An ecosystem services approach to the quantification ...

Ecosystem Services 9 (2014) 204–215

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Ecosystem Services journal homepage: www.elsevier.com/locate/ecoser

An ecosystem services approach to the quantification of shallow mass movement erosion and the value of soil conservation practices E.J. Dominati a,n, A. Mackay a, B. Lynch b,1, N. Heath b,1, I. Millner b,1 a b

AgResearch, Grasslands Research Centre, Tennent Drive, Private Bag 11008, Palmerston North 4442, New Zealand Hawkes Bay Regional Council, 159 Dalton Street, Private Bag 6006, Napier 4142, New Zealand

art ic l e i nf o

a b s t r a c t

Article history: Received 28 March 2014 Received in revised form 25 June 2014 Accepted 26 June 2014 Available online 5 August 2014

This study characterises the loss of ecosystem services from a grazed pasture following shallow mass movement erosion and subsequent recovery of services. The influence of space-planted trees, a soil conservation practice, on the provision of services, was also assessed. The economic value of the services provided by an uneroded steep pasture grazed by sheep and cattle was estimated at NZD 3717 ha 1 yr 1. This value dropped by 65% when the topsoil was lost in a single shallow mass movement. Fifty years after erosion, the services only recovered to 61% of uneroded value. In contrast, the same landscape type planted with soil conservation trees provided, after 20 years, additional ( þ22% in dollar value) services from the similar unprotected landscape. A benefit cost analysis of soil conservation practices showed planting conservation trees is only profitable if the trees are harvested for timber (age 20), and low discount rates ( o5%) are used. When the economic value of the extra services from conservation trees is included in the BCA, the Net Present Value of the investment is greatly positive at discount rates ranging from 0% to 10%. Analysis of this ecological infrastructure investment using an ecosystem service approach offers new insights for resource managers and policy makers. & 2014 Elsevier B.V. All rights reserved.

Keywords: Soil change Ecosystem services Natural capital Erosion recovery Soil conservation Benefit cost analysis

1. Introduction New Zealand is one of the only places in the world where hill country is grazed all year round. Almost 70% of the country has slopes greater than 12 deg., and is commonly called ‘hill country’. Geologically, the majority of the North Island, where this study takes place, is developed on soft rock and crushed soft rock terrain (McIvor et al., 2011). These combined with warm sub-tropical to cool temperate climates make the North Island hill country highly prone to shallow landslides, earthflow and gully erosion (Basher et al., 2008; McIvor et al., 2011). Nowadays, historical deforestation associated with pastoral land use, also plays a major role in determining erosion risk. The Water and Soil Conservation Act was passed in New Zealand in 1941 to address hill country erosion associated with post-European settlement and deforestation. Catchment Boards, directed by central government policies, were tasked with soil and water conservation until 1988. In 1988, Catchment Boards were absorbed into Regional or Unitary Councils responsible for broader

n

Corresponding author. Tel: þ 64 63518216. E-mail addresses: [email protected] (E.J. Dominati), [email protected] (A. Mackay), [email protected] (B. Lynch), [email protected] (N. Heath), [email protected] (I. Millner). 1 Tel.: þ64 68359209. http://dx.doi.org/10.1016/j.ecoser.2014.06.006 2212-0416/& 2014 Elsevier B.V. All rights reserved.

natural resource management, including soil erosion and flood control under the Resource Management Act (RMA) of 1991. Each year hill country erosion is estimated to cost between NZD 100 to 150 million (Eastwood et al., 2001). Part of this is through lost pasture production and nutrients (MfE, 2007), but does not include an estimate of the loss of soil natural capital stocks (Dominati et al., 2010). The investment in soil conservation continues today, as erosion remains a challenge threatening the long-term sustainability of agro-ecosystems. This is not unique to New Zealand but a threat to food security in many regions of the world (McBratney et al., 2014), heightened by uncertainties surrounding future climates. Soil conservation practices aim to reduce the risk of soil erosion in hill and steep land country, downstream costs associated with sediment loadings in waterways, and damage to productive farmland and towns through siltation. In New Zealand, erosion control measures for pastoral hill country are established in the presence of the grazing animal, with permanent retirement from grazing recommended only in the most extreme situations. Tree-based control measures, which stabilise mass flows, are most of the time, the only affordable option on the scale required. Poplar (Populus spp.) and willow (Salix spp.) are two of the most suitable and most used tree species (McIvor et al., 2011). Current evaluation of soil conservation policies are largely limited to the assessment of the reduction in soil erosion, soil loss, sediment, impacts on productive capacity and downstream

E.J. Dominati et al. / Ecosystem Services 9 (2014) 204–215

community. However, evaluation needs to be broader in order to capture the wider benefits of soil conservation. For example the value of the full range of ecosystem services (ES), above- and below-ground, lost following shallow mass movements need to be considered, beyond the loss of selected provisioning services. Until the full range of services is considered in the analysis, the cost of land degradation and loss of natural capital stocks will remain unknown and the full value of an ecological infrastructure investment in soil conservation practices will not be available for land use decision-making. The purpose of this study is to test and evaluate an ecosystem services approach for the quantification and economic valuation of the multiple benefits, provided by soil conservation practices using a study area on the East Coast of the North Island of New Zealand. In April 2011, Hawke's Bay on the East Coast of the North Island was affected by a heavy rain storm (200 to 650 mm of rain in 12 hours). Subsequent widespread shallow mass movement erosion occured on hill slopes, including slips and reactivated earth flows, as well as gullying, along a 250 km coastal strip predominantly under permanent pasture grazed by sheep and cattle. Following that storm, Hawke's Bay Regional Council used satellite imagery to estimate the proportion of land affected by landslides. Overall 43 km2 (4300 ha) of bare ground was classified from a total area of 5900 km2, including 86% new bare ground resulting from the storm (Jones et al., 2011). Estimates of damage to infrastructure and land, personal and commercial property was NZD 39 million. As part of a wider analysis, Hawke's Bay Regional Council was interested in investigating the long-term implications of the storm event on the region's natural resource base. This provided an opportunity to quantify and value the ecosystem services lost as a consequence of shallow mass movement erosion and at the same time the benefits of existing soil conservation practices. While this study produces economic values for ecosystem services, the main purpose of the research was to explore the merits of an ecosystem service approach in providing resource management decision makers with new insight into the long-term consequences of erosion and the costs and benefits of an ecological infrastructure investment. The objective was to determine if an ecosystem services approach offers a new tool for amending existing and shaping future policy. In order to look at the impacts of shallow mass movement erosion and soil conservation practices, the contribution of soils to the provision of all ecosystem services was assessed using the theoretical natural capital-ecosystem service framework of Dominati et al., (2010). That framework builds on the millennium ecosystem assessment (MEA, 2005), and differentiates ecological processes, particularly soil processes, from the flow of ecosystem services, and integrates the relationships between natural capital stocks, the impacts of external drivers, and the provision of ecosystem services with human needs. Within the context of this study, the framework is used to account for below and above ground contributions of natural capital stocks to the provision of ecosystem services. The ecosystem services considered include provisioning services such as the provision of food (quantity and quality), wood and fibre, the provision of support for human infrastructures and farm animals, and the provision of shade and shelter for livestock. The regulating services include flood mitigation, the filtering of nutrients and contaminants, the decomposition of wastes, net carbon accumulation in soils and conservation trees, nitrous oxide regulation, methane oxidation and the regulation of pest and disease populations. Cultural services (the nonmaterial benefits people obtain from ecosystems) (MEA, 2005) are recognised but not considered in this study. Their non-biophysical nature, requires the use of very different techniques for quantification and valuation and as such are outside the scope of this study.

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Table 1 Farm size, landscape type, soil classification and properties, and nutrient inputs to a sheep and beef operation.

Area (ha) Slope class Slope Relative productivity NZ soil classification Order Group NZ Soil Series US soil classification Order Suborder Olsen P (mg/kg) Anion storage Capacity (%) N fertiliser applied (kgN/ha/yr) P fertiliser applied (kgP/ha/yr)

Rolling landscape unit

Steep landscape unit

255 (45%) Easy hill 16–251 1.6

315 (55%) Steep hill 4261 1

Brown Orthic Waimarama sandy loam

Pallic Immature Wanstead clay loam

Inceptisols Dystrudepts 25 43 20

Inceptisols Eutrudepts 16 21 0

20

15

2. Methodology 2.1. Study site: East Coast hill country sheep and beef operation The dominant land use in Hawke's Bay region on the East Coast of the North Island of New Zealand is sheep and beef farming. Average farm monitoring data (MPI, 2012) for a summer-dry hill country breeding and semi-finishing sheep and beef operation were used for this study (Table 1). The farm characteristics include a 70:30 sheep to cattle ratio, 130% lambing rate, stocking rate of 10 stock units2 per ha, and pasture growth of 9 t of dry matter/ha/yr. Average yearly Rainfall is 1000 mm, and the climate is described as summer-dry. Soil and landscape information was provided by Hawke's Bay Regional Council (Table 1). The farm has two dominant landscape units described as the rolling landscape unit, being flat to easy rolling country, and the steep landscape unit, being moderate to steep hill country. To answer the study objectives, the following steps were taken:

Quantify and value the provision of ecosystem services for a sheep and beef operation (Section 2.3).

Quantify the impact of erosion on soil properties, the flow of ecosystem services and their recovery (Section 2.4).

Quantify and value the impact of soil conservation practices on

the flow of ecosystem services from a sheep and beef farm (Section 2.5). Undertake a benefit-cost analysis of an investment in soil conservation (Section 2.7).

For all the steps mentioned above, the provision of ecosystem services is calculated at the paddock scale (the ecosystem services source area) on a 1 ha basis. To quantify the provision of ecosystem services at the paddock scale (1 ha), information from existing data bases and tools that support existing planning in the region, including data collected as part of soil quality monitoring, land use capability class maps (Lynn et al., 2009) and the OVERSEERs nutrient budget model 2 A livestock unit (SU) is the feed requirement used as the basis of comparison for different classes and species of stock. It expresses the annual feed requirements, equivalent to one 55 kg ewe rearing a single lamb. 1 SU requires approximately 520 kg of good quality pasture dry matter per year.

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(version 6.0) (Wheeler et al., 2008) were used to generate the required data.

2.2. General principles for quantifying and valuing ecosystem services The methodology used for the quantification and economic valuation of ecosystem services follows the six guiding principles from Dominati et al., (2014):

The quantification of ecosystem services needs to be very

specific to the benefits directly useful to humans, not to be confused with the processes underlying ecosystem functioning and the supporting processes behind the formation and maintenance of natural capital stocks. To determine how ecosystems provide ecosystem services, the properties and processes at the origin of the provision of each service need to be described and quantified. Differentiate the contributions of the natural capital and added or built capital (e.g. infrastructures, inputs such as fertilisers or irrigation water) to the flow of ecosystem services when defining proxies to quantify each service. Identify where and how external drivers such as climate and land use impact on the provision of services through their impact on both natural capital stocks and processes. Analyse the impact of degradation processes, in this case erosion, in addition to the effect of soil conservation measures, on natural capital stocks and thereby the flows of ecosystem services. Base the economic valuation on the bio-physical measures of the services that are relevant for the chosen scale and land-use.

These principles align with the model described by DEFRA (2007) and used by Posthumus et al. (2013).

2.4. Impact of shallow mass movement erosion on soil properties, the flow of ecosystem services and recovery The main forms of erosion in New Zealand are (Cairns et al., 2001; Lynn et al., 2009):

Mass movement erosion – which occurs when heavy rain or

The erosion form addressed here which occurred in Hawke's Bay region in April 2011, is mass movement erosion. On unstable slopes, thousands of landslides can be triggered by high-magnitude/low-frequency storms with estimated return periods exceeding 50 years (McIvor et al., 2011). In New Zealand, the distinction is made between shallow (in the soil, easily stabilised) and deep (in underlying regolith, where stabilisation is difficult) mass movements (Crozier, 1986; Cairns et al., 2001; Lynn et al., 2009). Four forms of shallow mass movement are common in the New Zealand landscape:

Slips – These are shallow landslides in soil or weathered

2.3. Quantification of the provision of ecosystem services for a sheep and beef operation The quantification and economic valuation of ecosystem services was realised at the paddock scale (1 ha) for each of the two landscape units (rolling and steep) of the hill-country sheep and beef operation studied. For each service, a proxy to measure the service was determined based on ecosystem services definitions and Dominati et al. (2010) framework. The data needed to calculate the proxy was generated from the OVERSEERs nutrient budget model (version 6.0) (Wheeler et al., 2008) or sourced from the literature. Changes in the value of each proxy over time were based on published research results when available, or assumptions. The economic valuation was then based on the measure of the proxy for each service. Although, ‘precise’ economic values are presented as results here, we recommend treating them as ‘orders of magnitude’. The detail of the steps in the quantification and economic valuation for each of the provisioning and regulating ecosystem services are presented in Table 2. Definitions of each of the provisioning and regulating services are given by Dominati et al. (2010). The reader can also refer to Dominati et al., (2014) which presents in details the methodology applied to a dairy grazed system. In this study, where research knowledge and data were lacking, assumptions were made regarding changes to the measure of each service. Assumptions are listed in Table 4. The cultural services associated with a sheep and beef operation were not considered in this study.

earthquakes cause whole slopes to slump, slip or landslide. Storms are the primary triggers. This is the most common form of erosion in the hill country. Fluvial erosion – which occurs when running water digs shallow channels or deeper gullies into the soil. Surface erosion – which occurs when wind, rain or frost detach soil particles from the surface, allowing them to be washed or blown off the paddock (occurs largely outside the hill country.) Sediment erosion – activities involved in earthworks, plantation forests, cropland and pasture management may result in sediment loads being mobilised and entering watercourses.

regolith on steep slopes or on low-angle slopes where regolith is susceptible to rupture. Earth flows – These are shallow flows on low-angle slopes, where regolith is susceptible to plastic deformation, or in colluvium (slip debris) on foot slopes. Debris avalanches—shallow, rapid landslides in regolith on upper mountain slopes. Debris flows – shallow, rapid flows in colluvium (avalanche debris) on lower mountain slopes.

In this study no distinction was made between the different types of shallow mass movements; therefore when the term landslide is used, we refer to shallow mass movement erosion. Future studies could examine in more detail the impact difference types of erosion have on soil properties and recovery. 2.4.1. Loss of ecosystem services from erosion The first step in quantifying the loss of ecosystem services from an erosion scar and subsequent recovery was establishing the state of the soil natural capital stocks after a shallow landslide. Existing research on the soil properties of erosion scars and soil recovery after landslides (Rosser and Ross, 2011a; Rosser and Ross, 2011b), as well as assumptions made by the authors, were used to describe the status of the natural capital stocks remaining after a shallow mass movement, and how that, and the provision of ecosystem services from those natural capital stocks, recover in subsequent years. Recovering services were re-quantified and re-valued using the methods presented in the previous section (Table 2). The quantification and valuation of ecosystem services was recalculated for a paddock in a steep landscape unit immediately after a shallow mass movement. It was assumed the soil displaced during the shallow mass movement was deposited elsewhere, e.g. downslope covering existing pasture, and therefore still able to provide services. It would be very interesting to consider the


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Table 2 Detailed description, quantification, proxy and economic valuation of provisioning and regulating ecosystem services.

Table 3 Details of the additional ecosystem services provided by wide-spaced trees to a grazed pasture system, their quantification, proxy and economic valuation. Services

What to measure

Parameters/indicators used to quantify the service

Feed Quantity Tree

Tree growth

Amount of foliage animals can eat, sustainable harvest (literature)

Wood- Fibre

Tree growth

Provision of shade to animals Provision of shelter to animals Net carbon accumulation (tree)

Impact of shade on animal growth Impact of shelter on young animal survival rates Net C flows

Formula

ES measure

Amount of foliage animals can eat/year Equivalent in pasture dry matter Amount of wood grown (literature) Amount of wood harvested Wood kg/ha after 20–30 years Grazing time / day (literature) Growth with increased dry matter kg meat/ha/yr utilisation with shade normal growth Lamb and calf losses (literature)

C stocks and variations (literature)

Number of young animals with shelter normal number of young animals Net C accumulation in wood

Valuation method used at the catchment scale Pasture dry matter market price Wood market price Meat market price

kg meat/ha/yr

Meat market price

t C/ha/yr

Market prices of C

208


Table 4 Assumptions made for the base analysis and sensitivity analysis. Ecosystem services

Soil recovery–base assumptions (% of pasture baseline)

Soil recovery–sensitivity Wide spaced trees-base analysis assumption (% of pasture baseline) (% of pasture baseline)

Years following erosion

Years following erosion

Age of trees

0 Food quantity pasture (%) 0 Food quantity tree (%) NA Food quality-pasture (%) 0 Wood-fibre (%) NA Provision of support for human infrastructures (%) 0 Provision of support for farm animals (%) 0 Provision of shade (%) NA Provision of shelter (%) NA Flood mitigation (%) 30 Filtering of nutrients and contaminants (%) 30 Decomposition of wastes (%) 5 Net carbon accumulation (soil) (%) 0 Net carbon accumulation (tree) (%) NA Nitrous oxide regulation (%) 0 Methane oxidation (%) 0 Regulation of pest and disease populations (%) 113

Wide spaced trees-sensitivity analysis (% of pasture baseline) (% of initial assumption) Age of trees

20

50

0

20

50

10 years

20 years

10 years

20 years

60 NA 60 NA 0 50 NA NA 40 40 60 1000 NA 40 40 100

80 NA 80 NA 0 50 NA NA 50 50 80 83 NA 50 50 100

0 NA 0 NA 0 0 NA NA 15 15 3 0 NA 0 0 50

30 NA 30 NA 0 25 NA NA 20 20 30 500 NA 20 20 50

40 NA 40 NA

85

75

100

100

0 100

0 100

110 110 105 60

130 130 110 60

240 200 100

240 200 100

70 50 100 50 0 100 50 50 100 100 100 30 50 100 100 100

50 50 100 50 0 100 50 50 100 100 100 30 50 100 100 100

25 NA NA 25 25 40 42 NA 25 25 50

NA: Not applicable

provision of ecosystem services from such areas which potentially present improved natural capital after the landslide. However, this was beyond the scope and boundaries of the present study. It was assumed that immediately after the landslide (in the first year), soil would not grow pasture, support infrastructures or animals, accumulate carbon (C), or regulate greenhouse gases (GHG) (Rosser and Ross, 2011b). Therefore, the provision of those services was assumed to be nil (Table 4). After a shallow landslide water storage by the soil profile is reduced by 70% (Rosser and Ross, 2011b; Stavi and Lal, 2011). It was also assumed that the capacity of bare ground to filter nutrients and contaminants decreased by 70%. On bare ground, it was assumed than only 5% of the dung deposited was decomposed properly. Since no pasture would grow on bare ground immediately after a shallow landslide, the provision of regulation of pest and disease populations was assumed to be maximal (Table 4).

2.4.2. Recovery of the ecosystem services after erosion A number of studies have quantified the recovery rates of some topsoil properties and pasture characteristics following mass movement erosion (Lambert et al., 1984; Sparling et al., 2003; Rosser and Ross, 2011b). This data was used to calculate the recovery of ecosystem services provision as the bare ground weathered and other pedological processes initiate soil development under a pastoral vegetation cover. It has been shown that after a shallow landslide, pasture recovery beyond 80% of uneroded level is unlikely even after 50 years (Lambert et al., 1984; Rosser and Ross, 2011b). It was assumed here that the provision of food quantity (pasture growth) and quality would recover linearly to 60% after 20 years, 80% after 50 years and then plateau (Table 4). The loss of depth of soil profile during a shallow landslide leads to reduced drainage and water holding capacity. As mentioned earlier, some of the soil material would be deposited elsewhere, but this was not considered in this study. This means landslides scars are likely to be wetter for extended periods limiting the support service. It was assumed that the support to animals would recover up to 50% in 20 years and then further improvements would be very slow. Similarly, flood mitigation and the filtering of nutrients and contaminants are highly-dependent on the depth of

the soil profile and the nutrient status of the topsoil. It was assumed that these services will recover up to 40% in 20 years, 50% beyond 50 years, with further recovery tied to pedological time scales. Microbiological activity in the new topsoil recovers up to 80% of the original topsoil in around 30 years (Sparling et al., 2003). To simulate the recovery of the decomposition of wastes, we assumed a recovery of 60% after 20 years, and 80% after 50 years (Table 4). The new top soil being formed accumulates C faster than uneroded soils (Lambert et al., 1984; Sparling et al., 2003; Page et al., 2004) even though total C levels may never recover to the uneroded top soil levels within human lifetimes (Sparling et al., 2003; Rosser and Ross, 2011b). It was assumed that net C accumulation rates were 10 times faster than uneroded levels in the first 10 years following a shallow landsliding, decreasing overtime, leading to total C levels in the new topsoil reaching 80% of uneroded levels after 45 years (Sparling et al., 2003; Page et al., 2004). The regulation of greenhouse gas (GHG) emissions also depends on drainage and nutrients status, so it was therefore assumed that this service will recover up to 50% in 20 years (Table 4). As pasture re-establishes on the landslide scar, and topsoil starts forming and accumulating, pests also return. It was assumed that, like a newly sown pasture, initial infestation rate will be high between year 2 and 5 after the landslide, before declining to the levels found on an uneroded site. 2.5. Impact of soil conservation practices on the flow of ecosystem services: Options for reducing the risk of shallow mass movements in hills country usually are (Crozier, 1986; Cairns et al., 2001; Lynn et al., 2009)

Spaced planting of trees in pasture, to provide the soil with root reinforcement, or

Land use change from grazing to commercial timber trees, or Reversion to native scrub cover. The first option is examined in this study. The ability of wide-spaced tree planting to reduce the occurrence of shallow landslides is well understood. Douglas et al., (2011) showed that


wide-spaced trees can reduce landslide occurrence by 95%, compared to pasture only landscapes. It was assumed in the examination of the benefits of wide-spaced tree planting (50 stems per hectare, 30% canopy cover) on the provision of ecosystem services from soils under a pastoral use, that once the trees are planted, no more landslides occur for the next 20 years, on 100% of the area covered. In reality the trees will have limited utility against erosion until 5–10 years of age. Additional services from trees, that would otherwise not be available in a grassland system, were included in the analysis (Table 3). 2.5.1. Changes to ecosystem services provision under trees Data from the literature were used to quantify the influence the wide-spaced planted trees had on the provision of ecosystem services from the pasture soil. The following assumptions were made. Wide-spaced trees reduce pasture production by 15% when trees are aged 5–10 years and by 25% when trees are aged 10–20 years and over (Guevara-Escobar et al., 2007; McIvor and Douglas, 2012). The trees have no influence on pasture quality (Table 4). Trees intercept 5–35% of rainfall (Guevara-Escobar et al., 1998; Benavides et al., 2009) depending on tree density and crown size, therefore we assumed that runoff was decreased by 10% under young trees (o10 years) and 30% under older trees ( 420 years). Increased water uptake by pasture in combination with trees and improved soil physical characteristics (Benavides et al., 2009) were assumed to improve the soil's filtering capacity by 10% under young trees ( o10 years) and 30% under older trees ( 420 years) (Table 4). Increase in carbon (C) and nitrogen (N) mineralization (Benavides et al., 2009) and increased soil pH and exchangeable cations (Ca, K and Mg) (Guevara-Escobar et al., 2002) were identified under wide-spaced trees compared to open pastures. Therefore we assumed the decomposition of wastes was 5% greater under young trees (o10 years) and 10% greater under older trees (420 years). Soil C and N content (top 10 cm) are greater in open pasture than in a poplar-pasture system (Guevara-Escobar et al., 2002; Benavides et al., 2009); therefore the net C accumulated in soils by a pasture-tree system was assumed to be 60% of open pasture (Table 4). Since pasture-tree systems are usually better drained (Benavides et al., 2009) it was assumed that N2O emissions would be 50% lower and CH4 oxidation would double compared to open pasture. The level of regulation of soil pest and diseases population was assumed to be the same for open pasture or under poplars (Table 4). 2.5.2. Additional ecosystem services from trees Conservation trees are an ecological infrastructure investment (Jury et al., 2011) that provides additional support and resilience to the pastoral ecosystem. Both McGregor et al. (1999) and Parminter et al. (2001) identified several benefits from introducing widespaced-planted trees into a pasture system, including shelter from extreme events for farm animals, shade throughout the year, soil stability, vista, food source for native birds, wood fibre and an alternate forage source for grazing animals during summer dry periods, which coincides with mating and hence provides insurance against decline in ovulation rates (Orsborn et al., 2003). These services were added to the list above (Table 3). Pollarding willow and poplar in the summer months can provide a supplementary fodder source in summer-dry hill country (Douglas et al., 2006). It was assumed here that trees could be used for forage when aged 5 years or older. The amount of forage produced was calculated for 10 and 20 year old trees assuming they have not been pollarded previously. The value of the available forage was added as a service. Timber is another service that trees provide as they can be harvested for their wood. The provision of

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shelter and shade for animals is an emerging reason for greater tree-planting on pastoral farms. Trees are an option for reducing the risk of stock losses from extreme climatic events and moderating extended weather extremes (e.g. temperature) to protect capital stock and sustain animal growth rates. The availability of shade has been shown to prolong animal grazing each day (Betteridge et al., 2012), so an increased dry matter utilisation was used to measure these services. Shelter has the potential to make a significant difference to lamb and calves survival rates and initial growth rates in the spring months, so it was assumed that lamb and calf losses were reduced by 25% between scanning and weaning (Parminter et al., 2001) for trees older than 5 years and 50% for trees older than 10 years. Established soil conservation trees provide an opportunity to claim carbon (C) credits under the New Zealand Emissions Trading Scheme. Net C accumulated in poplar trees was quantified (Table 3). The influence of a treepasture system on the GHG regulation and the regulation of above-ground pest and disease populations of the system is still poorly understood so we assumed here that there were no immediate additional benefits other than the ones considered in soils in the section above. 2.6. Sensitivity analysis A sensitivity analysis was included to investigate the influence the assumptions made in the quantification of the ecosystem services and calculation of economic value of the services had on the results and subsequent interpretations. Two separate sensitivity analysis were conducted. 1. The recovery of the ecosystem services after an erosion event was assumed to be only half of the original recovery rates (Table 4). The implications of this to the cost of erosion were then recalculated. 2. The wide-spaced trees only grow at half the rate and produce only half of the additional ecosystem services (Table 4). The implications of the benefit cost analysis of a soil conversation investment were also recalculated.

2.7. Economic and financial analysis To determine the economic value of each service at the farm scale neo-classical economic valuation techniques, including market prices, defensive expenditures, replacement cost and provision cost, were used (Tables 2 and 3). A mixture of valuation methods had to be used as many services, such as regulating services, are not traded in markets and therefore lack direct market values. Basing the economic valuation of each service on its direct biophysical measure, as opposed to looking at costs foregone with the soil conservation measure (Posthumus et al., 2013) enables a direct link to be made between soil properties, soil change and a change in the value of the service. When costs of infrastructure were used as proxies for the value of the service (defensive expenditures, replacement cost and provision cost), the construction costs of the infrastructure were annualised over its lifetime using a depreciation rate of 10% and added to annual maintenance costs to determine the annual value of the service. According to the information presented above on changes to the systems, the value of each service was calculated for a rolling landscape unit uneroded under pasture, a steep landscape unit uneroded under pasture, a steep landscape unit eroded immediately after a shallow landslide (year 0), and 20 and 50 years after, and finally a steep landscape unit with a pasture-tree system with 10 and 20 year old trees (Table 5).

210


Table 5 Economic value (NZD/ha/yr) of ecosystem services from uneroded pasture on rolling and steep landscape units, eroded steep land 0, 20 and 50 years after a shallow landslide and on a steep land pasture planted with wide-spaced trees. Soil service

Food quantity pasture Food quantity tree Food quality-pasture Wood-fibre Provision of support for human infrastructures Provision of support for farm animals Provision of shade Provision of shelter Flood mitigation Filtering of nutrients and contaminants Decomposition of wastes Net carbon accumulation (soil) Net carbon accumulation (tree) Nitrous oxide regulation Methane oxidation Regulation of pest and disease populations Combined value (NZD/ha/yr)

Uneroded

Eroded Steep landscape unit

Steep landscape unit with trees

Rolling landscape unit

Steep landscape unit

0 years

20 years

50 years

10 years

20 years harvested

745 NA 29 NA 204 53 NA NA 1155 2227 336 6 NA 2 0.08 328 5085

484 NA 29 NA 0 33 NA NA 911 1800 127 2.2 NA 1.2 0.08 328 3717

0 NA 0 NA 0 0 NA NA 273 634 21 0 NA 0 0 371 1299

290 NA 17 NA 0 17 NA NA 364 807 76 22 NA 0.5 0.03 328 1921

387 NA 23 NA 0 17 NA NA 456 978 102 1.8 NA 0.6 0.04 328 2293

411 105 29 104 0 33 58 9 938 1847 149 1.3 150 2.7 0.06 328 4165

363 210 29 104 0 33 58 19 990 1940 170 1.3 300 2.7 0.06 328 4548

NA: Not applicable

Fig. 1. Three scenarios – Scenario A – typical east coast hill country sheep and beef grazing; scenario B – Shallow mass movement erosion followed by recovery and scenario C – Soil conservation in hill country. (B-photo Brenda Rosser; C-photo Grant Douglas)

To investigate the impact of an investment in soil conservation on the provision of ecosystem services, the flows of services were considered on a per hectare basis over 20 years, using three scenarios illustrated in Fig. 1

Scenario A – Business as usual: provision of ecosystem services

from a sheep and beef farm for the two landscape units with no landslides and no addition of conservation trees, Scenario B – Shallow landslide followed by recovery: provision of ecosystem services from the steep landscape unit of a sheep and beef farm with a landslide in year 0 and subsequent recovery of soil and ecosystem services over 20 years, Scenario C – Soil conservation: Planting of conservation trees at 50 stems per hectare on the steep landscape unit at year 0 to reduce the risk of soil erosion, and subsequent tree development over 20 years.

For each of these scenarios the Present Value (PV) of the flow of ecosystem services (benefits) was calculated over 20 years. The present value (PV) of cash flows is a widely used criterion in benefit-cost analysis. It calculates the present value of a sum of money in year t (here the economic value of ecosystem services in year t) by discounting it at the rate r, arising between the present and a future date. The PV of a sum of money received in the future

is calculated using the following equation: PV ¼

Vt ð1 þ rÞt

where PV is the present value, t is the year, Vt is the value of the cash flow in year t, and r is the discount rate. Different discount rates, between 0% and 10%, were used in this study to show the sensitivity of the analysis to the choice of discount rate. The flow of services was assumed to be constant for 20 years for the two uneroded units, rolling and steep. For the steep unit with shallow landslide followed by recovery, it was assumed that the value of the services was recovering linearly from the value calculated for immediately after the landslide to the value after 20 years recovery (Table 5). For the soil conservation on steep landscape, it was assumed that the value of the services was similar to steep land with pasture only in the first 5 years, similar to the value calculated for a tree-pasture system in year 10 between year 6 and 10, and similar to the value calculated for a tree-pasture system in year 20 between years 10 and 20 (Table 5). Finally a benefit-cost (BCA) analysis of an investment in soil conservation on the steep landscape unit was realised. The extra costs and benefits associated with conservation trees were considered over 20 years on a per hectare basis, including the tree management costs and differences in the value of the ecosystem


services provided between a normal pasture and a tree-pasture system. The Net Present Value (NPV) of the investment on a per hectare basis was calculated by discounting the streams of benefits and costs at the rate r arising between the present and a future date, using the following formula: t

NPV ¼ ∑

Bt C t

t ¼ 0 ð1 þ rÞ

t

Where NPV is the net present value (PV of benefits PV of costs), t is the year (0–20), Bt is the value of the benefits in year t, Ct is the value of the costs in year t, and r is the discount rate. Four options were considered for the steep landscape unit as a combination of selling the trees for wood and considering or not the value of the ecosystem services provided in the BCA

Option 1: Trees are not sold for wood and the value of the ecosystem services provided are not considered in the BCA.

Option 2: Trees are sold for wood and the value of the ecosystem services provided are not considered in the BCA.

Option 3: Trees are not sold for wood, but the value of the ecosystem services provided are included in the BCA.

Option 4: Trees are sold for wood and the value of the ecosystem services provided are included in the BCA. For each option, the NPV of the investment was assessed using four discount rates between 0% and 10%.

3. Results and discussion

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thought and as a consequence do not feature in most assessments of land degradation following erosion. Until the impacts of erosion on these services have a recognised value, the long-term consequences of erosion on human well-being will continue to be under-estimated. Planting wide-spaced trees on pasture on the steep landscape unit, while decreasing the provision of specific services (Table 5), in general increased the provision of baseline ecosystem services and added extra services including forage from trees, wood, provision of shade and shelter for animals and net C accumulation in wood (Fig. 2). The presence of soil conservation trees increased the economic value of the ecosystem services (as calculated according to Tables 2 and 3) provided from NZD 3717 ha 1 yr 1 to NZD 4165 ha 1 yr 1 after 10 years and NZD 4548 ha 1 yr 1 after 20 years (Table 5), an increase in value of 12% and 22%, respectively. Had the conservation trees only grown at half the rate and associated services also halved, the economic value of ecosystem services would be 91% (NZD 3781 ha 1 yr 1) and 84% (NZD 3817 ha 1 yr 1) of the value using initial assumptions for the steep land with 10 and 20 year old trees, respectively (Fig. 3). While not included in the analysis, conservation trees planted on-farm will impact on the provision of cultural services, such as vista and landscape aesthetic values (Swaffield and McWilliam, 2014), land prices, and sense of stewardship of the land. Further, conservation trees also impact on habitat. For example poplar is known to be a food source for some native bird species in early spring (Yao and Kaval, 2010). Inclusion of these services would add to the economic value of the ecosystem services of a sheep and beef operation which has soil conservation plantings as part of the farms ecological infrastructure.

3.1. Economic Value of the provision of ecosystem services 3.2. Net present value of the flow of ecosystem services The economic value of ecosystem services (as calculated according to Tables 2 and 3) provided by uneroded pasture grazed by sheep and cattle was NZD 5085 ha 1 yr 1 for the rolling landscape unit and NZD 3717 ha 1 yr 1 for the steep landscape unit (Table 5). Of the eleven ecosystem services, filtering of nutrients and contaminants had the greatest economic value (44–49%), followed by flood mitigation (23–25%) and then the provision of food (13–14%). Regulating services, which are usually not traded in markets or considered in decision making, had an economic value four-times that of the provisioning services for the rolling landscape unit and almost six-times greater for the steep landscape unit (Fig. 2). The economic value of ecosystem services provided by bare ground following a shallow landslide on steep land was NZD 1299 ha 1 yr 1, a decrease of 65% of the services from uneroded land (NZD 3717 ha 1 yr 1), based on the assumptions in Table 4. Based on the work of Rosser and Ross (2011b) and the assumptions made, the initial recovery of provisioning services following a shallow landslide on the steep landscape unit was relatively quick with 59% recovery in the first 20 years, and after 50 years up to 78% of the economic value of the services of the uneroded landscape (Fig. 2). Further recovery is linked to longterm pedological processes. Recovery of the regulating services was predicted to be much slower, with only a 50 and 59% recovery of these services after 20 years and 50 years, respectively (Fig. 2). In the first few years after the erosion event, soil C accumulation and pest regulation rates are high, before slowing and stabilising. Assuming the recovery of the soil on the eroded land was only half of what was used in the initial analysis, the economic value of ecosystem services would drop to 53% (NZD 1013 ha 1 yr 1) and 53% (NZD 1205 ha 1 yr 1) of the value using initial assumptions for the steep land recovering after 20 and 50 years respectively (Fig. 3). The long-time frames required for the recovery of regulating services after an erosion event (Fig. 2) are generally given little

The three scenarios Scenario A – Business as usual, Scenario B – Shallow landslide followed by recovery, and Scenario C – Soil conservation were considered over 20 years to explore the Net Present Value of an investment in soil conservation. For each of these scenarios the NPV of the flow of ecosystem services, that is the flow of benefits coming from the land, was calculated over 20 years, using initial assumptions and different discount rates. When considering the recovery of the steep landscape unit after a landslide (100% eroded), it was assumed that the economic value of the ecosystem services provided was NZD 1299 ha 1 yr 1 in the first year (economic value of the services after erosion), increasing linearly to NZD 1922 ha 1 yr 1 over 20 years (economic value of services after 20 years recovery) (Table 5). When considering the provision of ecosystem services from the steep landscape unit under wide-spaced trees, it was assumed that the provision of ecosystem services was similar to pasture in the first 5 years (NZD 3717 ha 1 yr 1) (Table 5). Then, the provision of ecosystem services was assumed to be worth NZD 4165 ha 1 yr 1 between year 6 and 10, and NZD 4548 ha 1 yr 1 between years 10 and 20 (Table 5). The NPV of the ecosystem services from the steep landscape unit, for each of the three scenarios, are presented in Fig. 4, with different discount rates. When calculating the NPV of the flow of ecosystem services over 20 years for the three scenarios, changing the discount rate from 10% to 0% more than doubles the NPV (Fig. 4). Trees planted on the steep landscape unit increased the NPV of the ecosystem services by around 10% (using initial assumptions) but most importantly, by preventing soil erosion, reduces the risk of a loss of NPV of the flow of ecosystem services of around 58% over 20 years. Comparison of the provision of ecosystem services from the 20 year wide-spaced-planted tree-pasture system with the open pasture 50 years on from a landslide (Table 5), serves to highlight the key role of trees in protecting soil natural capital stocks and securing key regulating services.

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Fig. 2. Economic value (NZD/ha/yr) of ecosystem services provided by pasture grazed by sheep and cattle, on uneroded rolling and steep land, steep land immediately after a shallow landslide and following 20 and 50 years of recovery, and steep land planted with 10 and 20 year old wide spaced trees.

Fig. 3. Combined economic value (NZD/ha/yr) of ecosystem services provided by pasture grazed by sheep and cattle, on uneroded rolling and steep land, steep land immediately after a shallow landslide and following 20 and 50 years of initial and half initial assumed recovery rates, and steep land planted with 10 and 20 year old wide spaced planted trees growing at the initial or half initial growth rate.

After the sensitivity analysis, the NPV of ecosystem services was around 53% and 91% of the initial values presented in Fig. 4, for Scenario B and Scenario C respectively. After a single landslide, the economic value of the ecosystem services provided by the landscape under a sheep and beef operation decreased by 64% from NZD 3717 ha 1 yr 1 to NZD 1299 ha 1 yr 1 (Table 5). Assuming the land recovers over 20 years without another landslide, this represents a loss of NPV of NZD 33,989/ha (from NZD 59,020 to NZD 25,031 ha 1) over 20 years using a 3% discount rate (Fig. 4). Application of this loss in NPV across the 4300 ha affected by the April 2011 storm equates to NZD 146 million. This is an indication of the NPV of the ecosystem services permanently lost from the storm event, including the loss of pasture production and income to the farmer. This must be added to the actual material costs of the storm, estimated at NZD 39 million for infrastructure, and land, personal and commercial damage claims. This gives a total cost of the April 2011 storm in excess of NZD 185 million or more than four times the material clean-up cost that is quoted as the storm cost incurred by the community.

Soil conservation trees not only reduce the risk of erosion conserving the equivalent of NZD 33,989 ha 1 of NPV (Fig. 4), but also add extra services that when valued add an extra NZD 7043 ha 1, from NZD 59,020 ha 1 for uneroded steep pasture to NZD 66,063 ha 1 for steep pasture with trees (using a 3% discount rate) (Fig. 4). If half of the 4300 ha affected by the storm on the east coast on April 2011 had been planted in trees, reducing the risk of erosion, the added NPV of the ecosystem services provided by that land would have amount to NZD 15.1 million over 20 years using a discount rate of 3%. Further, those wide-spaced trees, by preventing erosion, would potentially have preserved a value of NZD 108 million (NZD 73 million of NPV of ecosystem services not lost, NZD 15.6 million of NPV of potential extra ecosystem services from trees, and NZD 20 million in costs that could have been avoided), for an original investment of NZD 1.6 million in planting 2150 ha, based on costs of NZD 736 ha 1 to plant the trees at 50 sph. In this case, every NZD 1 spent on soil conservation trees is worth NZD 68 of NPV of avoided infrastructure costs and lost services, and extra services provided. These calculations put the cost of soil erosion and the value of soil conservation in a different light from conventional evaluation


213

Fig. 4. Net present value (NPV; NZD/ha) over 20 years of the flow of ecosystem services associated with uneroded steep land (Scenario A), eroded steep land recovering (Scenario B) and steep land planted with conservation trees (Scenario C), for different discount rates.

Fig. 5. Net present value (NPV; NZD/ha) over 20 years of four different management options at for four discount rates.

of soil conservation practices. Accepting a number of the assumptions in the quantification and valuation of the services are open to interpretation and can be challenged, the methodology provides insights into the long-term effects of land degradation on the provision of the whole range of ecosystem services and the increased resilience brought with the addition of conservation trees on land at risk of erosion in storm events. This information should be available to decision makers for informing policy development and investment in soil conservation practices. 3.3. Soil conservation benefits and benefit-cost analysis The benefit-cost analysis of an investment in soil conservation explored four options that included a combination of selling the

trees for wood and considering or not the economic value of the ecosystem services (as calculated according to Tables 2 and 3, and reported in Table 5). For each option, the NPV of all the extra benefits and costs associated with conservation trees discounted over 20 years was calculated using four discount rates between 0% and 10% (Fig. 5). The economic value of the additional ecosystem services provided by the wide-spaced trees (Table 5), beyond the service provision from the open pasture, is not usually considered in Benefit cost analysis of soil conservation practices. Here, all the costs associated with growing poplars, including planting, pruning, pollarding and harvesting were considered, alongside the additional services this ecological infrastructure investment provided.

214


The option of selling the trees for timber after 20 years was also considered despite this option having limited applicability at present. The costs associated with tree planting were assumed to be NZD 736 ha 1 (for 50 stems per hectare) as plantation costs and NZD 200 ha 1 in year 7 and 17 for pruning. The cost of harvesting and transport was subtracted from the net revenue coming from the timber sale in year 20. The NPV of the different options are presented in Fig. 5. Planting soil conservation trees without harvesting them (Option 1) is not profitable if the extra provision of ecosystem services is not considered (NPV ¼ NZD 998 at 3%) (Fig. 5). However, if the extra provision of ecosystem services is considered (Option 3) the NPV of the investment is positive regardless of the discount rate used (Fig. 5). If the trees are harvested for timber after 20 years but the ecosystem services not considered (Option 2), the investment is only profitable for discount rates below 5%. Again, if considering the economic value of the extra provision of ecosystem services from trees as well as timber sale (option 4), the NPV of the investment is greatly positive regardless of the discount rate (Fig. 5). The sensitivity analysis showed that with lower tree growth rates, the economic value of the ecosystem services provided by the steep landscape unit with conservation trees is reduced, and the investment is only profitable if the trees are harvested for wood. With slower growth rates it would be necessary to extend the analysis out beyond 20 years to assess the merits of a space planted trees versus other soil conservation investment options. Ultimately, the final decision on a soil conservation investment will be dependent on the type, risk and severity of erosion, the time period over which the investment is investigated, in addition to the decision to include the impact on ecosystem services and the economic value associated with each service. Further, the allocation of ecosystem services economic values to private versus public beneficiaries and the level of the return on the next best use of the money invested in soil conservation will also be factors. This study shows how much of a difference inclusion of the whole range of ecosystem services and their economic value makes to the estimated NPV of conservation practices. The ecosystem services approach described here offers a more complete picture of the “costs” of soil erosion, its long-term effects on the provision of services, and the ‘value’ of soil conservation practices (Robinson et al., 2014).

4. Conclusion This study through a live case on the East Coast of the North Island of New Zealand addresses soil conservation and shows how an ecosystem services approach can be used to quantify and value the long-term costs of soil erosion, here shallow mass movements, and the multiple benefits from soil conservation. While this study is based on current research knowledge, some assumptions were made where data was lacking regarding changes to the measure of ecosystem service from eroded landscapes and landscapes with soil conservation trees. When data becomes available this study can be revisited. The methodology tested in this study demonstrates that using existing and detailed information about ecosystems change at a small scale to quantify the provision of provisioning and regulating ecosystem services from agro-ecosystems is not only feasible, but also extremely informative for decision makers for not only policy development, but also for evaluating the impacts of policy. Understanding how current and future investments in ecological infrastructure, such as soil conservation trees, are likely to change the flow of ecosystem services from managed landscapes is critical to assess the efficiency, cost-effectiveness and sustainability of

resource management policies, and to increase political and public awareness of the value of land and the long term costs of land degradation.

Acknowledgement Funding for this study was provided by the Envirolink tools programs through Hawke's Bay Regional Council, as well as the Rutherford foundation of the Royal Society of New Zealand and the Sustainable Land Use Research Initiative. Thank you to colleagues and reviewers for their valuable comments and suggestions.

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