Habitat loss/degradation on breeding grounds
In recent decades, significant changes to important breeding habitats in Europe have been implicated in the population declines of many birds, including A-P migrants. This is particularly true in agricultural areas, but also in woodland and wetland habitats (Table S2). Key changes in agricultural habitats include the homogenization of the agricultural landscape, the intensification of crop management (e.g. changes in fertilizer, pesticide, grazing and cutting/harvesting regimes) and the loss of marginal, natural or semi-natural habitats (e.g. Donald et al. 2001, Benton et al. 2003, Newton 2004b). The increasing homogeneity of farmland has reduced its suitability as foraging or nesting habitat for several A-P migrants that require a combination of habitat types for breeding or feeding, including the Hoopoe Upupa epops (Barbaro et al. 2008), Eurasian Wryneck Jynx torquilla (Weisshaupt et al. 2011), Common Redstart Phoenicurus phoenicurus (Schaub et al. 2010), Red-backed Shrike Lanius collurio (Brambilla et al. 2010), Lesser Grey Shrike Lanius minor (Wirtitsch et al. 2001) and Yellow Wagtail Motacilla flava (Gilroy et al. 2010).
The intensification of cropped habitats has also impacted several species. For example, the switch from hay to silage with subsequent earlier and more intensive cutting caused severe adult, nest and chick mortality of Corncrakes Crex crex and Whinchat (Green et al. 1997, Grübler et al. 2008), and intensive cultivation of cereal crops and grasslands may have reduced the breeding success of Yellow Wagtails (Bradbury & Bradter 2004, Gilroy et al. 2009).
The loss of marginal, natural or semi-natural habitats may have impacted farmland A-P migrants because they rely on these habitat features for foraging and nesting. For example, Woodchat Shrikes Lanius senator require scattered trees in farmland as nest-sites and perches for hunting (Schaub 1996), and the Ortolan Bunting (Menz et al. 2009), European Roller Coracias garrulus (e.g. Avilés et al. 2000) and Red-backed Shrike require hedges or isolated boundary trees as song posts and/or feeding perches (Brambilla et al. 2007). Similarly, European Turtle Doves require large mature hedges for nesting close to weed-rich habitats for foraging (Browne & Aebischer 2004). Yellow Wagtails nest in crops in arable landscapes, but forage on emergent insects from water-filled boundary ditches (Bradbury & Bradter 2004, Gilroy et al. 2009) and species such as Common Whitethroat forage extensively in invertebrate-rich uncultivated fallows and wildflower areas (e.g. Birrer et al. 2007).
Conversely, in other parts of Europe, abandonment of agricultural land has become an issue for some A-P migrant species. In the northwest Mediterranean region, land abandonment was associated with increases in resident/short-distance migrants, largely woodland species, but declines in A-P migrants, such as Sylvia warblers favouring open semi-natural habitats (Sirami et al. 2007, Fonderflick et al. 2010). For some species, such as Red-backed Shrike, intermediate levels of farming intensity offer the optimal habitat composition, with complete abandonment also likely to cause a severe decline in breeding populations (Brambilla et al. 2007, 2010).
Böhning-Gaese and Oberrath (2003) showed that long-distance migrants, as a group, favour open breeding habitats (e.g. agricultural land) to a greater extent than residents and short-distance migrants, a preference most likely inherited from savannah-breeding African ancestors. This dependence on open breeding habitats by a larger subset of long-distance migrants may predispose them to being disproportionately affected by changes in European agricultural ecosystems.
Changing patterns of woodland management, such as reduction in traditional coppicing practices, and natural processes of succession and recent increases in deer browsing, have meant that young-growth woodland has been increasingly lost throughout much of Europe, with a shift in woodland age structure towards intermediate-aged stands (e.g. Fuller et al. 2007, Hopkins & Kirby 2007). Since many A-P migrants favour early successional habitats to a greater extent than residents (e.g. Helle & Fuller 1988, Fuller & Crick 1992), they may have been particularly sensitive to such changes. In Fennoscandia, intensive forestry has led to the loss of old-growth stands and their replacement with structurally less diverse planted monocultures. The latter may be much less suitable for many bird species, including A-P migrants; they lack a herbaceous understorey required by ground-nesters and may be too dense for aerial-sallying flycatchers (Hausner et al. 2003).
The A-P migrants that breed in freshwater wetlands are threatened by direct habitat loss, particularly due to drainage for agricultural reclamation. The globally threatened Aquatic Warbler Acrocephalus paludicola, for example, has lost 90% of its favoured sedge fen mire breeding habitat in southwest Belarus since 1960 due to drainage (Kozulin & Flade 1999), and loss of Phragmites reed beds from inland wetlands may have contributed to declines in Great Reed Warbler Acrocephalus arundinaceus (Graveland 1998). Commercial harvesting of reeds leads to reduced arthropod abundance compared with unmanaged reed beds (Schmidt et al. 2005), reducing the suitability for a number of species, including the Eurasian Reed Warbler Acrocephalus scirpaceus (Graveland 1999) and Purple Heron Ardea purpurea (Barbraud et al. 2002). Deterioration in water quality has been shown to impact species such as Black Tern Chlidonias niger (e.g. Beintema 1997) and changes in water table management, reduction in standing water and eutrophication have been linked to population trends of resident and migrant marshland birds in The Netherlands (Van Turnhout et al. 2010b).
Drought and habitat change in the non-breeding season
Key wintering and staging areas for many A-P migrants in Africa are the wetlands, savannahs and savannah woodlands of the Sahel zone where ecological conditions are intimately linked to levels of precipitation in the wet season from July to September. If these rains have been good, then food resources will be also be good when migrants arrive. Although rainfall in North Africa has increased since the 1990s (Fontaine et al. 2011), drought conditions predominated during the last three decades of the 20th century (Nicholson 2000), almost certainly causing near-irreversible changes in habitats of this region (Zwarts et al. 2009). The Sahel drought has been frequently linked to declines of many A-P migrants, including those dependent on both seasonal freshwater and terrestrial habitats in these non-breeding areas (e.g. Peach et al. 1991, Zwarts et al. 2009).
In wetland habitats, the abundance of Sedge Warblers Acrocephalus schoenobaenus and Sand Martins Riparia riparia in Britain, for example, varies in relation to rainfall in West Africa. Low rainfall is associated with low annual survival rates and reduced population size in the subsequent breeding season, possibly because drought causes density-dependent overwinter mortality by reducing wetland habitat and hence the carrying capacity of the wintering area (e.g. Peach et al. 1991, Norman & Peach 2013). Other A-P migrants (e.g. Purple Heron, Squacco Heron Ardeola ralloides, Black-crowned Night Heron Nycticorax nycticorax) using seasonal freshwater habitats in West Africa show similar correlations with indices of the wet season (den Held 1981, Zwarts et al. 2009).
A-P migrants exploiting terrestrial habitats that have also been shown to be sensitive to climatic fluctuations, particularly drought in their staging or wintering areas, include Common Whitethroats (e.g. Baillie & Peach 1992), White Stork Ciconia ciconia (Schaub et al. 2005), Barn Swallow Hirundo rustica (Robinson et al. 2008), Lesser Kestrel (Mihoub et al. 2010) and Red-backed Shrike (Pasinelli et al. 2011; Table S2). Some studies have noted a close association between African rainfall and invertebrate abundance (e.g. Dingle & Khamala 1972, Sinclair 1978). For these terrestrial bird species, the impact of rainfall is probably an indirect one, mediated through change in food availability, but in the absence of field research in key wintering and staging areas, the underlying mechanisms remain unclear. In the Nearctic–Neotropical flyway, a strong positive association has been shown between arthropod biomass and winter warbler abundance (Johnson & Sherry 2001). Furthermore, field studies and experimental manipulation of winter food availability for Ovenbirds Seirus aurocapilla in wet and dry years has demonstrated a direct relationship between non-breeding season precipitation, food availability and physical condition of the birds (Strong & Sherry 2000, Brown & Sherry 2006).
Poor conditions on the wintering grounds can also carry over to affect arrival dates and reproductive performance on the breeding grounds. Thus, late arrival of Barn Swallows in years when rainfall and primary productivity are low in their southern African wintering grounds may reflect poor body condition and slower moult in dry years, resulting in delayed departure and, ultimately, reduced productivity (Saino et al. 2004a,b). Similarly, African rainfall is positively correlated with the migration phenology and the eventual breeding success of White Storks in Germany (Dallinga & Schoenmakers 1987, Bairlein & Henneberg 2000). Once again, evidence for such a carryover effect has been demonstrated in the Nearctic–Neotropical flyway; American Redstarts Setophaga ruticilla originating from high-quality tropical winter habitat arrive earlier and have higher reproductive success on the breeding grounds than individuals coming from low-quality winter habitat (Marra et al. 1998, Norris et al. 2004).
Drought conditions have almost certainly compounded the impact of other human-induced habitat changes in vulnerable zones. For example, Sahelian wetlands have been subject to damming, exploitation for irrigated crops, the conversion of natural floodplain woodlands to plantations of exotic species, changes in grazing regimes, and increased hunting activity. These pressures differ across the major wetlands in the Sahel/Sudan zone (Senegal Delta, Inner Niger Delta, Hadejia-Nguru floodplains, Lake Chad Basin and the Sudd region) and thus impact different bird species and populations at different times in different places (Zwarts et al. 2009). Just as these changes are complex, so are the responses of A-P migrants to them. Whereas Ruff and Black-tailed Godwit, which forage in rice crops, apparently benefit from the conversion of floodplains to rice cultivations, Sedge Warblers, which rarely use cultivated rice fields, do not (Zwarts et al. 2009).
Wooded savannah dominated by Acacia species is a major habitat of the Sahel zone (Morel & Morel 1992) and has undergone widespread deterioration from activities such as clearance for agriculture, wood fuel and grazing exacerbated by changes in climate and the prevalence of drought (e.g. Wilson & Cresswell 2006, Yiran et al. 2012). Information about land cover or land cover change in sub-Saharan West Africa from remote sensing earth observation data is patchy and often at too broad a scale to allow assessment of changes that affect migrant distribution. Where information exists, it suggests extensive loss of forest/woodland habitats. In Senegal, for example, the extent of wooded savannahs and forests is estimated to have declined from 78% of the country's land area in 1965 to 72% in 2000 (a loss of c. 33 000 ha per year), with much of the remaining habitat exhibiting moderate to severe degradation from charcoal production (Tappan et al. 2000, 2004). Severe deforestation of Sahelian wooded savannah has also been documented at smaller spatial scales in Senegal (Morel & Betlem 1992, Gonzalez 2001), northeastern Nigeria (Geomatics 1998, Cresswell et al. 2007), northern Ghana (Yiran et al. 2012) and northeastern Somalia (Oduori et al. 2011). The same may be true for Mali, Niger, Burkina Faso and Sudan (Grimmett 1987), and this forest loss and degradation is predicted to continue throughout much of sub-Saharan West Africa (Gaiser et al. 2011, Heubes et al. 2011).
Forest loss and degradation is likely to have a significant negative impact on populations of A-P migrants dependent on them (e.g. Jones 1985, Morel & Morel 1992, Jones et al. 1996, Wilson & Cresswell 2006), although documented responses to habitat change differ between species. Deforestation has been shown to result in decreases in Common Whitethroat and Subalpine Warbler Sylvia cantillans but in increases in Western Bonelli's Warbler Phylloscopus bonelli, Yellow Wagtail (Cresswell et al. 2007), Northern Wheatear (Wilson & Cresswell 2010) and Whinchat (Hulme & Cresswell 2012). Dry season farmland, much of which was open savannah woodland in West Africa, can support good numbers of open country species, such as Whinchat and Wheatear. It has been suggested that Whinchats may even benefit from anthropogenic change on their wintering grounds (Hulme & Cresswell 2012) in the same ways that some agricultural habitats, specifically tree crops such as Citrus spp. and shaded coffee Coffea sp., can be valuable for Nearctic–Neotropical migrants (e.g. American Redstart, Johnson et al. 2006).
Habitat degradation in the Sahel zone may also have impacts on a suite of species for which this is a refuelling stage before or after crossing the Sahara desert. In the absence of field studies in sub-Saharan West Africa, knowledge on where species fatten and the ecological strategies they adopt remains poor. Recent studies suggest A-P migrants may need less fat reserves to cross the Sahara than previously thought (Salewski et al. 2010a,b) and have considerable flexibility in fattening strategy (Dierschke et al. 2005, Delingat et al. 2006), but the balance of evidence still supports the view that spring and autumn passages over the Sahara are periods of extremely high energy demand for many species (Newton 2006, 2008). The refuelling strategies that different species employ on migration (e.g. Bairlein 1985, Schaub & Jenni 2000, Arizaga et al. 2013) will influence the extent to which conditions in the Sahel affect their survival. Some apparently increase their fuel accumulation closer to the border of the Sahara, with the most important pre-migratory fattening sites in spring around the southern edge of the Sahel (e.g. Ottosson et al. 2002, Salewski et al. 2002). Other species may fatten gradually much further south, e.g. Garden Warbler Sylvia borin, Pied Flycatcher and Yellow Wagtail, making them less dependent on a small number of high-quality stopover sites in the Sahel (e.g. Bell 1996, Salewski et al. 2002, Ottosson et al. 2005, Bayly & Rumsey 2010, Jenni-Eiermann et al. 2011).
Similarly, during the southern migration, some species migrating from Portugal to Senegal refuel en route (e.g. Common Grasshopper Warbler Locustella naevia, Bayly et al. 2011), some accumulate fat deposits at staging sites adjacent to the boundary of the Sahara (e.g. Eurasian Reed Warbler and Common Whitethroat, Schaub & Jenni 2000; Garden Warbler, Bairlein 1991), and others do not accumulate appreciable fat deposits at all during their migration and may rely on foraging opportunities within the Sahara itself (e.g. Spotted Flycatcher, Schaub & Jenni 2000).
For some species on northward spring migration, the fruits of shrubs, such as Salvadora persica, are a major dietary component during pre-migratory fattening (Jones 1985, Stoate & Moreby 1995). If fruit and berry abundance in the Sahel are reduced by habitat degradation, this may be reflected in reduced body condition (e.g. Common Whitethroat, Stoate 1995) or may force individuals to refuel earlier and further south, effectively increasing the Sahara crossing distance and so perhaps also the energetic cost (Ottosson et al. 2002, Wilson & Cresswell 2006).
To date, the great majority of studies have highlighted drought and habitat degradation in the Sahel as a key issue. However, the recently suggested declines of A-P migrants wintering in the Guinea forest-savannah and Guinea moist forest zones of Africa (Hewson & Noble 2009, Thaxter et al. 2010) indicate that other factors may now also be operating. Although these species pass through and may refuel in the Sahel, their more recent pattern of decline suggests that factors in the more southern zones may have become increasingly important. Forests in these more southerly zones have also declined in extent, and possibly in quality, for migrant birds (Brink & Eva 2008, Yiran et al. 2012). Agricultural expansion is the most significant cause of deforestation. Remotely sensed land cover data for sub-Saharan Africa suggests a 57% increase in agriculture between 1975 and 2000 at the expense of natural vegetation, a loss of almost 5 million hectares of forest and non-forest natural vegetation per year (Brink & Eva 2008). In countries such as Cote D'Ivoire, Ghana and Nigeria, 83% of what were once dense forests are now forest-agricultural mosaics and the production of the major food crops, such as cassava and plantain, has increased markedly in the humid lowlands (Norris et al. 2010).
Finally, the quality of important staging areas in northern Africa and Europe may also be declining for migrants. For example, refuelling rates of Ruff in grasslands in The Netherlands, the western major staging site, have declined as it has become intensively managed for dairy production. This has been reflected in a global redistribution of breeding Ruffs, with a decline in the western flyway population breeding in Europe and the European Arctic, and an increase in the eastern one, staging in Belarus and breeding in Western and Central Siberia (Verkuil et al. 2012). Similarly A-P migrants staging in southern Morocco during spring northbound migration ultimately rely on good feeding conditions for continuing migration (Maggini & Bairlein 2011, Arizaga et al. 2013).
Effects of climate change on breeding and non-breeding grounds
Bioclimatic models predict a shift in the geographical ranges of A-P migrants, suggesting that their potential future breeding ranges (under climate scenarios for 2070–2099) in Europe might be on average only 89% of their present range, and potential future and present distributions might only overlap by 42% (Huntley et al. 2008). Other studies have suggested that trans-Saharan passerine migrants may suffer winter range reductions in the future, but the current range data for Africa on which these studies are based are generally poor (Barbet-Massin et al. 2009). A study of Sylvia warblers demonstrated that, whereas potential breeding ranges were anticipated to move northwards, potential non-breeding ranges showed no consistent directional shift. Importantly, however, migration distances, and thus energetic costs, between breeding and wintering grounds were projected to increase (Doswald et al. 2009), although in practice a few species have shown the opposite trend (e.g. Barn Swallow, Ambrosini et al. 2011).
Climatic change may also disrupt the synchrony of bird–prey dynamics. Migrants may be particularly vulnerable to phenological mismatch (Both & Visser 2001, Both et al. 2006) because their arrival dates on European breeding sites may be constrained by conditions in non-breeding areas, and there are likely to be fitness consequences of arriving too early or too late. Many studies of spring migration phenology report a greater advancement in short-than long-distance migrants (e.g. Rubolini et al. 2007, Saino et al. 2011), a trend that could cause differences in susceptibility to mismatch (Knudsen et al. 2011). However, some studies have found no difference (Hüppop & Hüppop 2003, Zalakevicius et al. 2006), or even the reverse pattern (Stervander et al. 2005, Jonzén et al. 2006). These inconsistencies may be real or a result of using different methods (Lehikoinen & Sparks 2010), or even sampling effects whereby smaller population sizes lead to apparent later arrival.
The key question, however, is whether these observed shifts in migration and breeding phenology have been sufficient to track changes in resource peaks in breeding areas, usually food, although for Common Cuckoo, host nests may also be important (Møller et al. 2011, but see Douglas et al. 2010). Studies of Pied Flycatchers breeding in The Netherlands (Both & Visser 2001, Both et al. 2006) showed that populations have advanced the onset of egg-laying by as much as 10 days over two decades, but some populations no longer lay synchronously with the window of peak food availability (Both & Visser 2001), with populations nesting late relative to food peaks declining more than those nesting relatively early. Pied Flycatchers may use endogenous responses to environmental cues (e.g. photoperiod) to initiate spring migration in Africa and because these cues do not necessarily reflect conditions on the breeding grounds, this may constrain their capacity to respond adaptively to climate change. Alternatively, it may also be that different climatic trends in breeding and passage areas act as a climatic barrier, similarly disrupting synchrony between timing of breeding and peak food availability (Hüppop & Winkel 2006).
Insectivorous A-P migrant species in The Netherlands declined strongly in forests, a habitat characterized by a short spring food peak, whereas those living in marshes with less seasonal food peaks have declined less. Within these forests, species arriving later in spring declined most, probably because mismatch with the peak food supply was greatest (Both et al. 2010). Furthermore, A-P migrants in forests have declined more severely in Western Europe, where springs have become markedly warmer, compared with northern Europe, where temperatures during spring arrival and breeding have increased less (Both et al. 2010). Mismatch has also been suggested as one of the possible causes of the relationship between timing of migrant declines in the UK and wintering latitude, with species wintering further south tending to exhibit more recent declines (Thaxter et al. 2010, Ockendon et al. 2012).
Overall, however, few studies explicitly address the effects of trophic mismatches on A-P migrants, and the generality of its link to population declines has not been established (Knudsen et al. 2011). For example, in The Netherlands, Pied Flycatcher populations are showing clear evidence of trophic mismatch, whereas further east in Germany they are doing well, and even show evidence of a climate-related increase in productivity (Winkel & Hudde 1997). The advancement of the timing of caterpillar peaks appears to have had no measurable effect on nestling development, productivity or population trends of Wood Warblers in the UK and Poland (Mariarz & Wesolowski 2010, Mallord et al. 2012). Furthermore, broader evidence suggests that A-P migrant population trends are more strongly correlated with migration distance than with phenological mismatch of temperature trends in the breeding and wintering area, although migration distance and phenological mismatch may be linked (Jones & Cresswell 2009, Knudsen et al. 2011).
A slightly different form of phenological mismatch has been suggested to occur in the long-distant migrant Common Cuckoo Cuculus canorus. Although it parasitizes other long-distant migrants, resident and short-distance migrant birds are also hosts and if these advance their lay dates, with increasing spring temperatures, their phenology may be increasingly be poorly matched with that of the Cuckoo (Møller et al. 2011). However, support for changes in host availability as a major driver of Cuckoo declines is mixed (Douglas et al. 2010).
Changes in weather conditions on passage may also affect arrival dates and breeding phenologies (Newton 2010). For example, the median arrival date of a Finnish population of Pied Flycatcher was negatively associated with temperatures in northern Germany, an important passage area. However, although the birds arrived early in response to increasing temperatures in northern Germany, they did not initiate breeding attempts any earlier, probably because temperatures on the breeding ground remained consistent during the study period (Ahola et al. 2004). In years of high productivity in passage sites in the western Mediterranean (indicating higher vegetation production and possibly therefore higher arthropod prey availability), trans-Saharan migrants tended to delay departure. This may be because they trade off early arrival with the certainty of accumulating sufficient energy reserves, or because good conditions allow more poor quality individuals to survive and these tend to have later migration (Robson & Barriocanal 2011).
Climatic change for migrants may also disrupt competitive relationships between resident and migrant bird assemblages or create completely new species interactions (Böhning-Gaese & Bauer 1996, Wiens et al. 2009). It has been suggested that migrants are inferior competitors compared with resident species, and that they are excluded from habitats in which residents occur at high densities (Herrera 1978). It is also possible that higher winter temperatures associated with climate change might improve the overwinter survival of resident species, increasing their numbers in the breeding season and hence the severity of inter-specific competition with migrants. Although, in a limited number of sites, the fraction of the bird community consisting of migrants can be predicted from observed changes in climatic parameters (Lemoine & Böhning-Gaese 2003, Lemoine et al. 2007), there is little empirical evidence demonstrating such antagonistic competition between residents and migrants in Europe. Furthermore, species interactions may vary in relation to the density of potential competitors and switch from positive to negative along environmental gradients (Mönkkönen et al. 2004) such that interactions may, in some cases, be positive (Mönkkönen & Forsman 2005).
Hunting on passage and non-breeding grounds
A-P migrants are widely hunted (shot and trapped) during both spring and autumn passage through the Mediterranean region (e.g. Magnin 1991, McCulloch et al. 1992, Stronach et al. 2002). This is particularly evident in areas of southern France, northern Iberia, Italy, Greece, and Turkey, the islands of Cyprus and Malta, and the Maghreb region of northwest Africa (McCulloch et al. 1992), where large numbers have been reported killed in the past (Magnin 1991). Studies of individual quarry species suggest large numbers may be hunted annually: for example, 116 000 and 205 000 Common Quail Coturnix coturnix in two consecutive autumns (1989 and 1990) in the North Sinai, Egypt (Baha el Din & Salama 1991), and 2–4 million European Turtle Doves across a number of EU countries, a sizeable fraction of the estimated total European population (3.5–7.2 million pairs, BirdLife International 2004, Boutin 2001).
Some species are also hunted on the non-breeding grounds, although the capture and killing of large numbers of birds for consumption in sub-Saharan Africa is relatively restricted to certain species and locations, e.g. terns caught in considerable numbers in coastal West Africa (Everett et al. 1987, Grimmett 1987), and large numbers of White Storks hunted in the Sudan (Grimmett 1987) and Mali (Thauront & Duquet 1991). It has also been suggested that the cause of reduced survival of migrant wildfowl, such as Garganey, in years with little flooding is due to increased hunting pressure as birds become concentrated in a few accessible localities (Zwarts et al. 2009). In the Inner Niger Delta, a hotspot of such activity, the annual catch of Garganey has been estimated as 1800–27 000 birds, an annual loss of 1–15% of the western wintering population (Senegal–Mali) and 20 000-80 000 Ruff, 15–60% of the population wintering there. Catches are higher in dry years when birds concentrate in remaining wetlands and the impact of hunting on survival is difficult to distinguish from that of drought (Zwarts et al. 2009).
The extent of flooding in large deltas such as the Inner Niger has declined in recent years as dams are constructed and water extracted, particularly for crop irrigation. This further concentration of birds in remaining wetlands alongside increased human population pressure may have increased hunting mortality (E. Wymenga pers. comm.). In addition to those species for which there are estimates of numbers, species such as Barn Swallow, European Turtle Dove, European Honey-buzzard Pernis apivorus and Yellow Wagtail are also regularly hunted on their African wintering grounds (Grimmett 1987).
Few studies have shown population-level impacts of this considerable hunting pressure, largely because the relevant data do not exist (e.g. McCulloch et al. 1992), and hence the impact is almost impossible to quantify. Although hunting mortality in autumn may in theory be compensated for by density-dependent reduction in winter mortality, there is far less scope for such compensation for spring hunting mortality. For some key, declining, quarry species, such as Turtle Dove, analyses of survey and demographic data from European breeding grounds may be valuable in establishing the extent to which reduced overwinter survival, perhaps linked to hunting mortality, is a key driver of the decline. However, although the effects of hunting on some species may well be considerable, particularly in the eastern Mediterranean, it is probably not an important driver of declines for a large number of A-P migrants.
Other factors – predation, collision with infrastructure and pesticide use
Nest predation has been implicated in the decline of Neotropical migrants and it has been suggested that, because A-P migrants lay fewer clutches than closely related sedentary congeners, they may be particularly vulnerable to nest predation (e.g. Bruderer & Salewski 2008). However, there are few convincing examples of this, except perhaps the Black-tailed Godwit, although here increased numbers of predators have occurred simultaneously with agriculture-related habitat changes that may have reduced cover and food availability (Schekkerman et al. 2009).
Particularly during the period of migration, collision with human infrastructure (e.g. wind turbines, television masts, electricity pylons, lighthouses) can sometimes be a source of substantial mortality (Newton 2006). The increasing number of wind turbines is a cause for concern and may be an increasing risk for migrating raptors throughout the flyway. However, demonstrating population impacts of collisions with such infrastructure is difficult. There is some evidence that mortality of adult Egyptian Vultures, as a result of collision with wind turbines, may have contributed to declines in European breeding numbers (Carrete et al. 2009), and power line collision accounts for as much as 25% of the White Stork juvenile mortality in Switzerland (Schaub & Pradel 2004).
Finally, pesticide use has long been suggested as a possible driver of declines of A-P migrants (Berthold 1973, Mullie & Keith 1993), particularly the large quantities used over vast areas to control populations of orthopterans, e.g. plagues of Desert Locust Schistocerca gregaria (Dallinga & Schoenmakers 1987, Newton 2004a, Sanchez-Zapata et al. 2007). These impacts are likely to be indirect, e.g. reducing abundance of important prey, rather than resulting in direct mortality, but to date no studies have linked pesticide use to population-level effects (although see White Stork below). In one study, the abundance of Afrotropical birds (not A-P migrants specifically) on plots treated with pesticides (aerial spraying of locust-control pesticides) declined significantly compared with control plots over a period of 3 months (Mullie & Keith 1993), but only a fraction of the observed bird mortality (c. 7%) on the plots could be attributed to direct contamination. The implementation of locust-control measures in Sahelian Africa has been associated with a decline in both the occurrence and the extent of plague events (Rainey et al. 1979, Van Huis et al. 2007). For the White Stork, abundance in Central Europe is higher following winters with high numbers of locusts, compared with summers following poor locust years (Dallinga & Schoenmakers 1987), and it is conceivable that reduced food (locust) availability in Africa may have contributed to the declines of other large trans-Saharan migrants. However, Moreau (1972) suggested that locust peaks are sporadic and irregular, and in the intervening period there is a greater dependency on local solitary orthoptera (e.g. Montagu's Harrier, Trierweiler et al. 2013). Nevertheless, it is also possible that the disruption of natural cycles of the Desert Locust may have cascading effects in the Sahelian ecosystem, causing disruption over several subsequent years (Sanchez-Zapata et al. 2007).