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Eur J Wildl Res DOI 10.1007/s10344-012-0673-8 ORIGINAL PAPER The importance of hunting pressure, habitat preference and life history for population trends of breeding waterbirds in Finland Hannu Pöysä & Jukka Rintala & Aleksi Lehikoinen & Risto A. Väisänen Received: 21 May 2012 / Revised: 18 September 2012 / Accepted: 27 September 2012 # Springer-Verlag Berlin Heidelberg 2012 Abstract Populations of migratory species have undergone dramatic changes in recent decades, but little is known about the factors actually driving those changes. Of particular concern are quarry species such as migratory ducks (Anatidae), many of which have an unfavourable conservation status in Europe. By including both quarry and nonquarry species, as well as habitat preference and life history characteristics of the species, we investigated the relative importance of hunting pressure, both in Finland and at the European level, in explaining population changes of 16 species of migratory waterbirds in Finland during 1986– 2011. Ban of lead shot in 1996 resulted in considerably lower annual hunting bags in Finland thereafter. Species which had the highest hunting pressure had the most negative slopes in population trends from 1986 up to 1997, suggesting that hunting probably limited those populations. However, in general population trends of the species were not strongly associated with hunting pressure in Finland or in Europe. Nor were basic life history characteristics (body mass and clutch size) associated with population trends of the species. In contrast, recent population declines were Communicated by C. Gortázar H. Pöysä (*) Finnish Game and Fisheries Research Institute, Joensuu Game and Fisheries Research, Yliopistokatu 6, FI-80100 Joensuu, Finland e-mail: [email protected] J. Rintala Finnish Game and Fisheries Research Institute, P.O. Box 2, FI-00791 Helsinki, Finland A. Lehikoinen : R. A. Väisänen Finnish Museum of Natural History, University of Helsinki, P.O. Box 17, FI-00014 Helsinki, Finland associated with habitat preferences of the species: those breeding mainly in eutrophic lakes had more negative population trends than those breeding in oligotrophic lakes or generalist species. Reasons for the relatively poor status of species preferring eutrophic lakes probably include overeutrophication of nutrient-rich lakes, resulting in less abundant food resources, and increased nest depredation. Keywords Body mass . Clutch size . Habitat preference . Hunting impact . Migratory species . Waterbirds Introduction Population change in migratory species has been a focus of considerable research during the last few decades. One reason for the increased research activity is the notion that climate change has negatively affected bird migration phenology and breeding performance, which in turn may have contributed to long-term population declines of several species (reviews by Møller et al. 2010; Knudsen et al. 2011). Populations of migrants breeding at northern latitudes have been thought to be under particular threat because individuals that migrate longer distances have to cope with climate change effects in geographically widely separated areas during their annual cycle. While climate change undoubtedly has affected population trends of some bird species, there are other anthropogenic factors that also influence population size, such as hunting. At the extreme end of effects, hunting by prehistoric humans contributed significantly to the extinction of several birds (e.g., Duncan et al. 2002), while more recently, less extreme effects of hunting on bird abundance have been reported frequently (e.g., Newton 1998). Considering all European breeding birds, recent reports imply alarmingly Eur J Wildl Res high total hunting kills in relation to population size for several species listed in Annex II (i.e., hunted species) of the EU Birds Directive (Hirschfeld and Heyd 2005; Mooij 2005). Many of these heavily hunted species have an unfavourable conservation status at the European level (BirdLife International 2004). In this context, it is particularly worrying how little we know about the role of hunting pressure in driving population trends of quarry species in Europe (e.g., Elmberg et al. 2006; Devineau et al. 2010). For example, many Anatidae (ducks and geese) are important quarry species in Europe and elsewhere, yet apart from the information-based management of waterfowl harvest in North America (e.g., Nichols 1991; Nichols et al. 2007), we are aware only of two quantitative analyses addressing the impact of hunting on the status of Anatidae populations. Green (1996) focused on globally threatened Anatidae species and concluded that hunting was an important determinant of the threat category of the species included in the analysis. Long et al. (2007) studied 154 species of Anseriformes (Anatidae and screamers) and did not find hunting to be an important factor influencing the population trends of the species concerned. However, both of these analyses were undertaken at a global scale and were based on a very crude classification of hunting impact (a given species was classified as hunted or not) without considering speciesspecific differences in harvest rate, i.e., how big a proportion of the population was annually taken. However, estimation of total hunting kill for quarry species at the European scale is problematic (see Hirschfeld and Heyd 2005; Mooij 2005), and we urgently need more robust analyses of hunting impacts on Anatidae and other waterbird populations in Europe. In addition, some recent studies suggest that habitat type may also play a role in population declines of migratory species, but their results are divergent. The intensification of agriculture has had negative effects on many farmland bird species (Herzon et al. 2008, 2011), and this intensification has also increased the eutrophication of wetlands in agricultural environments (Stoate et al. 2009). Sanderson et al. (2006) analysed population trends of European breeding birds and found that, in closely related pairs of species, negative trends were more frequent in long-distance migrants than in shortdistance migrants or residents, irrespective of breeding habitat. On the other hand, Both et al. (2010) showed that longdistance migrants breeding in seasonal temperate forests in The Netherlands have declined more than long-distance migrants breeding in less seasonal marshes, while no difference in population trends between habitats was found for residents and short-distance migrants. Cormont et al. (2011) also studied population trends of Dutch breeding birds and showed that species breeding in the same habitat (forest or marshland) may differ in terms of weather-correlated population dynamics depending on life-history traits. Recent analyses of habitat dependence of bird population trends have been based on rather crude habitat categories. For example, in Sanderson et al.’s (2006) study, the category ‘inland wetland’ pools together species breeding in any type of wetland (lakes, ponds, rivers, etc.). This approach certainly is useful for continent-wide analyses incorporating all migratory species but it may mask important differences between species specialized to different types of wetlands. In his review of factors explaining an increase phase of the breeding bird fauna in northern Europe, von Haartman (1973) listed several examples of habitat specialists that increased in eutrophic lakes only; the author mentioned that the avifauna of oligotrophic lakes had suffered simultaneous reductions in numbers. Because divergent trends may occur even within same habitat type (e.g., eutrophic lakes), we need to consider differences in species’ habitat specialisation when analysing population trends. Here, we focus on the importance of hunting pressure and habitat specialisation in explaining long-term (1986–2011) trends in numbers of breeding waterbirds (i.e., ducks, divers, grebes and common coot Fulica atra) in Finland. Standardized monitoring of the numbers of breeding waterbirds in Finland started in 1986 (Koskimies and Väisänen 1991; Pöysä et al. 1993, 2011). For six out of the 16 species included in the present study, the estimated breeding numbers in Finland represent over half of the European breeding population, as defined and reported by Hagemeijer and Blair (1997), and for the other species the proportion of birds breeding in Finland from the total European breeding population is also high (see Hagemeijer and Blair 1997). Hence, the population trends based on the Finnish long-term monitoring data should have significance at the European level. In addition, a considerable proportion of waterbirds harvested each year in Europe are shot in Finland (Hirschfeld and Heyd 2005; Mooij 2005), meaning that changes in waterbird harvest in Finland may bring about considerable changes in overall hunting impact at the European level. We will first describe population changes for 16 species differing in hunting status and breeding habitat specialization. Then we study whether differences between the species’ long-term trends can be explained by differences in hunting pressure and habitat specialization. The species studied here include both protected (i.e., no legal hunting) and quarry species. We use two approaches to assess the impact of hunting. First, we estimate overall harvest rate, both in Finland and at the European level, for each of the quarry species. Using this information, and by also including protected species, we test if there is an association between population trend and hunting pressure. Second, the ban of lead shot in Finland in 1996 resulted in a considerable reduction in waterbird harvest. For example, the mean annual total kill of waterbirds (ducks breeding in inland waters and the common coot) decreased by about Eur J Wildl Res 27 % from the period of 1990–1994 to the period of 1996– 2000, and has remained at a low level since 2000 (see Materials and methods). This magnitude of permanent change in harvest rate sets up a unique opportunity to study the impact of hunting on duck populations. Hence, we test the prediction that, assuming hunting mortality before 1996 has not been completely compensatory (a reasonable assumption, e.g., Smith and Reynolds 1992; Nichols 1991; Pöysä et al. 2004) in Finland, populations of the quarry species should show a more positive or less negative trend after 1996 as compared with the trend before 1996, if hunting is an important factor driving population trends of the quarry species in Finland. Corresponding change in trend should not occur in non-quarry species. Evidence from North America suggests that duck populations respond rapidly to changes in harvest rate (e.g., Patterson 1979; Reynolds and Sauer 1991). As life-history traits such as body size and clutch size potentially are important components in determining species’ vulnerability to over exploitation and other types of threat (e.g., Patterson 1979; Owens and Bennett 2000; Olden et al. 2007), we also examine if long-term population trends of waterbirds are associated with these life-history traits. Materials and methods Fig. 1 Distribution of waterbird census sites totaled according to 20km uniform grid squares Census data The Finnish Game and Fisheries Research Institute (FGFRI, in 1988–2011) and the Finnish Museum of Natural History (FMNH, in 1986–2011) have coordinated monitoring of breeding waterfowl numbers in Finland (Pöysä et al. 1993). In our dataset, the total number of monitored sites is 3,431and the annual mean number of sites containing pair estimates is 740. Point counts are the basic method used in pair surveys, but round counts encircling wetlands are also used (methods explained by Koskimies and Väisänen 1991). The censuses are carried out in May (generally two visits per site to cover both early and late nesting species). Spatial distribution of the census sites is shown in Fig. 1. Trend indices The estimation of annual population change indices of the 16 species (see Table 1; common names are according to BirdLife International 2012) was based on log-linear modelling on site-by-year count data (TRIM; Pannekoek and van Strien 2005) of breeding pairs in 1986–2011. TRIM bases on generalized linear models as the framework of the estimation of model parameters (McCullagh and Nelder 1989), and it is a commonly used tool in European bird monitoring schemes (see www.ebcc.info). In the estimation of standard errors of the slopes of index trends, over-dispersion in counts was controlled for in order to reduce the risk of type I statistical error (rejecting correct null hypothesis). All possible change points were allowed in the description of the annual variations in population change indices. Our main purpose was to compare index trends of the species estimated over the periods 1986–1997 and 1997–2011 (see below). Estimation of hunting pressure Estimates of species-specific hunting bags, especially at the European level (see Hirschfeld and Heyd 2005; Mooij 2005), and population sizes are crude and sensitive to errors. Therefore, instead of using actual harvest rate estimates, we ranked the species in terms of harvest rate, separately at the European and Finnish level, and used the ranks in the analyses as explained below. The most recent and comprehensive estimates of annual bag size for ducks and other waterbird species in Europe are provided by Hirschfeld and Heyd (2005) and Mooij (2005). Species-specific estimates differ between these two sources (discussed by Hirschfeld and Heyd 2005) but in general there is a strong correlation between the estimates for the species included in this study (r00.994, p<0.001, n012 duck species; note that Mooij 2005 does not give bag statistics for Eur J Wildl Res Table 1 Body mass in grams (BO), clutch size (CL), breeding habitat preference (HA) (G generalist, E eutrophic lakes, O oligotrophic lakes) and hunting pressure rank (EU Europe, FI Finland) of study species Species Latin name BO CL HA EU FI Mallard Common teal Eurasian wigeon Common goldeneye Northern Pintail Northern shoveler Garganey Tufted duck Common pochard Common merganser Red-breasted merganser Arctic diver Great crested grebe Red-necked grebe Anas platyrhynchos Anas crecca Anas penelope Bucephala clangula Anas acuta Anas clypeata Anas querquedula Aythya fuligula Aythya ferina Mergus merganser Mergus serrator Gavia arctica Podiceps cristatus Podiceps grisegena 1077 326 699 858 838 635 384 737 921 1438 1091 2389 937 835 8.9 8.7 7.8 9.0 8.1 9.6 9.0 9.7 7.3 9.2 9.1 2.0 4.2 4.0 G G E G E E E E E O O O E E 11 12 8 6 9 10 4 5 7 2 1 0 0 0 11 8 10 6 9 7 12 3 2 5 1 0 0 0 Horned grebe Common coot Podiceps auritus Fulica atra 403 782 4.5 7.7 E E 0 3 0 4 the common coot). We calculated the mean of the values provided by the two sources and used the mean values when estimating harvest rate (estimate of the annual bag for the common coot was taken from Hirschfeld and Heyd 2005). Information about autumn or winter population (hereafter, hunted population) size for each species was derived from BirdLife International (2004) and Delany and Scott (2006). Again, the estimates vary depending on the source. Because of this, and because we do not know to what extent birds originating from different non-breeding ranges (see Delany and Scott 2006) constitute the bag at the European level, we related the estimated annual bags to three different estimates of the hunted population for each species, i.e., the EU25 wintering population estimate from BirdLife International (2004), the non-breeding population estimate for North-west Europe from Delany and Scott (2006) and the non-breeding population estimate for North-west, Central and Eastern Europe from Delany and Scott (2006). For example, the estimated annual bag for the Eurasian wigeon (Anas penelope) was 621,278 individuals (i.e., the mean of the estimates from Hirschfeld and Heyd 2005; Mooij 2005) and the estimated non-breeding population size for North-west Europe was 1,500,000 individuals according to Delany and Scott (2006), resulting in a harvest rate of 41.4 % (621,278/1,500,000×100). The procedure gave us three estimates of harvest rate for each quarry species at the European level. We calculated the mean of the three harvest rate values, ranked the species-specific mean values and used the mean rank in the analyses. In addition, our analysis included four protected species at the European level, representing ‘zero’ hunting pressure (Table 1). Considering harvest rate in Finland, we used annual species-specific bag statistics from the period 2003–2009 compiled by the Finnish Game and Fisheries Research Institute (Finnish Game and Fisheries Research Institute 2009, 2010) and calculated mean annual kill for each quarry species. The period considered is the only period from which the annual kill has been reported for all the species included in this study. Based on wing samples of duck bag from Finland in 1969–1970 (Pirkola and Lindén 1972) and 2005–2007 (Alhainen et al. 2010) the proportions of different species have remained remarkably similar between 1969–1970 and 2005–2007 (r00.985, p<0.001, n09 duck species), implying that relative annual kill of the species has been stable over the past few decades. To estimate the population size at the beginning of the hunting season in Finland (20 August) for each of the quarry species, we multiplied the number of breeding pairs reported in the latest Finnish bird atlas (Väisänen et al. 2011) by 3; this procedure aims to include immature birds in the population (see Delany and Scott 2006). Species-specific estimates of harvest rate in Finland were calculated as explained above at the European level, except that we needed to calculate only one estimate. We ranked the species-specific harvest rates and used the ranks in the analyses. There were four protected species also at the level of Finland, and these species were assigned to ‘zero’ hunting pressure (Table 1). The ban of lead shot in Finland in 1996 (before the 1996 hunting season) resulted in an abrupt reduction in waterbird harvest: the mean annual total kill of waterbirds (ducks breeding in inland waters and the common coot) was estimated at 801,150 individuals (range 734,500–850,850) during 1990– 1994 (Finnish Game and Fisheries Research Institute 1998), while it was only 586,120 individuals (range 526,700– 636,500) during 1996–2000 (Finnish Game and Fisheries Research Institute 2002; note that bag statistics are not available for 1995), i.e., a reduction of about 27 % in total Eur J Wildl Res waterbird harvest. The decrease was consistent across species: the corresponding change for individual species or species group (at that time, annual bags were not reported for all species separately) was 28.6 % for mallard (Anas platyrhynchos), 25.9 % for common goldeneye (Bucephala clangula), 18.8 % for common teal (Anas crecca) and garganey (Anas querquedula) combined, 31.5 % for red-breasted merganser (Mergus serrator) and common merganser (Mergus merganser) combined, and 36.5 % for other species combined. Moreover, except in 1 year since 2000, the annual total kill has been below the mean value for the period 1996–2000 (Finnish Game and Fisheries Research Institute 2007, 2011). Considering the magnitude and consistency of the decrease in waterbird harvest in Finland, we expect it to be reflected in the population trends of the quarry species, if hunting pressure is an important driver of population dynamics of these species. We would like to note that the reduction in hunting mortality may not have been the only positive effect resulting from the ban of lead shot. Indeed, it may also have had direct positive effects on waterbird populations through reduction in lead poisoning (e.g., Pain 1990; Mateo 2009; MartinezHaro et al. 2011). However, it seems unlikely that these effects would override changes in hunting mortality. More important, changes in both mortality factors (i.e., hunting mortality and mortality due to lead poisoning) should have affected trends in the same direction in the present study. Habitat specialization and life history To characterise habitat specialization, we divided the species into three groups: (1) species breeding mainly in eutrophic lakes, (2) species breeding mainly in oligotrophic lakes and (3) generalists which can be found in considerable numbers in both types of lakes. Eutrophic lakes included both naturally eutrophic lakes as well as formerly oligotrophic lakes that have been eutrophicated by agriculture during recent decades. The species-specific classification (Table 1) was based on breeding densities and habitat distribution of the species, as documented in Kauppinen (1993) and Kauppinen and Väisänen (1993), updated with information from Väisänen et al. (1998). The habitat definition (i.e., eutrophic lakes or oligotrophic lakes) used by these authors was based on midsummer coverage of dominant helophytes at the lakes (for more details, see Kauppinen and Väisänen 1993). We used body size (body mass) and clutch size as life history variables (Table 1). Mean body mass values were taken from Cramp and Simmons (1977). Clutch size information of Finnish birds was taken from Solonen (1985). Statistical analyses Clutch size (CL), adult body mass (BO), habitat preference (HA), and Finnish and European hunting pressure (FI and EU, respectively) were used as explanatory variables (Table 1) for the species-specific slopes estimating mean annual population change from 1986 up to 1997 and from 1997 up to 2011 (Fig. 2). Thus, one slope for each period and species was estimated. The period effect was indicated by a factorial variable (PE). For statistical analyses, natural logarithms of the life history (CL and BO) as well as of the hunting (EU and FI; 1 was added to each hunting variable before log transformation) variables were taken. Normally distributed additive slopes, indicating trends in log population index, were estimated using TRIM software (Pannekoek and van Strien 2005); two change-points (1986 and 1997) were set to modelling options. Note that, as an example, the index of 1997 is the result of addition of the slope value of 1996 to the index of 1996. The ban of lead shot took effect in 1996 before the start of hunting season. We set the change-point of population indices at the year 1997, assuming a delayed (1 year) population dynamical response (e.g., Newton 1998) of quarry population recovery to the decreased hunting pressure. Assuming no delayed dynamics, our experimental set-up should be conservative for the effect of hunting pressure during the first period as it includes population census from the spring 1997. In any case, the shifting of the change point by 1 year would not have apparent effects on the slope values (cf. Fig. 2). We used Bayesian mixed modelling and Markov chain Monte Carlo (MCMC) posterior sampling (Brown and Zhou 2010) in the estimation of generalized linear mixed models. Species identifier was set as random variable, controlling for between cluster (species) variation in slopes and the dependency of slopes within the clusters. Fifteen competing models were estimated using all possible combinations of the following factors or factor combinations: (1) CL + BO, (2) EU, (3) HA, and (4) the interaction and main effects of FI × PE. In particular, our aim was to investigate if the ban of lead shot (in 1996) had had effect on slopes; thus, we designed the analysis in order to probe primarily the interaction term (4; i.e., FI × PE). In order to reduce the number of models, we analysed the joint effect of the life history parameters (1). In MCMC posterior sampling, non-informative priors were used. A total of 10,000 burn-in iterations (up to convergence of coefficients) and additive 90,000 iterations were run in order to estimate the posterior distribution of model parameters. Posterior distributions were derived from thinned iteration chains (every 50th iteration recorded). We conducted the Bayesian analyses with the R (R Development Core Team 2012) interfaces glmmBUGS (Brown and Zhou 2010) and R2WinBUGS (Sturtz et al. 2005) (for WinBUGS 1.4.3; Spiegelhalter et al. 2003). For details of MCMC posterior sampling, see, e.g., Gelman et al. (2003) and Lunn et al. (2000; 2009). The differences of deviance information criterion ΔDICi values (Spiegelhalter et al. 2002) were estimated for models (i); DIC weights wi Eur J Wildl Res Common Teal Eurasian Wigeon Common Goldeneye ANACRE 863 ANAPEN 531 BUCCLA 1539 1.1 0.6 0.8 Tufted Duck ANAQUE 47 AYTFUL 547 0.5 1.5 1.0 1.5 2.0 1.0 0.6 0.5 1.0 1.5 2.0 Garganey ANACLY 141 Red−br. Merganser Arctic Diver MERMER 175 MERSER 89 GAVARC 88 1.0 1.1 0.6 0.9 0.7 0.5 0.7 0.9 1.1 1.3 Common Merganser AYTFER 225 1.4 Common Pochard Common Coot PODAUR 57 FULATR 516 1.0 1.0 0.6 0.6 0.9 0.7 0.5 1.0 1.5 2.0 Horned Grebe PODGRI 173 1.4 Red−necked Grebe PODCRI 607 1.4 Great Crested Grebe 1.1 1.3 0.9 0.9 1.2 1.0 1.0 0.8 Northern Shoveler ANAACU 113 1.4 Northern Pintail 1.5 2.5 1.2 Mallard ANAPLA 1074 1990 2010 Fig. 2 Population change indices as estimated with TRIM. Thin line shows index series based on all possible change-points and imputation, which notifies both estimated and observed counts as the basis of index calculus. Bold line expresses model based indices (based on only estimated counts) and allowing two change-points, at 1986 and 1997. The mean of each index series is scaled at one. Horizontal dash line is placed at index values 0.5, 1.0, and 1.5. Numbers after scientific name abbreviations (first three letters from genera and species names, see Table 1) indicate mean number of pairs per study year (Wagenmakers and Farrell 2004; Lunn et al. 2009) were calculated as models in which that factor was included, ji 01 (otherwise, ji 00). expð$DICi =2Þ wi ¼ P expð$DICi =2Þ Results i Over all competing models, ∑wi 01. We used the DIC weights for each model as the indication of how well that model was supported, and model averaging to determine the relative importance of factors (1–4 above) in models by summing, ∑wiji, which is the sum of DIC weights from all The trend analysis based on two change-points at 1986 and 1997 revealed significant long-term trends (Fig. 2, Table 2). Seven out of the 16 species increased and two species (common teal and northern pintail) decreased statistically significantly (t-test, p<0.05) during the period 1986–1997, whereas Eur J Wildl Res Table 2 Additive slopes of indices (slope), their standard errors (SE) and statistical significances (p) based on two change-points at 1986 and 1997 in 16 waterbird species in Finland (Fig. 2) 1986–1996 1997–2011 Species Latin name Slope SE p Slope SE p Mallard Common teal Eurasian wigeon Common goldeneye Northern pintail Northern shoveler Garganey Tufted duck Common pochard Common merganser Anas platyrhynchos Anas crecca Anas penelope Bucephala clangula Anas acuta Anas clypeata Anas querquedula Aythya fuligula Aythya ferina Mergus merganser 0.002 −0.013 −0.001 0.022 −0.039 −0.016 −0.004 0.002 0.038 0.026 0.004 0.004 0.005 0.003 0.010 0.008 0.012 0.006 0.008 0.010 0.636 0.002 0.814 <0.001 <0.001 0.051 0.730 0.747 <0.001 0.007 0.009 −0.002 −0.029 −0.017 −0.033 −0.0002 −0.087 −0.065 −0.121 −0.020 0.003 0.003 0.004 0.002 0.008 0.007 0.013 0.006 0.009 0.007 <0.001 0.605 <0.001 <0.001 <0.001 0.976 <0.001 <0.001 <0.001 0.004 Red-breasted merganser Arctic diver Great crested grebe Red-necked grebe Horned grebe Common coot Mergus serrator Gavia arctica Podiceps cristatus Podiceps grisegena Podiceps auritus Fulica atra 0.046 0.041 0.001 0.075 0.010 0.011 0.015 0.012 0.005 0.007 0.011 0.005 0.002 0.001 0.756 <0.001 0.405 0.042 −0.035 0.003 −0.025 −0.059 −0.070 −0.072 0.012 0.008 0.004 0.007 0.014 0.005 0.003 0.698 <0.001 <0.001 <0.001 <0.001 Here, slope estimates mean annual population change and it is derived from the trend-line of logarithmic indices (indexyear+1 0 slope + indexyear) estimated for each of the two periods. Mean annual multiplicative change is calculated as exp(slope) seven species showed no significant trend (Table 2). During the period 1997–2011, 12 species declined significantly, three species showed no significant trend and the mallard increased significantly. The northern pintail decreased significantly during the first and second period. Six species increased significantly during the first period and then declined significantly during the second period (Table 2). On the basis of the Bayesian analysis and DIC weights, the model that contained only the habitat preference (HA) and the interaction term (FI × PE) with the main effects was superior compared to the other models, and the relative importance of those factors scored highest values (Table 3). Slopes were less positive, indicating more declining or less increasing trends, among the species of eutrophic waters than among the habitatgeneralist species. The slopes among the species preferring oligotrophic waters were not different from those of habitatgeneralist species (Table 4). Slopes tended to be smaller during the second period than during the first period (Table 4, cf. Table 2). Hunting pressure in Finland (FI) had a negative effect on slopes, but the positive interaction term (FI × PE) suggests that the negative effect was compensated for during the second period (Table 4, Fig. 3). This indicates that the ban of lead shot, resulting in a smaller national bag, may have had positive effects on population trends of some species. In particular, the less-hunted species tended to decline more than the more-hunted species during the second period, i.e. being opposite to the pattern observed in the first period (Fig. 3). Discussion The findings of this study indicate that the breeding populations of seven out of the 16 species examined increased during the first period (1986–1997), but three quarters of the species declined significantly during the second period (1997–2011). The role of hunting The population trends during the first period (1986–1997) and alarming population declines in many waterbird species during the second period (1997–2011) were not associated with the average hunting pressure in Europe. However, hunting in Finland seems to have affected the population trends during the first period, i.e., before the decrease in hunting bag size due to the ban of lead shot. Species that had the highest hunting pressure had the lowest slopes of population trend during the first period, suggesting that hunting probably limited those populations at that time. After the ban of lead shot, such an effect was no longer observed. We would like to note, however, that the drastic change in population trend (from positive to negative slope) from the first to the second period in some of the non-hunted species (e.g., red-necked grebe Podiceps grisegena) and in species with relatively low hunting pressure (e.g., red-breasted merganser and common pochard Aythya ferina) (see Table 1 and Eur J Wildl Res Table 3 Model weights (wi) and averaging based on deviance information criterion (DIC) The lower/higher DIC/w, the more supported model CL clutch size, BO body mass, EU European hunting pressure, FI Finnish hunting pressure, HA habitat preference, PE study period; for further information see Materials and methods Model (i) Factor1 Factor2 Factor3 1 2 3 4 5 6 7 8 9 10 11 12 13 14 CL + BO CL + BO CL + BO CL + BO EU EU HA 15 Relative importance CL + BO 0.196 EU 0.259 HA 0.857 Factor4 EU HA FI × PE CL + BO CL + BO CL + BO EU HA FI × PE EU EU EU Fig. 2) have had a strong effect on the interaction between hunting pressure and period (i.e., the FI × PE term). The change to a positive direction from the first to the second period was relatively small or did not exist in species ranked as having relatively high hunting pressure (e.g., garganey, mallard, Eurasian wigeon and common teal; Table 2). Thus we suggest that, in general, hunting in Europe and in Finland is unlikely to have been the major driver of the observed recent large-scale population declines of breeding waterbirds in Finland. HA FI FI FI FI FI HA HA HA × PE × PE × PE × PE × PE FI × PE 1.000 DICi ΔDICi wi −106.5 −106.6 −107.8 −124.2 −104.1 −104.7 −123.9 −106.5 −122.7 −128.8 −126.7 −125.5 −121.8 −102.7 22.3 22.2 20.9 4.6 24.7 24.1 4.9 22.3 6.1 0.0 2.1 3.3 6.9 26.1 7.75E−06 8.24E−06 1.52E−05 0.05 2.35E−06 3.18E−06 0.05 7.76E−06 0.03 0.54 0.19 0.10 0.02 1.15E−06 −123.0 5.8 0.03 Our indices of hunting pressure at the European level were based on incomplete bag statistics. For example information on bag statistics from European countries outside the EU, notably Russia, was missing (see Hirschfeld and Heyd 2005; Mooij 2005). Also, we were not able to examine the hunting pressure outside Europe, and yet hunting in Africa could be significant for some species. However, only two of the study species, garganey and northern pintail, are mainly wintering in Africa, and thus hunting outside Europe cannot explain the generally declining pattern. Interestingly, both of Table 4 Bayesian analysis based on MCMC posterior sampling (A) (B) Parameter Mean SD 2.5 % 50 % 97.5 % BO CL HA_oligotrophic HA_eutrophic (0a) FI (0b) EU PE_97_11 (0c) FI × PE_97_11 (0d) HA_oligotrophic 0.0115 −0.0076 −0.0031 −0.0094 −0.0302 −0.0132 −0.0864 0.0248 0.0004 0.0184 0.0251 0.0153 0.0283 0.0178 0.0159 0.0196 0.0114 0.0209 −0.0245 −0.0577 −0.0339 −0.0652 −0.0649 −0.0456 −0.1250 0.0024 −0.0413 0.0116 −0.0077 −0.0030 −0.0096 −0.0300 −0.0131 −0.0864 0.0248 0.0007 0.0475 0.0425 0.0260 0.0459 0.0059 0.0187 −0.0480 0.0476 0.0412 HA_eutrophic (0a) FI (0b) PE_97_11 (0c) FI × PE_97_11 (0d) −0.0306 −0.0197 −0.0863 0.0249 0.0158 0.0086 0.0186 0.0108 −0.0620 −0.0372 −0.1230 0.0035 −0.0307 −0.0196 −0.0862 0.0250 0.0012 −0.0033 −0.0498 0.0460 Degree of belief Pr.(a<0)00.954 Pr.(b<0)00.816 Pr.(c<0)01.000 Pr.(d>0)00.985 Pr.(a<0)00.970 Pr.(b<0)00.988 Pr.(c<0)01.000 Pr.(d>0)00.989 Posterior mean, standard deviation and limits from lower to upper 95 % credible interval. Degree of belief is the probability that a hypothesis is true. Posteriors reported for (A) the model with all parameters, and (B) the most supported model by DIC (Table 3) CL clutch size, BO body mass, EU European hunting pressure, FI Finnish hunting pressure, HA habitat preference, PE study period; for further information, see Materials and methods 0.02 0.06 −0.04 Slope (1986−1996) Eur J Wildl Res 1.0 1.5 2.0 2.5 −0.06 0.00 0.5 The role of life history parameters and habitat type −0.12 Slope (1997−2011) 0.0 considered in the present study, despite legislative requirements for the collection of such data (e.g., in the African– Eurasian Migratory Waterbird Agreement). Reasonably long series on annual breeding numbers, reproductive output and bag statistics in Finland are presently available for mallard, common teal and common goldeneye, and we will address the effects of hunting pressure on breeding numbers of these species in more detail elsewhere (Pöysä et al., in preparation). 0.0 0.5 1.0 1.5 2.0 2.5 Hunting pressure (Finland) Fig. 3 Effect of hunting pressure (Finland) on species specific slopes of 1986–1996 (upper panel) and 1997–2011 (lower panel) (Table 2) based on Bayesian analysis. Mean of estimated slopes (open circles) with Bayesian 95 % confidence intervals (credible intervals). Dots denote observed slopes. Regression line is drawn on the basis of estimated mean values. Hunting pressure values of 0 have been slightly random jittered in order to help distinction between species. Hunting pressure values have been transformed as log(FI + 1) (for original values, see Table 1) the trans-African migrants did show a rapid population decline during the second period, and similar declines have been observed in many other long-distance migrants during recent decades (Sanderson et al. 2006; Brommer 2008). Thus, the decline of garganey and northern pintail could be partly linked with the general trend of deteriorating conditions along the migration route and/or on wintering grounds (Sanderson et al. 2006; Pearce-Higgins et al. 2009). On the other hand, Rendón et al. (2008) did not find any long-term changes (from 1978 to 2005) in the numbers of northern pintail wintering in Doñana, south-west Spain. From the other six Anatidae species included in both Rendón et al. (2008) and this study only the common teal and Eurasian wigeon showed a long-term decline in wintering numbers at Doñana. Due to limited sample sizes, our analysis omitted geese which generally have been increasing in Finland, except the bean goose Anser fabalis whose population is most probably declining (Valkama et al. 2011). Hunting pressure towards the bean goose is high in Finland (Finnish Game and Fisheries Research Institute 2011) especially considering its low population size (Väisänen et al. 2011). Finally, whilst our study is the first quantitative analysis of hunting impacts on waterbird populations in Europe, our study design has shortcomings due to limitations in data availability. A more powerful approach for exploring the effects of hunting pressure might be to study directly the effect of annual harvest rate on annual changes in population size (e.g., Nichols 1991; Reynolds and Sauer 1991). Unfortunately, such data are not available at the European level for any of the species Neither the clutch size nor the body mass of the study species were associated with the observed population trends. However, habitat type had a clear effect: species that prefer eutrophic lakes had generally declined more as compared with generalist species or species preferring oligotrophic lakes. This finding is in line with that observed in wetlands classified as Important Bird Areas (IBA) in Finland. The conservation value of several important wetlands has been decreasing significantly since the early 1980s, but the decline has been much weaker in areas where notable improvements to management actions have been made (Ellermaa and Lindén 2011). Since the areas from which the present data come are not part of the Finnish IBA network, our findings, together with those of Ellermaa and Lindén (2011), suggest that the status of breeding waterbirds in eutrophic wetlands in Finland has generally been impaired. Several reasons have been proposed to explain the declining trend of waterbirds in eutrophic lakes. One of the suggested causes is further eutrophication of the water systems (Ellermaa and Lindén 2011) that has been going on for decades (Kauppi et al. 1993; Simola et al. 1995; Ekholm and Mitikka 2006). Such development should have been more pronounced in naturally eutrophic lakes and hence affected especially those species that prefer those habitats. Eutrophication has been suggested to cause population declines of common goldeneye, common coot and velvet scoter Melanitta fusca in the Archipelago Sea in SW Finland (Rönkä et al. 2005), but its effects on waterbird populations has not been directly examined at inland lakes. Most of the eutrophic lakes in Finland are situated on lowlands in the southern part of the country, where agriculture is also at its most intensive (von Haartman 1973). Ekholm and Mitikka (2006) studied 20 lakes situated in agricultural landscapes in south and central Finland and found six of them to be eutrophic and 14 to be hypertrophic. Over half of the same lakes also showed signs of increasing eutrophication during 1976–2002, supporting the idea that nutrient inputs from agriculture have also continued during our study period. Eutrophication has been shown to cause increases of phytoplankton, turbidity and cyprinid populations (Olin et al. 2002; Ekholm and Mitikka 2006), which all Eur J Wildl Res can affect negatively the abundance of underwater vegetation and invertebrates (Giles 1994; Hanson and Butler 1994). Hence, strong eutrophication can decrease food availability for waterbirds, which may lead to population decline. Indeed, food abundance has been shown to be the key factor determining species number and density of waterfowl communities in Northern Europe (Elmberg et al. 1993). On the other hand, eutrophication of nutrient-poor oligothrophic lakes may improve breeding opportunities for dabbling ducks, such as the mallard (Sjöberg et al. 2000), that prefer moderately eutrophic wetlands. Interestingly, Hilli-Lukkarinen et al. (2011) found that the decrease of agricultural areas around their study lakes in central Finland from 1955–56 to 2002–2003 did not have clear effects on waterfowl populations. It is also important to keep in mind that climate change may enhance eutrophication. In Northern Europe, warming climate will likely increase rainfall especially during winter time (Jylhä et al. 2004), which increases nutrient flow to the water systems from the catchment area (Meier et al. 2012). This highlights, that both direct and indirect consequences of climate change on waterbird populations should be examined in the near future. Additional reasons for declining waterfowl numbers may be increasing abundance of invasive predators (Mikkola-Roos et al. 2010) such as the raccoon dog Nyctereutes procyonoides and American mink Mustela vison (Nordström et al. 2002, 2003; Väänänen et al. 2007) and the decline of a keystone species, the black-headed gull Larus ridibundus (MikkolaRoos et al. 2010; Väänänen 2011). Indeed, several duck species are known to have higher breeding densities in blackheaded gull colonies (Cramp and Simmons 1977; Cramp 1985) and there is experimental evidence that the breeding colonies provide safety from nest predators (Väänänen 2000). Increased nest predation may have decreased the breeding success of waterbirds and resulted in population declines. However, the gull population has been declining since the 1970s (Valkama et al. 2011), whereas the majority of the waterbird populations have shown negative trends only in more recent years. While the decrease of gull populations mostly has been gradual and lags of several years may occur in the response of duck populations to changes in predation pressure, these findings suggest that other factors are behind the overall decline of waterbirds in Finland. Finally, changes in fish–duck interactions may be one of the potential causes of waterbird population declines, as fish may affect waterbirds through both competitive and predatory effects (e.g., Elmberg et al. 2010; Nummi et al. 2012; Väänänen et al. 2012 and references therein). Conclusions We examined the importance of hunting pressure in explaining long-term trends of breeding numbers of waterbirds in Finland by analyzing simultaneously population trends of quarry and non-quarry species. Our findings indicate that hunting pressure in Finland may have limited the abundance of some species during the period of high hunting pressure, though the evidence was not strong. However, in general the present analysis indicates more strongly that hunting pressure in Finland or in Europe does not explain the general declining trend of most breeding waterbirds during the last 15 years. These declines were most severe in species that prefer eutrophic lakes for breeding. One potential reason for this is over-eutrophication of naturally nutrient-rich lakes, but the pattern should be studied in more detail. We encourage further investigation into the effects of exploitation on population sizes of quarry species and into the connection between water quality and waterbird populations. To that end, we urgently need more accurate and up-to-date hunting bag data from all countries along the main European flyways of waterfowl. 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