LUCY R. MALPAS1*, JENNIFER SMART1, ALLAN DREWITT2, ELWYN SHARPS1,3,4and ANGUS GARBUTT3
1Royal Society for the Protection of Birds, UK Headquarters, The Lodge, Sandy, Beds SG19 2DL, UK;2Natural
England, Eastbrook, Shaftesbury Road, Cambridge, CB2 8DR, UK;3Centre for Ecology and Hydrology,
Environment Centre Wales, Deiniol Road, Bangor, Gwynedd LL57 2UW, UK and4Bangor University, School of
Ocean Sciences, Menai Bridge, Anglesey LL59 5AB, UK
Capsule: Over 50% of saltmarsh breeding Common Redshank have been lost since 1985, with current conservation management having only limited success at halting these declines.
Aims: To update population size and trend estimates for saltmarsh-breeding Redshank in Britain, and to determine whether conservation management implemented since 1996 has been successful in influencing grazing intensity and Redshank population trends.
Methods: A repeat national survey of British saltmarsh was conducted in 2011 at sites previously visited in 1985 and 1996. Redshank breeding density and grazing pressure were recorded at all sites; the presence of conservation management was additionally recorded for English sites. Results from all three national surveys were used to update population size and trend estimates, and to investigate changes in grazing pressure and breeding density on sites with and without conservation management.
Results: Of the 21 431 pairs breeding on saltmarsh in 1985, 11 946 pairs remained in 2011, with the highest proportion of this population found in East Anglia. From 1985, British breeding densities
declined at a rate of 1 pair km−2year−1, representing a loss of 52.8% of breeding pairs over 26 years,
although regional trends varied across different time periods. Grazing pressures did not change markedly with conservation management. Redshank declines were less severe on conservation-managed sites in East Anglia and the South of England where grazing pressures remained low, though were more severe on conservation-managed sites in the North West where heavy grazing persisted.
Conclusion: Saltmarsh-breeding Redshank declines continue and are likely to be driven by a lack of suitable nesting habitat. Conservation management schemes and site protection implemented since 1996 appear not to be delivering the grazing pressures and associated habitat conditions required by this species, particularly in the North West of England, though habitat changes may not be linked to unsuitable grazing management in all regions. An in-depth understanding of grazing practices, how conservation management guidelines could be improved, and the likely success of more long-term management solutions is needed urgently.
The Common Redshank Tringa totanus is a species of conservation concern in the UK and Europe, having lost more than a quarter of its population in 25 years, a decline attributed to the loss and degradation of important wetland breeding sites (Tucker & Heath 1994, BirdLife International 2004a, 2004b, Eaton et al. 2009). Great Britain is an internationally important
breeding ground for this species, supporting over 18% of the estimated 100 000–172 000 pairs breeding in northwest Europe (Piersma 1986, Batten et al. 1990, BirdLife International 2004a, 2004b). Redshank breed on lowland wet grassland and upland rough pasture habitats in Britain (Hale 1988), although a nationally and internationally important proportion of the population nests on saltmarsh, with over 45% of breeding pairs found on the coast (Brindley et al.
*Correspondence author. Email: [email protected]
1998). National surveys of British saltmarsh identified a 22.9% reduction in the density of nesting pairs over 11
years, with the estimated saltmarsh breeding
population declining from 21 022 in 1985 to 16 433 pairs in 1996 (Allport et al. 1986, Brindley et al.
1998). Although Redshank populations in the
Wadden Sea – one of the most internationally important saltmarsh sites in continental Europe – have
remained relatively stable or experienced only
moderate declines since 1991 (Koffijberg et al. 2006, Hötker et al. 2007a, 2007b, JMBB 2010), surveys of individual sites in Britain indicate that the decline here may have continued since 1996 (e.g. Smart et al. 2003, Tyas 2008). Whether this decline has continued at a national scale is unknown however.
Reductions in breeding success are known to drive population declines of other wader species (e.g.
Eurasian Oystercatcher Haematopus ostralegus,
Northern Lapwing Vanellus vanellus, Black-tailed Godwit Limosa limosa, Eurasian Curlew Numenius arquata) nesting in both inland and coastal habitats across northwest Europe (Roodbergen et al. 2012). Low breeding success due to a reduction in the availability of saltmarsh nesting habitat in suitable condition was implicated as a key driver of the decline in British Redshank breeding density between 1985 and 1996 (Brindley et al. 1998, Norris et al. 1998). Such reductions in saltmarsh habitat condition are often attributed to environmental changes such as sea level rise and erosion (Woodworth et al. 1999, Natural England 2006), though increases in detrimental management practices (Pye & French 1993) – particularly changes in grazing management (Bakker 1985, Kiehl et al. 1996, Norris et al. 1997, Bouchard et al. 2003, Hannaford et al. 2006) – are likely to have heavily influenced the loss of Redshank pairs between 1985 and 1996. Lower Redshank breeding densities are found on heavily grazed or ungrazed sites compared to those lightly or moderately grazed, and densities declined more severely on sites where grazing intensity increased (Norris et al. 1998). Possible solutions to this problem could therefore involve encouraging the recovery of degraded saltmarsh by mitigating the effects of environmental processes and implementing more suitable grazing management (Davidson et al. 1991, Environment Agency 2005).
Site designation is a key mechanism through which threatened habitats and species may be protected, with the majority of saltmarsh sites in Britain designated as Sites of Special Scientific Interest (SSSIs) and forming part of larger national or international site designations
such as Special Areas of Conservation (SACs), Special Protection Areas (SPAs) and Ramsar sites (Gaston et al. 2006, Natural England 2011). Additionally, management on nature reserves and private land in the form of agri-environment schemes (AES) with saltmarsh-specific options have been implemented since 1996 with the aim of improving saltmarsh habitat condition. These schemes focus largely on the implementation of sympathetic grazing practices, as well as the prevention of surface-damaging activities, cessation of fertilizer inputs and management of accumulated tidal debris (DEFRA 2004, Natural England 2010a). Emerging evidence suggests that
similarly targeted environmental stewardship or
protected area management, implemented on lowland wet grassland habitats in Britain and elsewhere in Europe where Redshank also breed, have had a
positive effect on site occupancy and wader
productivity (Schekkerman et al. 2008, Dillon et al. in prep.). However, little is known about whether similar management has been successful in halting Redshank population declines on saltmarsh.
Traditional saltmarsh management practices vary
greatly by region around the British coast
(Environment Agency 2005), as do saltmarsh habitats and the vegetation communities characterising them (Burd 1989, Natural England 2006). As such, it is likely that the drivers impacting on Redshank populations will vary between regions, and that a
regionally-specific management approach will be
required to facilitate population recovery. In this study we aimed to generate up-to-date national and regional estimates of Redshank breeding population abundance and trends on British saltmarsh by conducting a repeat national survey in 2011 at sites visited in 1985 and
1996. Using grazing pressure and management
information, we then aimed to determine whether management schemes implemented since 1996 have successfully improved grazing practices or effectively influenced Redshank population trends, and whether grazing and Redshank responses to this conservation management differ between regions.
METHODS
Survey and analysis methods used in 2011 matched those used in 1985 and 1996 as far as possible to ensure compatibility between historical and contemporary results. These methods are described in detail by Brindley et al. (1998) and Norris et al. (1998), with key elements and minor changes reported below.
Study sites
Seventy-seven saltmarsh sites were surveyed in 1985, encompassing 2548 ha (about 6%) of vegetated saltmarsh and representing the range and distribution of vegetation types in Great Britain (Allport et al. 1986, Brindley et al. 1998). Of these original sites, 64 were resurveyed in 1996 (83%) and 57 were resurveyed in 2011 (74%). Data from all three years for an additional site (Langstone Harbour, Hampshire, Ordnance Survey grid reference SU693046; Site 78, Fig. 1) became available in 2011, collected using the same survey methods and with Redshank densities
similar to those found elsewhere in the same region. This site was also included in analyses (final site count: 1985 = 78, 1996 = 65, 2011 = 58).
Brindley et al. (1998) and Norris et al. (1998) grouped study sites into nine British regions based on Nature Conservancy Council (NCC) administration areas. We reduced this number to five broader geographic regions that incorporated the boundaries of the original nine (East Anglia [EA], South of England [SE], North West of England [NW], Wales and Scotland; Fig. 1) to better represent the separation between major saltmarsh biophysical types (Burd 1989) and the likely scale at which management solutions will be required. The majority of sites surveyed were SSSIs, with the exception of one site in Wales and seven sites in Scotland which had no statutory protection. Many sites also formed part of large-scale national and international protected areas, with 38, 39 and 34 sites in England (total sites = 43), 13, 14 and 4 sites in Scotland (total sites = 22) and 10, 10 and 13 sites in Wales (total sites = 13) forming part of Ramsar sites, SPAs and SACs respectively.
Where it was sensible to do so, to minimize the variation in habitat and management within study units, sites were split into sub-plots separated by changes in topography (e.g. the presence of major creeks) which represented barriers between different grazing regimes and homogeneous habitat zones (mean no. of plots per site = 1.7, range = 1–5; mean plot area = 31.8 ha, range = 4.7–257 ha; total no. of sites/plots surveyed in 1985 = 130, 1996 = 111, 2011 = 96). In 2011, three teams of two surveyors covered sites in East Anglia (sites 1–13), the South of England (sites 14–21, 27–29 and 78) and the North West of England (sites 43–55); the remaining sites in the South of England, Scotland and Wales were surveyed by experienced volunteers.
Breeding Redshank surveys
Breeding Redshank were surveyed using standard methods developed by Green et al. (1984) and Green (1986) as described in Brindley et al. (1998). All sites were visited three times between mid-April and the end of May with at least 10 days between survey visits, with surveyors walking a standard route developed for each site which approached 100 m of all areas. The timing of surveys was governed by the tide, with access to most sites only possible when tides were low or receding. No surveys were undertaken in rain or winds greater than Gale Force 5. For each site or plot (for larger sites), the mean total number of Redshank counted on survey visits up to 31 May, excluding flocks of more than six individuals and
Figure 1. Locations of saltmarsh sites surveyed in Great Britain in 1985, 1996 and 2011, and the regional divisions used in analyses. All sites were surveyed in 1985 (n = 78); sites 1–8, 10–31, 36, 38– 60, 67 and 69–78 were re-surveyed in 1996 (n = 65); sites 1–22, 24–33, 40, 42–57, 67–74 and 78 were re-surveyed in 2011 (n = 58).
pairs behaving as if they had young, was used as an estimate of the peak number of nesting pairs (Green et al. 1984, Smart et al. 2006). Numbers of pairs at the plot-level were divided by the area of the plot to give plot-level Redshank breeding density estimates (pairs km−2). Site-
level density estimates (pairs km−2) were calculated by
summing the number of pairs across all plots within a site and dividing this total by the area of the site.
Ordnance Survey (OS) maps with hand-drawn site/ plot boundaries were used to record historical survey data, with site areas (ha) estimated by eye from these maps in 1985. To provide blank maps for surveyors in 2011, historical boundaries were digitally redrawn by hand and the site areas (Cartesian ha) estimated within a GIS programme (MapInfo Professional v.6 2000). Although historical site boundaries were digitized as accurately as possible in 2011, there were small differences in the site areas estimated from the 1985 and 2011 maps (paired t-test: mean difference in area = 2.7 ha ±1.17 se, t = 2.31, P < 0.05). Though these small differences are likely to result from the greater accuracy of GIS area estimation compared to visual estimation techniques, a true difference in the area surveyed would affect density estimates if the same area estimate was used in all years. To control for this, the site areas estimated by eye from historical survey maps were used to estimate breeding densities in 1985 and 1996, with 2011 GIS area estimates used to calculate 2011 breeding densities.
Estimating Redshank abundance and conservation status
Regional density estimates were calculated by averaging observed site-level Redshank breeding density (pairs
km−2) across sites in England, Scotland, Wales and
within the three English regions (EA, NW and SE). A bootstrapping procedure was then used to calculate 95% confidence limits around these observed means: breeding densities at sites in each region in each survey year were randomly re-sampled with replacement 999 times, a mean density value was calculated for each of the 999 re-samples in each region/year, and the 2.5% and 97.5% percentiles from the resulting distributions of mean densities used as the 95% confidence limits for the observed mean density for that region/year (Brindley et al. 1998). Estimates for Great Britain were calculated
by summing observed means and bootstrapped
confidence limits from England, Scotland and Wales. Regional density estimates were then divided by the proportion of the area surveyed that was vegetated in
1985 (Redshank nest on vegetated saltmarsh only;
proportion vegetated in Great Britain = 75.93%,
England = 83.51%, Scotland= 88.91%, Wales =
50.17%, EA = 96.86%, NW = 84.25%, SE = 70.87%; Brindley et al. 1998). We found no evidence to suggest that the area of vegetated saltmarsh had changed significantly between survey years on our study plots (1985–1996: t = –0.20, df = 82, P > 0.05; 1985–2011: t = –0.19, df = 73, P > 0.05; 1996–2011: t = 0.51, df = 66, P > 0.05). Finally, these corrected density estimates were multiplied by the total area of vegetated saltmarsh as reported by the 1989 Saltmarsh Survey of Great Britain (Burd 1989) to provide an estimate of the total number of breeding pairs present in each region in each year. The extent of vegetated saltmarsh across England and Wales has not changed substantially since 1989, although individual estuaries have experienced differing levels of accretion/erosion (Phelan et al. 2011).
Sites re-surveyed in Scotland in 1996 had significantly higher densities in 1985 than the sites which were not resurveyed (see the Endnote and Brindley et al. 1998). To prevent this bias resulting in inflated population estimates for Scotland, density estimates expected on sites not resurveyed in 1996 in Scotland were estimated using the linear equation y = 0.378x + 4.273, relating breeding density at all sites surveyed in both 1985 (x) and 1996 (y); this equation differs slightly to that reported in Brindley et al. (1998) due to corrections made to the dataset in 2011 and the addition of Langstone Harbour. Although not all sites in Wales were resurveyed in 1996, and not all sites in Wales or Scotland were resurveyed in 2011, we found no evidence to suggest that these sites had significantly different densities in 1985 to those resurveyed (see Endnote). The majority of sites in England (42 of 43) were resurveyed in both 1996 and 2011.
Trends in regional Redshank populations were determined by calculating the absolute change in site- level breeding density at each site (e.g. 1996 density minus 1985 density) across three time periods (1985– 1996, 1996–2011, 1985–2011), using data from sites surveyed in both years only. The observed mean changes, averaged across all sites within each region, were then calculated and 95% confidence intervals around these means estimated using bootstrapping procedures as above (Brindley et al. 1998). Mean density changes and confidence limits were again corrected for the area of vegetated saltmarsh by dividing by the proportion of the area surveyed that was vegetated in 1985. Density estimates for sites in Scotland not resurveyed in 1996 (based on 1985
densities and the regression equation as above) were not included.
Redshank abundance and conservation status
estimates were recalculated for the 1985 and 1996 survey years and 1985–1996 survey period due to the inclusion of data from the additional site in the South of England (Langstone Harbour).
Grazing and management
We collected basic information on grazing and protected area/AES management during breeding Redshank surveys. Grazing pressure on each site/plot was scored on a 4-point scale in all three survey years (0 = no grazing,
1 = light grazing, 2 = moderate grazing, 3 = heavy
grazing) in relation to the vegetation communities present, the physical evidence of grazing (height of vegetation, hoof prints, faeces) and the number and type of livestock present, following the methods described by Allport et al. (1986). The grazing levels characterizing these categories could be described as: 0 = matted vegetation, no standing crop removed; 1 = majority of standing crop not removed; 2 = majority of standing crop removed; 3 = all standing crop removed, sward height < 10cm (Environment Agency 2005). Although it represents an indirect measure of grazing intensity, this scoring method has been shown to reflect actual livestock activity reasonably well (Norris et al. 1997). As grazing pressure information was recorded in less than 10 sites/plots in Scotland and Wales in 2011, these regions were excluded from region-specific investigations of grazing pressure change.
The inclusion of English sites in AES and the date when such schemes began (all after 1996) were provided by
Natural England (an English Governmental
organization), as were the designation of sites as National Nature Reserves (NNRs) managed by Natural England, or privately managed nature reserves (which may or may not be part of NNRs). Only sites with active Higher Level Stewardship AES agreements (options HP5/HP6:
Maintenance/Restoration of coastal saltmarsh) or
Countryside Stewardship AES agreements (option IT1: Managing inter-tidal habitats) were considered to be implementing AES management relevant to breeding
Redshank ecology (such as sympathetic grazing
management; DEFRA 2004, Natural England 2010a). Privately managed nature reserves either had active AES agreements with options HP5/HP6 or IT1, or were considered likely to be implementing beneficial saltmarsh management similar to that prescribed by these options.
Management on Natural England-managed saltmarsh NNRs is also similar to that prescribed under the AES options HP5/HP6 and IT1 (Natural England 2010b). Due to these similarities, all sites with operational AES agreements or managed as part of NNRs or other nature
reserves were categorized as implementing some
sympathetic conservation management (conservation- managed: ConsMan) compared to sites without any
formal conservation management scheme (no-
ConsMan). No reliable AES or nature reserve management information was available for Scotland or Wales, so these regions were not included in management investigations.
To determine whether grazing management had changed since 1996 across England, regionally and on plots implementing conservation management, we ran a multivariate linear mixed model (LMM) using the lme function in R (package nlme; Pinheiro et al. 2009, R Development Core Team 2009) with plot-level grazing pressure as a continuous response variable. This allowed the inclusion of study plot (PLOT) as a random intercept term to control for the non-independence of repeated grazing pressure measures between survey years on the same site/plot. Models included the independent effects of survey Year (1996, 2011), conservation management (ConsMan, no-ConsMan), Region (EA, NW, SE) and all two- and three-way interactions between these factors. Pairwise contrasts between all category combinations were conducted by changing the reference level of the three-way interaction term (Trebilco et al. 2010, Mueller et al. 2011). To determine whether breeding Redshank had responded to changes in grazing or conservation management, we repeated this modelling process using the number of Redshank breeding pairs per plot (rounded to the nearest pair) as the response variable in a generalized linear mixed model (GLMM) with a Poisson error distribution, and with the natural log of plot area (km2) as an offset term. Models specified in this way return estimates of breeding density
(pairs km−2). We used the function glmmPQL (in R,
package MASS; Venables & Ripley 2002), which accounted for zero-inflation (overdispersion) in the response variable by estimating and incorporating an overdispersion parameter in the model.
RESULTS
Redshank abundance and conservation status We estimate the saltmarsh breeding Redshank population to have been 21 431 pairs in 1985, declining to 17 007
pairs in 1996 and further declining to 11 946 pairs in 2011 (Table 1). Breeding densities were significantly higher in