegaard, J.A., So, H.B. and Troedson, R. J. 1993 Effect of compaction on the growth of pigeonpea on clay soils. III. Effect of soil type and water regime on plant response. Soil and Tillage Research, 26, 2, 163-178.

Lancaster, K. J. 1966. A new approach to consumer theory. Journal of Political Economy 74, 132-157.

Lloyd, T. A. Rayner, A. J. and Orme, C. D. 1991. Present value models of land prices in England and Wales. European Review of Agricultural Economics 18,141-166. Loomis, R. S. and Connor, D. J. 1996 Crop ecology: productivity and management in agricultural systems. Cambridge University Press.

Loveland P. J. and Thompson T. R. E. 2002. Identification and development of a set of national indicators for soil quality. R&D Project Record P5-053/PR/02. Pub. Soil Survey and Land Research centre, Silsoe, UK. http://www.defra.gov.uk/science/project_data/DocumentLibrary/SP0512/SP0512_432 _FRP.pdf

Lumby, S. 1988. Investment Appraisal and Financial Decisions . Van Nostrand Reinold.

Lussier, G. R. Baker, L. and Thomassin, P. J. 2001. Implicit prices for resource quality investments in Quebec’s agricultural land market. Canadian journal of Regional Science , 24, 295-311.

Macmillan, D. C. 2003. The Value of Managing Forest to Protect Archaeology. In ‘ The Social and Environmental Benefits of Forests in Great Britain’. Report to Forestry Commission, Edinburgh. Centre for Environmental Appraisal and Management, University of Newcastle upon Tyne.

Macphail, R. I. and Courty, M. A. 1985 Interpretation and significance of urban deposits. Edgren I. and Junger H. (Eds.). Proceedings of the Third Nordic Conference on the Application of Scientific Methods in Archaeology. ISKOS 5. Finnish Antiquarian Society, Helsinki, 71-83.

MAFF 1999. FCDPAG3 Flood and Coastal Defence Project Appraisal Guidance: Economic Appraisal . MAFF.

MAFF 2001a. FCDPAG1 Flood and Coastal Defence Project Appraisal Guidance; Overview MAFF.

MAFF 2001b. FCDPAG3 Flood and Coastal Defence Project Appraisal Guidance; Economic Appraisal. MAFF.

MAFF 2000a Fertiliser Recommendations for Agricultural and Horticultural Crops . Reference Book 209, 7 th Edition. Ministry of Agriculture, Fisheries and Food. The Stationery Office, London, 177pp.

MAFF 2000b Good Practice for Handling Soils. MAFF & Crown Copyright: Contains fifteen Guidance Sheets on soil stripping and handling and four Guidance Sheets on remedial works.

Page 93 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Maddison, D. 2000. A hedonic analysis of agricultural land prices in England and Wales. European Review of Agricultural Economics 27, 519-532.

Marshall D. C. W. and Bayliss A. C. 1994. Flood estimation for small catchments. Institute of Hydrology, Report No. 124. Publ.. Institute of Hydrology, Wallingford, Oxon., UK. 1994. 73pp.

McGrath, S. P. 1984 Metal concentrations in sludges and soil from a long-term field trial. Journal of Agricultural Science, Cambridge 103, 25-35.

McGrath, S. P., Giller, K.E. and Brookes, P.C. 1987 Rothamsted Experimental Station Report for 1986.

Miranowski, J.A., and B. D. Hammes 1984. Implicit Prices of Soil Characteristics for Farmland in Iowa. American Journal of Agricultural Economics 66, 745-49.

NAEI 2005. National Atmospheric Emissions Inventory. www.naei.org.uk/index.php .

NEGTAP 2001 Transboundary air pollution, acidification, eutrophication and ground- level ozone in the UK. National Expert Group on Transboundary Air Pollution, Defra, London.

Nunes, P.A.L.D., J.C.J.M. van den Bergh, 2001. "Economic Valuation of Biodiversity: Sense or Nonsense" , Ecological Economics 39, no. 2 203-222.

OECD (2006)Palmquist, R. 1989). Land as a differentiated factor of production: a hedonic model and its implications for welfare measurement. Land Economics 65, 23- 28.

Palmquist, R. B., and L.E. Danielson 1989. A Hedonic Study of the Effects of Erosion Control and Drainage on Farmland Values American Journal of Agricultural Economics 71, 943-46.

Pearce, David 2005a. The Social Cost of Carbon, in Dieter Helm (ed.) Climate Change Policy. Oxford University Press, Oxford.

Pearce, D.W. 2005b. “Does European Union environmental policy pass a cost-benefit test?”. World Economics 5, 115-138.

Pearce, David, Atkinson, Giles and Mourato, Susana 2006. Cost-Benefit Analysis and the Environment. OECD , Paris.

Penning-Rowsell E and Fordham M (eds.) 1984. Floods Across Europe: Flood Hazard Assessment, Modelling and Management. London: Middlesex University Press

Potash Development Association 2005 Potash for Cereals. PDA leaflet 11, York. 15pp.

Powe, N. A., G. D. Garrod and K. G. Willis 1995. Valuation of urban amenities using an hedonic price model. Journal of Property Research 12, 137-147.

Pretty J.N., Brett C., Gee D., Hine R. E., Mason C.F., Morison J.I.L., Raven H., Rayment M.D. and van der Bijl G. 2000. An assessment of the total external costs of UK agriculture. Agricultural Systems 65, 113-136.

Renwick, A., Revell, B. and Hodge, I. 2003. Preliminary Analysis of Partial Decoupling . Report for Defra.

Page 94 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Report of Field Trials: April 1983-March 1986 (SDA 9166 SLD). WRc Report No. 1047-M/2. WRc Medmenham, Marlow.

Ribaudo, M. O., D. Colacicco, A. Barbarika and C. E. Young, 1989 The Economic Efficiency of Voluntary Soil Conservation Programs , Journal of Soil and Water Conservation, Jan.-Feb.

Ribaudo, Marc O. and C. Edwin Young, 1989. Estimating the Water Quality Benefits from Soil Erosion Control. Water Resources Bulletin Vol. 25, no. 1, 71-78.

Rickman R., Douglas C., Albrecht S. & Berc J. 2002. Tillage, crop rotation, and organic amendment effect on changes in soil organic matter. Environmental Pollution, 116, 405-411.

Roberts, D., MacDonald, D. Kampas, T., Shannon, P, Potts, J. and Barroclough, F. (2002). Nature Conservation Designations and Land Values . Scottish Executive Central Research Unit, Edinburgh.

Roka, F. and Palmquist, R. 1997. Examining the use of national databases in a hedonic analysis of regional farmland values. American Journal of Agricultural Economics , 75, 1651-1656.

Rural Development Service 2005. Environmental Stewardship . Rural Development Service, Defra.

Russell, E.W. 1973 Soil Conditions and Plant Growth 10 th Edition, Longmans, London, 89pp.

Saunders, C. 1996. Financial, Public Exchequer and Social Value of Changes in Agricultural Output . University of Newcastle upon Tyne, Department of Agricultural Economics and Marketing Working Paper 20.

Schipper L. A. and McLeod M. 2002. Subsidence rates and carbon loss in peat soils following conversion to pasture in the Waikato Region, New Zealand. Soil Use and Management, 18, 91-93.

Skinner, R. J., Church, B.M. and Kershaw, C.D. 1992 Recent trends in soil pH and nutrient status in England and Wales. Soil Use and Management 14, 162-169.

Smith, C.V. 1977 Work days from weather data. The Agricultural Engineer. 32, 95-97.

Smith, S.R. 1996 Agricultural Recycling of Sewage Sludge and the Environment. CAB International, Wallingford. pp. 365.

Smith P., Powlson D. S., Glendining M. J. and Smith J. U. 1997. Potential for carbon sequestration in European soils: preliminary estimates for five scenarios using results from long-term experiments. Global Change Biology, 3, 67-79.

Smith P., Powlson D. S., Smith J. U., Falloon P. and Coleman K. 2000a. Meeting the UK’s climate change commitments: options for carbon mitigation on agricultural land. Soil Use and Management , 16, 1-11.

Smith P., Goulding K. W. T., Smith K. A., Powlson D. S., Smith J. U., Falloon P. and Coleman K. 2000b. Including trace gas fluxes in estimates of the carbon mitigation potential of UK agricultural land. Soil Use and Management , 16, 251-259.

Page 95 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Smith P., Milne R., Powlson D. S., Smith J. U., Falloon P. and Coleman K. 2000c. Revised estimates of the carbon mitigation potential of UK agricultural land. Soil Use and Management , 16, 293-295.

Soane, B. D., Godwin, R. J. Marks, M.J. & Spoor, G. 1987. Crop and soil responses to deep incorporation of phosphorus and potassium fertilizers and subsequent soil management on a range of soil types. Part I: Response to arable crops. Soil Use and Management 3, 3, 123 - 130.

Spurgeon, J. Colarullo, G. Radford, A and Tingley, D. 2001. Economic Evaluation of Inland Fisheries Module B: Indirect Economic Values Associated with Fisheries . R and D project Record, W2-039/PR/2. Report for the Environment Agency.

Stewart, L. Hanley, N. and Simpson, I. 1997. Economic valuation of the agri- environment schemes in the United Kingdom. Unpublished Report by the Environmental Economics Research group, University of Stirling, to HM Treasury and MAFF.

Tisdall, J.M. & Oades, J.M. 1982. Organic matter and water-stable aggregates in soils. Journal of Soil Science , 33, 141-163.Tol, Richard S. J. (2002a). Estimates of the damage costs of climate change: Part 1: benchmark estimates. Environmental and Resource Economics 21, 47-73.

Tol, Richard S. J. 2002a. Estimates of the damage costs of climate change: Part 1: benchmark estimates. Environmental and Resource Economics 21, 47-73.

Tol, Richard S. J. 2002b. Estimates of the damage costs of climate change: Part 2: dynamic estimates. Environmental and Resource Economics 21, 135-160.

Tol, Richard S. J. 2005. The marginal damage costs of carbon dioxide emissions: an assessment of the uncertainties. Environmental and Resource Economics 21, 135- 160.

Traill, W. B. 1979. An empirical model of the UK land market and the impact of price support policy on land values and rents. European Review of Agricultural Economics , 6, 209-232.

Li, Y.X, Tullberg, J.N., and Freebairn, D. M. 2006 Wheel traffic and tillage effects on runoff and crop yield . Soil and Tillage Research, In Press, Corrected Proof, Available online 2 February 2006 .

Vigerust, E. and Selmer-Olsen, A.R. 1986 Basis of metal limits relevant to sludge utilisation. In: Davis, R. D., Haeni, H. and L’Hermite, P. (eds.), Factors Influencing Sludge Utilization Practices in Europe. Elsevier Applied Science Publishers Ltd., Barking. Pp. 26-42. Watson, C.A., Atkinson, D., Gosling, P., Jackson, L.R. & Rayns, F. W. 2002 Managing soil fertility in organic farming systems. Soil Use and Management, 18: 239-247.

Weersink, A., Clark, S., Turvey, C. G. and Sarker, R. 1999. The effect of agricultural policy on farmland values. Land Economics 75, 425-430.

Whitby, M. and Adger, N. 1996. Natural and reproducible capital and the sustainability of land use. Journal of Agricultural Economics 47, 50-65.

Willis K. G. 1990. Valuing Non-Market Wildlife Commodities: an evaluation and comparison of benefits and costs. Applied Economics 22, 13-30.

Page 96 Economic Valuation of Soil Functions Phase 1: Literature Review and Method Development

Willis, K. G. and Garrod, G. D. 1992. Amenity value of forests in Great Britain and its impact on the internal rate of return from forestry. Forestry 65 (3), 331-346.

Willis, K. G., Garrod, G.D. and Saunders, C. M. 1993. Valuation of the South Downs and Somerset Levels and Moors Environmentally Sensitive Area Landscapes by the General Public . A Report to the Ministry of Agriculture, Fisheries and Food. Centre for Rural Economy, Department of Agricultural Economics and Food Marketing. University of Newcastle upon Tyne.

Withers, P.J.A., Unwin, R. J., Grylls, J. P. and Kane, R. 1994 Effects of withholding phosphate and potash fertilizer on grain yield of cereals and on plant-available phosphorus and potassium in calcareous soils. European Journal of Agronomy 1, 1- 96.

Page 97