AQUATIC CONSERVATION: MARINE AND FRESHWATER ECOSYSTEMS
Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) Published online 10 December 2008 in Wiley InterScience (www.interscience.wiley.com). DOI: 10.1002/aqc. 1007
Making agricultural landscapes more sustainable for freshwater biodiversity: a case study from southern England BELLA DAVIESa,b,, JEREMY BIGGSb, PENNY WILLIAMSb and STEWART THOMPSONa Spatial Ecology and Landuse Unit (SELU), School of Life Sciences, Oxford Brookes University, Gipsy Lane, Oxford OX3 0BP, UK b Pond Conservation: The Water Habitats Trust, c/o School of Life Sciences, Oxford Brookes University, Gipsy Lane, Oxford OX3 0BP, UK
a
ABSTRACT 1. Agriculture is known to have a range of deleterious impacts on freshwater habitats and biota and many countries have introduced measures to attempt to mitigate these impacts through agri-environment initiatives. Despite the protection they provide, water bodies (any discrete body of surface fresh water) in farmland landscapes commonly remain impaired by agriculture. In some areas of the UK there have been calls to halt farming completely, indicating that the measures offered for the widespread protection of aquatic systems, particularly the use of buffer strips, may not be extensive enough to provide sufficient protection for freshwater biota. 2. This study investigated whether existing agri-environment measures for the widespread protection of aquatic habitats could be better deployed to provide a higher level of protection for the aquatic macrophytes and macroinvertebrates of a study area in southern England. 3. Reserve selection procedures were used to reallocate the area of land that could be remunerated under the Environmental Stewardship scheme as buffer strips bordering water bodies, so that a higher level of protection was provided for both the richness and rarity of aquatic species in the study area. 4. Almost 395 ha were available for reallocation in the reserve selection process, which was found to provide a satisfactory level of protection for up to 90% of the surveyed species. 5. The results showed that the agri-environment scheme in England has a great deal of potential to provide more effective protection for the aquatic biodiversity of agricultural landscapes if measures are targeted. Copyright r 2008 John Wiley & Sons, Ltd. Received 4 December 2007; Revised 6 June 2008; Accepted 3 August 2008 KEY WORDS:
catchment area; reserve selection; aquatic biota; agri-environment; buffer strips; diffuse pollution; Marxan
INTRODUCTION Widespread declines in water quality and freshwater biodiversity have been observed as a result of agricultural activities, including: habitat loss through land drainage; pollution from chemicals, nutrients, animal waste and animal health by-products; and sedimentation resulting from soil erosion (Chuffney et al., 2000; Stoate et al., 2001; Allan, 2004; Foley et al., 2005; Declerck et al., 2006). To mitigate agricultural impacts on both freshwater and terrestrial environments, many European countries have introduced agri-environment initiatives. These are government schemes which provide financial incentives to farmers and land managers for adopting environmentally sensitive practices which maintain or enhance the natural resources, and in particular the
biodiversity, of agricultural land. Given the large scaleof agriculture, especially in Western Europe, these initiatives provide the most widespread opportunity for biodiversity protection in agricultural landscapes outside of specially designated areas and have entailed considerable expenditure (3.5 billion euros per year in Europe (Whitfield, 2006)). In England, agri-environment initiatives are also likely to be very important for the attainment of ecological targets set by the Water Framework Directive (2000/60/EC) because diffuse pollution from agriculture has been identified as a major pressure which will put water bodies at risk of failing to achieve ‘good ecological status’ by 2015. Despite the expenditure on agri-environment initiatives, there has been limited monitoring of their benefits for
*Correspondence to: Bella Davies, Spatial Ecology and Landuse Unit and Pond Conservation: The Water Habitats Trust, School of Life Sciences, Oxford Brookes University, Gipsy Lane, Headington, Oxford OX3 0BP, UK. E-mail:
[email protected]
Copyright r 2008 John Wiley & Sons, Ltd.
440
B. DAVIES ET AL.
protecting wildlife, and where monitoring has taken place their success has been found to be mixed (Carey, 2001; Kleijn et al., 2001, 2006; Wilson et al., 2007). Apart from wetland birds, e.g. snipe (Gallinago gallinago), redshank (Tringa totanus) and lapwing (Vanellus vanellus) (Wilson et al., 2007), monitoring has concentrated on terrestrial taxa, especially vascular plants, carabid beetles, bees, butterflies, grasshoppers, spiders and birds (see reviews by Kleijn and Sutherland, 2003 and Kleijn et al., 2006) and little is known about the effectiveness of agrienvironment schemes for aquatic species. Where data on the water quality of agricultural landscapes have been collected, they have often indicated that aquatic systems remain degraded. In eastern England the situation has been found to be so poor that calls have been made to halt farming, with the possible ‘rewilding’ of large areas unless ‘smart solutions’ can be found which enable the co-existence of both good water quality and agricultural production in the same landscape (ADAS, 2006; Clover, 2006; Moss, 2008). The English agri-environment initiative, ‘Environmental Stewardship’, aims to prevent widespread pollution of aquatic systems in agricultural areas through its lower tier, the Entry Level Scheme (ELS), with the use of buffer strips. These are areas of land adjacent to water bodies (any discrete body of surface water, including rivers, lakes, streams, ditches and ponds) which are left free from agricultural production with the intention of intercepting nutrients, chemicals and sediment that drain towards water bodies. The efficacy of buffer strips for water body protection has been widely investigated and most studies have suggested that they need to be in excess of 25 m wide to be effective (see reviews by Castelle et al., 1994; Environmental Law Institute, 2003; Hickey and Doran, 2004; Lovell and Sullivan, 2006). However, the maximum width remunerated under Environmental Stewardship is much narrower — 7 m for water bodies at the edge of a field (e.g. a 6 m buffer adjacent to 1 m of uncultivated land bordering a ditch) and 10 m for in-field ponds. In addition, almost all investigation of buffer strip effectiveness has been undertaken at the plot or field scale, with limited evidence that buffer strips make a difference to water quality at the catchment scale (Krutz et al., 2005). This has considerable implications for whether the ecological targets of the Water Framework Directive can be achieved using current measures because the Directive indicates that management should be undertaken on a whole catchment basis. Given the influence of a water body’s catchment area on its chemical and biological characteristics (Hynes, 1975; Allan et al., 1997; Allan, 2004), studies have recently begun to investigate the relationship between the proportion of different land-use types within a catchment and water-body quality. Many of these studies indicate that buffer strips are unlikely to incorporate a sufficient proportion of a catchment area to protect a water body effectively, with declines in biodiversity being observed when agriculture exceeds 30–50% of a catchment (Allan, 2004). In contrast, positive associations have been observed between a range of aquatic organisms and semi-natural habitats, including forests (Roth et al., 1996; Roy et al., 2003), deciduous, coniferous and mixed woodland and unimproved grassland (Davies et al., 2004), and wetlands (Roth et al., 1996). It is, therefore, likely that the land used as buffer strips for waterbody protection could be better ‘spent’ elsewhere by concentrating it to form areas of semi-natural land in locations where a maximum complement of species can be protected. Copyright r 2008 John Wiley & Sons, Ltd.
This was investigated using reserve selection techniques to ascertain whether current agri-environment resources for the widespread protection of aquatic habitats could be reallocated to give better protection for the aquatic plants and macroinvertebrates of a study area in southern England, based on scenarios of land-use deintensification within a water body’s catchment to levels identified from the literature as critical for aquatic biota.
METHODS The study area and its aquatic biodiversity The study area covered 143 km2 of lowland agricultural landscape located on the borders of Oxfordshire, Wiltshire and Gloucestershire in southern England. Arable cultivation dominated the landscape (75%) and predominantly comprised oil seed rape, cereals, maize, beans, potatoes, peas and permanent grassland (Brown et al., 2006). Other land cover included woodland (9%), unimproved grassland (7%), urban (2%) and water and bare rock (the remaining 7%). Surface water covered 205 ha of the study area and was categorized as ditches, lakes, ponds, rivers and streams (Table 1). A stratified random sample of 100 sites was selected for survey from the ditches, lakes, ponds, rivers and streams in the study area (20 in each water body type). Each survey site was area-limited to 75 m2 to enable the comparison of data from water bodies of very different sizes. The survey areas were designed to be representative of the habitats within the water body types. In linear water bodies the survey area was a rectangular section including both banks, and in standing (more circular) water bodies it was wedge-shaped with the apex at the centre of the water body and the base along the edge. Aquatic plants and macroinvertebrates were surveyed using the same methods at each site during a single visit, with half the sites in each water body type being surveyed in spring (April and May) and half in autumn (October and November). Table 1. Definitions of the water body types into which surface water was categorized (after Williams et al. (2004) and Davies et al. (2008a)) Water body type
Definition
Ditches
Man-made channels created primarily for agricultural purposes and which usually: (i) have a linear planform; (ii) follow linear field boundaries, often turning at right angles; and (iii) show little relationship with natural landscape contours. Bodies of water, both natural and man-made, greater than 2 ha in area (Johnes et al., 1994). Includes reservoirs, gravel pits, meres and broads. A body of water, both natural and man-made, between 25 m2 and 2 ha in area, which may be permanent or seasonal (Collinson et al., 1995). Relatively large lotic water bodies, created by natural processes. Marked as a double blue line on 1:25 000 Ordnance Survey (OS) maps and defined by the OS as greater than 8.25 m in width. Relatively small lotic water bodies, created by natural processes. Marked as a single blue line on 1:25 000 OS maps and defined by the OS as being less than 8.25 m in width. Streams differ from ditches by usually: (i) having a sinuous planform; (ii) not following field boundaries; and (iii) showing a relationship with natural landscape contours, usually by running down valleys.
Lakes Ponds Rivers
Streams
Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
MAKING AGRICULTURAL LANDSCAPES MORE SUSTAINABLE FOR FRESHWATER BIODIVERSITY
Macrophytes (marginal, emergent, floating-leaved and submerged plants, including charophytes and excluding bryophytes, as listed in PCTPR, 2002) were surveyed by wading all shallow areas and using a grapnel thrown from the bank or a boat. All macrophytes were identified to species level in the field and, where necessary, in the laboratory. Macroinvertebrates were surveyed using a standard 1 mm mesh hand-net for three minutes, with the total sampling time being divided equally between all major mesohabitats in the survey area, e.g. open water, bank and stands of vegetation with differing structure. The samples were live-sorted in the laboratory and identified to species level, except for Diptera larvae and Oligochaeta which were omitted from analysis. Each species was assigned a score according to its rarity or threat within Britain (see Nicolet et al. (2004) for sources of status data). Six categories were used (Williams et al., 2004):
Common (scoring 1). Local (2): Macrophytes: local species recorded from less than 25% (between 101 and 700) of the 10 10 km grid squares in Britain (Preston et al., 2002); Macroinvertebrates: species either (a) confined to limited geographical areas; or (b) of widespread distribution but relatively low population. Nationally Scarce (4): Recorded from 15–100 10 10 km grid squares in Britain. Red Data Book 3 (8): Vulnerable. Red Data Book 2 (16): Endangered. Red Data Book 1 (32): Critically Endangered.
Where the old IUCN system was in use for some invertebrate groups, the terms under the latest IUCN Red List were applied.
Catchment delineation and land-use composition The catchment area of a water body was considered to be the area over which surface runoff would flow into it. The subsurface catchment was not modelled because, although some throughflow and transport by field drains would have occurred, it was not considered to strongly influence the results owing to the predominantly impermeable clays that underlie the study area. Ordnance Survey (OS) MasterMaps data were used to identify all water bodies within the study area using ArcGIS 8.2. OS Landform Profiles data were used to create a digital elevation model (DEM) and the water bodies were burnt into it at their height value minus 10 m so that they were retained (and not filled in) during the catchment delineation process. The catchment area of each survey site was delineated separately using the ArcGIS extension ArcHydro Tools (Version 1.1 Beta 2; ESRI, 2001) by the method described in Davies et al. (2008a). An important part of this method was seeding standing water bodies (ponds and lakes) with ‘no data’ points to ensure ArcHydro Tools went against its underlying assumption that all water should flow to the edge of the DEM, thus retaining water in the standing water bodies. The river catchments extended beyond the limit of the data available and so their true size was estimated from information published for the nearest gauging station (Environment Agency, 2006). The proportion of agricultural land within each catchment area was assessed using Land Cover Map 2000 data from the Centre for Ecology and Hydrology (copyright NERC). Copyright r 2008 John Wiley & Sons, Ltd.
441
Improved and setaside grassland were included as agricultural land which ensured that any pasture land and land under crop rotation were included. For river sites, the proportion of the catchment under agricultural land use that could be modelled with the available data was taken to be representative of the whole catchment. Previous studies which had investigated a relationship between the proportion of a catchment area under agricultural use (both arable and pastoral) and the critical thresholds at which deleterious changes in aquatic communities occurred were identified from the literature so that these thresholds could be applied to the study area. Relatively few of these studies have been undertaken (and none in the UK), so it was difficult to identify catchments of comparable physicochemical characteristics (geology, local climate, land-use intensity, etc.) with the present study location. Therefore, the threshold above which biotic degradation might occur in the study area was taken to be the median threshold identified from all of the studies. For each survey site, the area of its catchment that would need to be deintensified to reach the median critical threshold was calculated.
Reallocation of agri-environment land using reserve selection Reserve selection is a process that aims to identify the optimum complement of sites to be included in a reserve network so that maximum biodiversity is incorporated for minimum costs (or within a specified budget), ensuring that conservation resources are used most effectively. Thus, each site has a value according to the species it supports and a cost which may be monetary, or otherwise, e.g. land area. The budget available for ‘spending’ in the present study was the total area of land that could have been remunerated for buffer strip implementation under the English ELS. This area was calculated using the GIS to buffer all edge of field water bodies (ditches, lakes, streams and rivers) by 7 m and all in-field water bodies (ponds) by 10 m, in accordance with the existing payment mechanism (Defra, 2005). Both the species richness of a site and the national rarity status of those species were incorporated into the reserve selection process so that the optimal reserves included as great a number of species and as many rare species as possible within the ‘spending’ allocation. This differed from most previous reserve selection queries which have traditionally designated reserves based on species richness or species rarity separately. Species rarity was measured by its national rarity status (see above), which was included in the reserve selection process by setting higher targets of representation in the reserve for rarer species (Table 2), with any site that included a Red Data Book species being automatically fixed in the reserve so that the species was always protected. The cost of including a site in the reserve was related to the land area involved in its protection, which was set according to two scenarios with different costs: Scenario 1. A water body is given as extensive protection as possible by completely deintensifying the agricultural land in its catchment, i.e. the cost of a site was the area of agricultural land within its catchment area. Scenario 2. A water body is given at least a satisfactory level of protection by deintensifying the area of land under agricultural use that exceeds the median critical threshold Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
442
B. DAVIES ET AL.
Table 2. Targets set in Marxan 8.1 for the selection of rare species in a reserve Status
Species Rarity Target of representation in the Score reserve
Common Locally Rare Nationally Scarce Red Data Book ‘Vulnerable’ Red Data Book ‘Endangered’ Red Data Book ‘Critically Endangered’
1 2 4 8
1 occurrence 420% of sampled occurrences 440% of sampled occurrences Site fixed in reserve
16
Site fixed in reserve
32
Site fixed in reserve
identified from the literature, above which aquatic biota has been found to be degraded, i.e. the cost of a site was the area under agricultural use that exceeded the median threshold (%) identified by previous studies, above which degradation of aquatic biota had been observed. The computer software Marxan 1.8. (Ball and Possingham, 2000; Possingham et al., 2000) was used to identify reserves for plants and macroinvertebrates separately and then for all plant and macroinvertebrate species together. Simulated annealing selection procedures were used followed by iterative improvement. Adaptive annealing was used to set the initial temperature and cooling factor, with 10 000 temperature decreases and one million iterations. A cost threshold was applied, which was the total maximum area that could be remunerated for water body protection by buffer strips through the ELS. Each selection procedure was run 500 times with the best reserve selected as a result.
RESULTS Within the study area there were 340 ditches, 8 lakes, 236 ponds, 97 streams and 3 rivers. The maximum area of land that would be remunerated for conversion to buffer strips under the ELS to protect these water bodies, and thus the area for reallocation using reserve selection, was 394.91 ha. This represented 2.7% of the study area. Across the study area, 361 aquatic species were surveyed, 94 of which were macrophytes and 267 were macroinvertebrates. On average, there were 37 species recorded at a site (ranging from 5–82), with 8 macrophytes (ranging from 1–19) and 28 macroinvertebrates (ranging from 2–67). The 94 macrophyte species surveyed represented 26% of the aquatic macrophyte species present in Great Britain (PCTPR, 2002), with 19% of British species occurring in the pond sites, 14% in the river sites, 13% in the lake sites, 11% in the stream sites and 8% in the ditch sites. The 267 macroinvertebrate species represented 33% of the aquatic macroinvertebrate species of Great Britain (Furse et al., 2003), with 21% of British species occurring in the pond sites, 19% in the river sites, 17% in lake sites, 15% in
stream sites and 11% in ditch sites. Thus, the species surveyed were not distributed equally between water body types and although ponds were the most species-rich water bodies at the landscape scale (i.e. gamma diversity), rivers supported the greatest species richness at the site level (i.e. alpha diversity), with ditches supporting the fewest species for both plants and macroinvertebrates (Table 3). Of the species surveyed, 17% had a rarity status greater than 1 (common). Of these species, 67.7% were Local, 30.6% were Nationally Scarce and 0.3% were Red Data Book (Table 4). Of the surveyed sites, ponds had the smallest mean catchment size followed by ditches, lakes, streams and rivers, which had the largest mean catchment size being over 160 times larger than the average for pond catchments (Table 5). The catchments of the lake survey sites comprised the lowest mean proportion of agricultural land, with ditch, pond, river and stream sites all having similar mean proportions of agricultural land (between 78% and 83%) in their catchments (Table 5). Six studies had identified a threshold at which the proportion of a catchment under agricultural use was related to a decline in aquatic biota. These varied from 1.3%–50% agriculture, with a median value of 27.5% (Table 6). Over 95% of the surveyed sites had catchments with a greater proportion of agricultural land use than this value. All of the sites with less than 27.5% agricultural land use in their catchments were lakes. With the cost threshold of 394.91 ha, the reserve selected under Scenario 1 (i.e. deintensification of all agricultural land within a site’s catchment) resulted in 18 sites being selected to protect macrophyte species, 24 for macroinvertebrate species and 25 for all species together (Table 7). These reserves protected 88% of the surveyed macrophytes (including all rare species), 82% of the surveyed macroinvertebrates and 81% of all the surveyed species. Under Scenario 2 (i.e. deintensification of agricultural land to less than 27.5% of a site’s catchment area), 24 sites were selected for inclusion in a reserve for macrophyte protection, 36 sites were selected for macroinvertebrate protection and 39 sites to protect all species together (Table 8). These reserves would provide a satisfactory level of protection for 90% of the surveyed
Table 4. The number (and %) and rarity status of the species surveyed in the study area Rarity status and score
All species
Macrophytes
Macroinvertebrates
Common (1) Local (2) Nationally Scarce (4) Red Data Book 1 (8) Red Data Book 2 (16) Red Data Book 3 (32)
299 42 19
83 11 0
216 31 19
(82.8%) (11.6%) (5.3%)
0 1 0
(0.3%)
(88.3%) (11.7%)
0
0
0
1
0
0
(81.0%) (11.6%) (7.1%)
(0.4%)
Table 3. Mean site richness by water body type, with range in parentheses
Macrophytes Macroinvertebrates
Ditches
Ponds
Lakes
Rivers
Streams
6.1 (1–14) 12.9 (2–35)
10.0 (2–17) 32.2 (5–67)
6.8 (1–14) 33.0 (21–49)
10.7 (6–19) 45.3 (18–66)
7.3 (1–17) 18.7 (3–50)
Copyright r 2008 John Wiley & Sons, Ltd.
Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
MAKING AGRICULTURAL LANDSCAPES MORE SUSTAINABLE FOR FRESHWATER BIODIVERSITY
443
Table 5. Mean catchment size and agricultural land-use composition of the surveyed water bodies, with range in parentheses
Catchment area (ha) Agricultural land use (%)
Ditches
Ponds
Lakes
Rivers
Streams
92.80 (1.77–474.63) 83.2 (53.9–99.9)
27.62 (1.04–89.94) 79.3 (33.6–99.7)
132.23 (16.16–361.07) 51.2 (18.7–75.5)
4569.91 (4372.54–6346.23) 77.7 (77.5–79.9)
281.22 (10.50–1359.92) 78.1 (60.9–97.6)
Table 6. Thresholds of agricultural land use within a catchment area above which declines in aquatic biota have been observed Study/Reference
Location
Organism
Index Species richness; EPT
Quinn and Hickey (1990)
New Zealand
Macro-invertebrates and fish
Wang et al. (1997) Fitzpatrick et al. (2001) in Allan (2004) Weigel (2003)
Wisconsin, USA Eastern Wisconsin, USA Wisconsin, USA: North ecoregion Driftless Ecoregion Ireland
Fish Fish
Donohue et al. (2006) Park et al. (2006)
Adour-Garonne Basin, France
Macro-invertebrates Macro-invertebrates Benthic macro -invertebrates Fish
Water body
Rivers scores; biomass of trout and sensitive macroinvertebrate species. Index of Biotic Streams IBI Streams Integrity (IBI) 14 metrics Streams
Threshold 30% improved pasture
50% agriculture 30% agriculture 7.5% pasture; 7.5% arable
14 metrics Quality Rating System
Streams Rivers
25% pasture; 20% arable 37.7% pasture; 1.3% arable
Assemblage types
Streams
40% agriculture
Ephemeroptera, Plecoptera, Trichoptera.
Table 7. Reserves selected under Scenario 1 providing a high level of aquatic biodiversity protection
Macrophytes
Area of Number of Number (and %) Number (and rarity) reserve (ha) sites included of species of omitted species included
Number of each water body type included in the reserve 10 ponds, 4 lakes, 3 ditches, 1 stream, no rivers
383
18
83 (88%)
11 (all Common)
Macroinvertebrates 383
24
219 (82%)
48 (34 Common, 11 Local, 12 ponds, 5 ditches, 4 lakes, 3 streams, no rivers 3 Nationally Scarce)
All species
25
293 (81%)
68 (51 Commom, 14 Local, 13 ponds, 5 ditches, 4 lakes, 3 streams, no rivers 3 Nationally Scarce)
392
Table 8. Reserves selected under Scenario 2 providing a satisfactory level of aquatic biodiversity protection Area of Number of Number (and %) Number (and rarity) Number of each water body type included in the reserve reserve (ha) sites included of species included of omitted species Macrophytes
386
24
85 (90%)
Macroinvertebrates 380
36
233 (87%)
All species
39
310 (85%)
389
9 (6 Common, 10 ponds, 9 lakes, 4 ditches, 1 stream, no rivers 3 Local) 34 (26 Common, 13 ponds, 12 lakes, 7ditches, 4 streams, no rivers 6 Local, 2 Nationally Scarce) 51 (39 Common, 15 ponds, 12 lakes, 7 ditches, 5 streams, no rivers 10 Local, 2 Nationally Scarce)
macrophyte species, 87% of the surveyed macroinvertebrates or 85% of all the species surveyed in the study area, although some rare species were omitted from each reserve. The reserve selection processes selected a similar proportion of the various water body types in each reserve (although the same sites were not always selected). Pond sites were selected most often (comprising 40% of all the sites that were selected: 52% of the sites selected under Scenario 1 and 38% under Scenario 2), followed by lake sites (comprising 24% of all sites selected: 18% of the sites selected under Scenario 1 and 33% under Scenario 2), ditch sites (comprising 17% of all sites Copyright r 2008 John Wiley & Sons, Ltd.
selected: 19% of the sites selected under Scenario 1 and 18% under Scenario 2), stream sites (comprising 9% of all sites selected and 10% of the sites selected under both scenarios) and finally river sites which were omitted from all reserves.
DISCUSSION Limited resources are available for biodiversity protection and, therefore, it follows that they should be allocated to locations where they will be most effective, i.e. they will give maximum Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
444
B. DAVIES ET AL.
biodiversity return for the cost and effort input (Pressey and Tully, 1994; Thompson et al., 1999; Wa¨tzold and Schwerdtner, 2005). In contrast to this, the entry level tier of the current English agri-environment scheme, Environmental Stewardship, takes a broad approach and any farmer can apply. As a result, the measures offered for the widespread protection of aquatic habitats have to be limited in extent as resources are potentially spread equally across the landscape. Consequently, most of the buffer strip widths offered are more than three times narrower than those suggested as necessary in the literature and so the level of protection given to aquatic biodiversity at the landscape scale will not be sufficient. Results of the present study demonstrate that a much higher level of protection could be given to aquatic biodiversity across agricultural landscapes with no additional remuneration, if the protection were targeted to create more extensive areas of semi-natural habitat around water bodies of high biodiversity value. Under this method, up to 90% of the species surveyed would have been satisfactorily protected, including most rare species, a remarkably high proportion of the biodiversity given the relatively small land area involved in protection. When identifying sites or areas for the protection of biological diversity, it is often difficult to take into account both its elements of species richness and the presence of rare species (Hamaide et al., 2006) because richness and rarity hotspots frequently do not coincide (Palmer, 1999; Orme et al., 2005). Thus, the method used here to incorporate both species richness and the degree of rarity of the species present into reserve design is a useful tool which has rarely been used (Arponen et al., 2005; Hamaide et al., 2006; Early and Thomas, 2007). In reserve design, rarity is often measured by the frequency of a species’ occurrence within a dataset, rather than by a designated status. Although this can be useful in preserving declining populations, it can also encourage the focusing of resources at common species that are naturally becoming locally extinct (but are still abundant elsewhere). Therefore, the identification of reserves according to nationally designated status can ensure that resources are targeted at species which are truly rare and can help meet national and international biodiversity targets. Traditionally, reserve selection problems are solved based on single habitat types, particularly for aquatic systems, e.g. ponds (Briers, 2002), rivers (Linke et al., 2007) and lakes (Virolainen et al., 1999), largely because data collection is oriented towards a single type of water body. However, many aquatic organisms use more than one water body type; indeed in the present dataset, 68% of species were found in more than one water body type and 13% were present in all five water body types. This demonstrates the importance of including a range of aquatic habitats in freshwater protection strategies and, as far as we are aware, this is the first time that such a range of water bodies has been included in reserve selection procedures. Ponds were included most often in every reserve, probably due to a combination of their high biodiversity value both for species richness and rarity and their small catchment sizes (enabling a number of sites to be included under the reserve threshold). Recent studies which have compared the biodiversity characteristics of different water body types have consistently found that, at the landscape scale, ponds support the most species, in both the UK (Williams et al., 2004; Biggs et al., 2007) and across Europe (Davies et al., 2008b; De Bie et Copyright r 2008 John Wiley & Sons, Ltd.
al., 2008). Ponds have also been found to be important habitats for rare aquatic species (Nicolet et al., 2004; Davies et al., 2008a, 2008b) and within the present study it was a pond which supported the rarest sampled species (a Red Data Book 2 water beetle). In addition, a recent comparison of the catchment sizes of different water body types found ponds to have the smallest catchments (Davies et al., 2008a) and so they were the ‘cheapest’ water body type to include in the reserves. With so many ponds included in the reserves, it would be advisable to take a precautionary approach when applying the scenarios involving catchment deintensification and deintensify as great a proportion of the catchment areas as possible (i.e. Scenario 1). This is because all of the studies that had investigated the relationship between the proportion of agricultural land use in a catchment and aquatic communities had been undertaken for lotic water bodies. In contrast ponds are small lentic systems and have less water available to dilute pollutants and higher rates of sediment (and pollutant) accumulation and, therefore, the critical thresholds at which pond communities are degraded may well occur at lower proportions of agriculture than for lotic systems. With respect to the catchment studies reported in the literature, it is also worth noting that they are very few in number and that they have all concentrated on fish or macroinvertebrates. It is likely that the communities of other taxa, e.g. plants, could show signs of degradation at different proportions of agriculture within a catchment. Thus, further investigation is needed of the responses of different types of water body and of different taxa to the proportion of agriculture within a catchment area. In contrast to the dominance of ponds in both reserves, river sites (and thus species that were unique to these water bodies) were completely omitted from every reserve because their large catchment sizes were always above the cost threshold. Thus, under current agri-environment spending, whether it is allocated as at present by the ELS, or reallocated to target protection in key areas, it is likely that the aquatic biota of rivers and some larger streams will not be protected effectively because the proportion of their catchment areas under agricultural land use remain above critical levels. However, even if resources were available for the large-scale deintensification of agricultural land within river catchments, it is unlikely that it would be practical or desirable because it would mean an increased reliance on imports of food and fuel crops as well as a loss of jobs in the agricultural sector and a change to the landscape character of many areas of England. Given that rivers are important ecosystems with many different socio-economic, environmental and ecological values, alternative methods must be sought to ensure their effective protection from diffuse pollution from agriculture. This will be particularly important if rivers are to meet their target of good ecological status set by the Water Framework Directive by 2015. The management of riparian areas is thought to be important for reducing land-use impacts on aquatic biota and is considered one of the most cost-effective actions to significantly improve the ecological status of rivers (Stewart et al., 2001; Gergel et al., 2002; Dobiasova et al., 2004). However, for a riparian zone to be effective for river protection, the width left free from agricultural production would need to be substantial and probably in excess of 100–200 m (Fitzpatrick et al., 2001; Gergel et al., 2002; Environmental Law Institute, Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
MAKING AGRICULTURAL LANDSCAPES MORE SUSTAINABLE FOR FRESHWATER BIODIVERSITY
2003). In addition, the spatial configuration of land parcels of different intensities (from intensive agriculture to semi-natural land) is also likely to affect aquatic biota (Johnson et al., 1997; Teels et al., 2006) and this may well have led to some of the differences between the critical thresholds of agriculture identified in the literature. The arrangement of land of differing cover and intensity within a catchment needs further investigation, as it could be important in improving protection for rivers, and for informing strategies of catchment deintensification, such as the two scenarios considered in the present study. There could also be a potential role for organic farming in the complement of less intensive land uses offered as an alternative to agricultural land for the protection of aquatic systems. Reserve selection procedures sometimes involve the aggregation of sites to enhance connectivity, encourage species dispersal and population sustainability (Williams et al., 2005; Van Teeffelen et al., 2006). This could not be incorporated into the methods used in the present study and it may be valuable in creating more sustainable aquatic populations. However, the result of such an approach would, in theory, be to concentrate land use into zones of semi-natural land and agriculture, which contradicts the aim of a ‘smart solution’ facilitating the coexistence of agricultural production and good aquatic biodiversity within the same landscape. In addition, the incorporation of spatial penalties into reserve selection tends to make reserves more costly (Briers, 2002) and with a relatively small amount of land available for reallocation in the present study, the inclusion of a spatial penalty is likely to have dramatically decreased the species richness and rarity value of the reserves. Thus, it may be difficult to make deintensified areas more connected without increasing the allocation of agri-environment resources for aquatic biodiversity protection. The deintensification of key areas to protect aquatic systems has a number of additional wildlife and other benefits. First, it provides the potential to create substantial areas of terrestrial habitat with an increased core area, which can support more permanent populations than could be achieved by narrow buffer strips spread through the landscape (Forman, 1995). (The functional benefits of buffer strips for terrestrial species, e.g. as corridors through the agricultural landscape, remain where they are implemented at field boundaries which do not adjoin water bodies.) Second, it provides the opportunity to create new wet habitats to further enhance aquatic biodiversity at the landscape scale, especially where such habitats have been lost and are sparse in the landscape. For example, Weigel (2003) found very poor macroinvertebrate assemblages to occur in catchments with less than 1% wetland and open water. Small water bodies, such as ponds, are relatively easy and cheap to create but provide potentially high biodiversity benefits even when they are relatively newly created (Williams et al., 2008). Third, concentration of semi-natural habitats, e.g. woodland, speciesrich hay meadows and wet grassland, may help to fulfil other aims of Environmental Stewardship — for example, enhancing landscape character, protecting the historic environment and natural resources and providing access and use of the countryside for the general public (Defra, 2005). Finally, there is the potential for additional benefits, including increased tourism, improved human health, flood reduction Copyright r 2008 John Wiley & Sons, Ltd.
445
and carbon sequestration (Sutherland, 2004; Pretty et al., 2007; Downing et al., 2008). To implement such a targeted approach, current measures offered under Environmental Stewardship would need to be extended. The higher level tier of the agri-environment scheme is based on a broadly targeted approach and so offers some scope for the inclusion of catchment deintensification as a method of protecting and enhancing aquatic biodiversity. More weight would be given to the concept if it were included within the programmes of measures identified to achieve Water Framework Directive targets. However, this relies heavily on agri-environment schemes which are taken up on a voluntary basis. Thus, where areas identified for deintensification cross farm boundaries, cooperation between land managers and a coordinated approach would be necessary to ensure the measure’s success. Such cooperation has been seen to work under a voluntary agri-environment scheme in Switzerland, where additional payments are available if farmers agree to link up areas managed for biodiversity (Herzog et al., 2005). Alternatively, a more regulatory approach may be seen as necessary if farmer cooperation is unlikely. This study has clearly demonstrated that the aquatic biota of agricultural landscapes could be given a higher level of protection using current agri-environment resources if they were targeted at key locations. It has also highlighted the potentially important role of small water bodies, and in particular ponds, in strategies to protect and enhance aquatic biodiversity because they support high levels of aquatic biodiversity at the landscape scale and are relatively cheap and easy to protect. We, therefore, advocate on-the-ground trials of such an approach accompanied by practical studies to advise on the best methods to protect rivers in agricultural landscapes, which will be particularly important for achieving obligations under the Water Framework Directive. The implementation of a more targeted approach to the protection of freshwater habitats in agricultural landscapes would provide one of the ‘smart solutions’ called for to enable the coexistence of good quality aquatic habitats and farming within the same landscape.
ACKNOWLEDGEMENTS We would like to thank NERC for funding the collection of biological data, Glen Hart and the Ordnance Survey for providing MasterMap and Landform Profile data and Geoff Smith and CEH Monks Wood for Land Cover Map 2000 data. We would also like to thank Hugh Possingham and Carissa Klein for advice on Marxan and the Jonathan and Polly Wood Family Trust for funding the reporting of this research.
REFERENCES ADAS. 2006. Food or water? Eastern farming hangs in the balance. Retrieved on May 2nd 2006, from: www.adas.co.uk/ news/news.html?podlet_id 5 40&article_id 5 149. Related to unpublished briefing Water versus Food, August 2006. Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
446
B. DAVIES ET AL.
Allan JD. 2004. Landscapes and riverscapes: the influence of land use on stream ecosystems. Annual Review of Ecology, Evolution and Systematics 35: 257–284. Allan JD, Erickson DL, Fay J. 1997. The influence of catchment land use on stream integrity across multiple spatial scales. Freshwater Biology 37: 149–161. Arponen A, Heikkinen RK, Thomas CD, Moilanen A. 2005. The value of biodiversity in reserve selection: representation, species weighting and benefit functions. Conservation Biology 19: 2009–2014. Ball IR, Possingham HP. 2000. MARXAN (V1.8.2): Marine Reserve Design Using Spatially Explicit Annealing, a Manual. Available online at: http://www.ecology.uq.edu.au/ index.html?page 5 29780. Biggs J, Williams P, Whitfield M, Nicolet P, Brown C, Hollis J, Arnold D, Pepper T. 2007. The freshwater biota of British agricultural landscapes. Agriculture Ecosystems and Environment 122: 137–148. Briers RA. 2002. Incorporating connectivity into reserve selection procedures. Biological Conservation 103: 77–83. Brown CD, Turner NL, Hollis JM, Bellamy PH, Biggs J, Williams P, Arnold DJ, Pepper T, Maund SJ. 2006. Morphological and physico-chemical properties of British aquatic habitats potentially exposed to pesticides. Agriculture, Ecosystems and Environment 113: 307–319. Carey PD. 2001. Schemes are monitored as effective in the UK. Nature 414: 687. Castelle AJ, Johnson AW, Conolly C. 1994. Wetland and stream buffer requirements: a review. Journal of Environmental Quality 23: 878–882. Chuffney TF, Meador MR, Porter SD, Gurtz ME. 2000. Responses of physical, chemical and biological indicators of water quality to a gradient of agricultural land use in the Yakima River Basin, Washington. Environmental Monitoring and Assessment 64: 259–270. Clover C. 2006. Farmland ‘must go back to the Middle Ages’. Last retrieved on October 4th 2007, from: www.telegraph. co.uk/news/main.jhtml?xml 5 /news/2006/04/18/nfarm18. xml&sSheet 5 /news/2004/04/18/ixhome.html. Collinson N, Biggs J, Corfield A, Hodson MJ, Walker D, Whitfield M, Williams PJ. 1995. Temporary and permanent ponds: an assessment of the effects of drying out on the conservation value of aquatic macroinvertebrate communities. Biological Conservation 74: 125–134. Davies BR, Biggs J, Lee JT, Thompson S. 2004. Identifying optimum locations for new ponds. Aquatic Conservation: Marine and Freshwater Ecosystems 14: 5–24. Davies BR, Biggs J, Williams P, Lee JT, Thompson S. 2008a. A comparison of the catchment sizes of rivers, streams, ponds, ditches and lakes: implications for protecting aquatic biodiversity in an agricultural landscape. Hydrobiologia 597: 7–17. Davies BR, Biggs J, Williams P, Whitfield M, Nicolet P, Sear D, Bray S, Maund S. 2008b. Comparative biodiversity of aquatic habitats in the European agricultural landscape|. Agriculture, Ecosystems and Environment 125: 1–8. De Bie T, Declerck S, De Meester L, Martens K, Brendonck L. 2008. A comparative analysis of cladoceran communities from different water body types: patterns in community composition and diversity. Hydrobiologia 597: 19–27. Declerck S, De Bie T, Ercken D, Hampel H, Schrijvers S, Van Wichelen J, Gillard V, Mandiki R, Losson B, Bauwens D et al. 2006. Ecological characteristics of small farmland ponds: associations with land use practices at multiple spatial scales. Biological Conservation 131: 523–532. Defra. 2005. Entry Level Stewardship Handbook. Downloaded September 2007 from: www.defra.gov.uk/erdp/pdfs/es/elshandbook.pdf. Copyright r 2008 John Wiley & Sons, Ltd.
Dobiasova M, Bacikova S, Wasson J-G, Villeneuve B, GarciaBautista A, Iital A, Timm H. 2004. Land use. In Relationships between Ecological and Chemical Status of Surface Waters. Work Package 4: Rivers, Friberg N (ed). Downloaded on 26 September 2007 from: www.rbmtoolbox.net/docstore/docs/3.1713.D13.pdf. Donohue I, McGarrigle ML, Mills P. 2006. Linking catchment characteristics and water chemistry with the ecological status of Irish rivers. Water Research 40: 91–98. Downing JA, Cole JJ, Middelburg JJ, Striegl RG, Duarte CM, Kortelainen P, Prairie YT, Laube KA. 2008. Sediment organic carbon burial in agriculturally eutrophic impoundments over the last century. Global Biogeochemical Cycles 22: GB1018. Early R, Thomas CD. 2007. Multispecies conservation planning: identifying landscapes for the conservation of viable populations using local and continental species priorities. Journal of Applied Ecology 44: 253–262. Environment Agency. 2006. Concise Register of Gauging Stations. Retrieved 29th March 2006 from www.nwl.ac.uk/ ih/nrfa/station_summaries/op/EA-Thames2.html. Environmental Law Institute. 2003. Conservation Thresholds for Land Use Planners. The Environmental Law Institute: Washington. ESRI. 2001. ESRI Support Centre: Hydro Data Model. Available for download from the World Wide Web at: http:// support.esri.com/index.cfm?fa 5 downloads.dataModels. filteredGateway&dmid 5 15. Fitzpatrick FA, Scudder BC, Lenz BN, Sullivan DJ. 2001. Effects of multi-scale environmental characteristics on agricultural stream biota in Eastern Wisconsin. Journal of the American Water Resources Association 37: 1489–1507. Foley JA, DeFries R. Asner GP, Barford C, Bonan G, Carpenter SR, Chapin FS, Coe MT, Daily GC, Gibbs HK et al. 2005. Global consequences of land use. Science 309: 570–574. Forman RT. 1995. Some general principles of landscape and regional ecology. Landscape Ecology 10: 133–142. Furse M, McDonald I, Abel B. 2003. Coded Checklist of Freshwater Animals of the British Isles: Version 3. Available to download from the World Wide Web at: www.ceh.ac.uk/ subsites/eic/ddc/furselist. Gergel SE, Turner MG, Miller JR, Melack JM, Stanley EH. 2002. Landscape indicators of human impacts to riverine systems. Aquatic Sciences 64: 118–128. Hamaide B, ReVelle CS, Malcom SA. 2006. Biological reserves, rare species and the trade-off between species abundance and species diversity. Ecological Economics 56: 570–583. Herzog F, Dreier S, Hofer G, Marfurt C, Schu¨pbach B, Spiess M, Walter T. 2005. Effect of ecological compensation areas on floristic and breeding bird diversity in Swiss agricultural landscapes. Agriculture, Ecosystems and Environment 108: 189–204. Hickey MBC, Doran B. 2004. A review of the efficiency of buffer strips for the maintenance and enhancement of riparian ecosystems. Water Quality Research Journal of Canada 39: 311–317. Hynes HBN. 1975. The stream and its valley. Verhandlungen der Internationalen Vereinigung fu¨r theoretische und angewandte Limnology 19: 1–15. Johnes P, Moss B, Phillps G. 1994. Lakes — Classification and Monitoring. Environment Agency R&D Note 253. Environment Agency: Bristol. Johnson LB, Richards CR, Host GE, Arthur JW. 1997. Landscape influences on water chemistry in Midwestern stream ecosystems. Freshwater Biology 37: 193–208. Kleijn D, Baquero RA, Clough Y, Dı´ az M, De Esteban J, Ferna´ndez F, Gabriel D, Herzog F, Holzschuh A, Jo¨hl R et al. 2006. Mixed biodiversity benefits of agri-environment Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
MAKING AGRICULTURAL LANDSCAPES MORE SUSTAINABLE FOR FRESHWATER BIODIVERSITY
schemes in five European countries. Ecology Letters 9: 243–254. Kleijn D, Berendse F, Smit R, Gilissen N. 2001. Agrienvironment schemes do not effectively protect biodiversity in Dutch agricultural landscapes. Nature 413: 723–725. Kleijn D, Sutherland WJ. 2003. How effective are European agri-environment schemes in conserving and promoting biodiversity? Journal of Applied Ecology 40: 947–969. Krutz LJ, Senseman SA, Zablotowicz RM, Matocha MA. 2005. Reducing herbicide runoff from agricultural fields with vegetative filter strips: a review. Weed Science 53: 353–367. Linke S, Pressey RL, Bailey RC, Norris RH. 2007. Management options for river conservation planning: condition and conservation re-visited. Freshwater Biology 52: 918–938. Lovell ST, Sullivan WC. 2006. Environmental benefits of conservation buffers in the United States: evidence, promise and open questions. Agriculture, Ecosystems and Environment 112: 249–260. Moss B. 2008. Water pollution by agriculture. Philosophical Transactions of The Royal Society B 363: 659–666. Nicolet P, Biggs J, Fox G, Hodson MJ, Reynolds C, Whitfield M, Williams P. 2004. The wetland plant and macroinvertebrate assemblages of temporary ponds in England and Wales. Biological Conservation 120: 261–278. Orme CDL, Davies RG, Burgess NM, Eigenbrod F, Pickup N, Olson VA, Webster AJ, Ding T-S, Rasmussen PC, Ridgely RS et al. 2005. Global hotspots of species richness are not congruent with endemism or threat. Nature 436: 1016–1019. Palmer M. 1999. The application of biogeographical zonation and biodiversity assessment of the conservation of freshwater habitats in Great Britain. Aquatic Conservation: Marine and Freshwater Ecosystems 9: 179–208. Park Y-S, Grenouillet G, Esperance B, Lek S. 2006. Stream fish assemblages and basin land cover in a river network. Science of the Total Environment 365: 140–153. PCTPR. 2002. A Guide to Monitoring the Ecological Quality of Ponds and Canals using PSYM. Ponds Conservation Trust: Policy and Research: Oxford. Possingham HP, Ball IR, Andelman S. 2000. Mathematical methods for identifying representative reserve networks. In Quantitative Methods for Conservation Biology, Ferson S, Burgman M (eds). Springer-Verlag: New York; 291–305. Pressey RL, Tully SL. 1994. The cost of ad hoc reservation: a case study in western New South Wales. Australian Journal of Ecology 19: 375–384. Preston CD, Pearman DA, Dines TD. 2002. New Atlas of the British and Irish Flora. Oxford University Press: Oxford. Pretty J, Peacock J, Hine R, Sellens M, South N, Griffin M. 2007. Green exercise in the UK countryside: effects on health and psychological well-being, and implications for policy and planning. Journal of Environmental Planning and Management 50: 211–231. Quinn JM, Hickey CW. 1990. The magnitude of the effects of substrate particle size, recent flooding, and catchment development on benthic invertebrates on 88 New Zealand rivers. New Zealand Journal of Marine and Freshwater Research 24: 411–427.
Copyright r 2008 John Wiley & Sons, Ltd.
447
Roth NE, Allan JD, Erickson DL. 1996. Landscape influences on stream biotic integrity assessed at multiple spatial scales. Landscape Ecology 11: 141–156. Roy AH, Rosemond AD, Paul MJ, Leigh DS, Wallace JB. 2003. Stream macroinvertebrate response to catchment urbanisation (Georgia, USA). Freshwater Biology 48: 329–346. Stewart JS, Wang L, Lyons J, Horwatich JA, Bannerman R. 2001. Influences of watershed, riparian-corridor, and reachscale characteristics on aquatic biota in agricultural watersheds. Journal of the American Water Resources Association 37: 1475–1487. Stoate C, Boatman ND, Borralho RJ, Carvalho CR, de Snoo GR, Eden P. 2001. Ecological impacts of arable intensification in Europe. Journal of Environmental Management 63: 337–365. Sutherland WJ. 2004. A blueprint for the countryside. Ibis 146: 230–238. Teels BM, Rewa CA, Myers J. 2006. Aquatic condition response to riparian buffer establishment. Wildlife Society Bulletin 34: 927–935. Thompson S, Larcom A, Lee JT. 1999. Restoring and enhancing rare and threatened habitats under agrienvironment agreements: a case study of the Chiltern Hills Area of Outstanding Natural Beauty, UK. Land Use Policy 16: 93–105. Van Teeffelen AJA, Cabeza M, Moilanen A. 2006. Connectivity, probabilities and persistence: comparing reserve selection strategies. Biodiversity and Conservation 15: 899–919. Virolainen KM, Virola T, Suhonen J, Kuitunen M, Lammi A, Siikamaki P. 1999. Selecting networks of nature reserves: methods do affect the long-term outcome. Biological Sciences 266:1141–1146. Wang LZ, Lyons J, Kanehl P, Gatti R. 1997. Influences of watershed land use on habitat quality and biotic integrity in Wisconsin streams. Fisheries 22: 6–12. Wa¨tzold F, Schwerdtner K. 2005. Why be wasteful when preserving a valuable resource? A review article on the costeffectiveness of European biodiversity conservation policy. Biological Conservation 123: 327–338. Weigel B. 2003. Development of stream macroinvertebrate models that predict watershed local stressors in Wisconsin. Journal of the North American Benthological Society 22: 123–142. Whitfield J. 2006. How green was my subsidy? Nature 439: 908–909. Williams JC, ReVelle CS, Levin SA. 2005. Spatial attributes and reserve design models: a review. Environmental Modelling and Assessment 10: 163–181. Williams P, Whitfield M, Biggs J. 2008. How can we make new ponds more biodiverse? A case study monitored over 7 years. Hydrobiologia 597: 137–148. Williams P, Whitfield M, Biggs J, Bray S, Fox G, Nicolet P, Sear D. 2004. Comparative biodiversity of rivers, streams, ditches and ponds in an agricultural landscape in Southern England. Biological Conservation 115: 329–341. Wilson A, Vickery J, Pendlebury C. 2007. Agri-environment schemes as a tool for reversing declining populations of grassland waders: Mixed benefits from Environmentally Sensitive Areas in England. Biological Conservation 136: 128–135.
Aquatic Conserv: Mar. Freshw. Ecosyst. 19: 439–447 (2009) DOI: 10.1002/aqc
