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Captive breeding and reintroduction of the Lesser kestrel Falco naumanni: a genetic analysis using microsatellites

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We used microsatellites to assess ongoing captive breeding and reintroduction programs of the lesser kestrel. The extent of genetic variation within the captive populations analysed did not differ significantly from that reported in wild populations. Thus, the application of widely recommended management practices, such as the registration of crosses between individuals in proper stud books and the introduction of new individuals into the genetic pools, has proven satisfactory to maintain high levels of genetic variation. The high rates of hatching failure occasionally documented in captivity can therefore not be attributed to depressed genetic variation. Even though genetic diversity in reintroduced populations did not differ significantly when compared to wild populations either, average observed heterozygosities and inbreeding coefficients were significantly lower and higher, respectively, when compared to the captive demes where released birds came. Monitoring of reproductive parameters during single-pairing breeding and paternity inference within colonial facilities revealed large variations in breeding success between reproductive adults. The relative number of breeding pairs that contributed to a large part of captive-born birds (50–56%) was similar in both cases (28.6 and 26.9%, respectively). Thus, the chances for polygyny (rarely in this study), extra-pair paternity (not found in this study) and/or social stimulation of breeding parameters do not seem to greatly affect the genetically effective population size. Independently of breeding strategies, the release of unrelated fledglings into the same area and the promotion of immigration should be mandatory to counteract founder effects and avoid inbreeding in reintroduced populations of lesser kestrels.
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SHORT COMMUNICATION
Captive breeding and reintroduction of the lesser kestrel
Falco naumanni: a genetic analysis using microsatellites
Miguel Alcaide Æ Juan J. Negro Æ David Serrano Æ
Jose
´
L. Antolı
´
n Æ Sara Casado Æ Manel Pomarol
Received: 2 February 2008 / Accepted: 15 January 2009 / Published online: 7 February 2009
Ó Springer Science+Business Media B.V. 2009
Abstract We used microsatellites to assess ongoing
captive breeding and reintroduction programs of the lesser
kestrel. The extent of genetic variation within the captive
populations analysed did not differ significantly from that
reported in wild populations. Thus, the application of
widely recommended management practices, such as the
registration of crosses between individuals in proper stud
books and the introduction of new individuals into the
genetic pools, has proven satisfactory to maintain high
levels of genetic variation. The high rates of hatching
failure occasionally documented in captivity can therefore
not be attributed to depressed genetic variation. Even
though genetic diversity in reintroduced populations did
not differ significantly when compared to wild populations
either, average observed heterozygosities and inbreeding
coefficients were significantly lower and higher, respec-
tively, when compared to the captive demes where released
birds came. Monitoring of reproductive parameters during
single-pairing breeding and paternity inference within
colonial facilities revealed large variations in breeding
success between reproductive adults. The relative number
of breeding pairs that contributed to a large part of captive-
born birds (50–56%) was similar in both cases (28.6 and
26.9%, respectively). Thus, the chances for polygyny
(rarely in this study), extra-pair paternity (not found in this
study) and/or social stimulation of breeding parameters do
not seem to greatly affect the genetically effective popu-
lation size. Independently of breeding strategies, the
release of unrelated fledglings into the same area and the
promotion of immigration should be mandatory to coun-
teract founder effects and avoid inbreeding in reintroduced
populations of lesser kestrels.
Keywords Genetic diversity Conservation genetics
Effective population size Founder effect
Mixed reproductive strategies
Introduction
Captive breeding of endangered species has become a
widespread practice to provide individuals for reintroduc-
tion or supplementation programs for extinct or declining
populations. Although traditional approaches have tried to
identify ecological and behavioural constraints affecting
the short-term success of these initiatives (e.g. Hirzel et al.
2004; Martı
´
nez-Meyer et al. 2006), most monitoring pro-
grams do not take full advantage of the potential afforded
by molecular markers. Monitoring population genetic
metrics can provide insights into relevant processes that are
difficult or impossible to study via traditional approaches
(e.g. Schwartz et al. 2006). For example, captive breeding
and reintroduction programs could potentially be counter-
productive if the genetic consequences of the various
M. Alcaide (&) J. J. Negro D. Serrano
Estacio
´
n Biolo
´
gica de Don
˜
ana (CSIC), Pabello
´
n de Peru
´
,
Avda. Ma Luisa s/n, 41013 Sevilla, Spain
e-mail: malcaide@ebd.csic.es
J. L. Antolı
´
n
DEMA, Defensa y Estudio del Medio Ambiente,
Crı
´
a del cernı
´
calo primilla, Ctra. Fuente del Maestre, km 17,
06200 Almendralejo, Spain
S. Casado
GREFA, Grupo de Rehabilitacio
´
n de la Fauna Auto
´
ctona y su
Ha
´
bitat, Apdo. 11, 28220 Majadahonda, Spain
M. Pomarol
Centro de Recuperacio
´
n de Torreferrusa, Servici Protecio
´
Gestio
´
Fauna. Carretera Sabadell-Sta, Perpetua de la Moguda,
Km. 4,5, 08037 Barcelona, Spain
123
Conserv Genet (2010) 11:331–338
DOI 10.1007/s10592-009-9810-7
management options are not fully considered (Woodworth
et al. 2002; Gilligan and Frankham 2003). In this respect,
loss of genetic variation linked to founder effects and
inbreeding may have serious fitness consequences and can
jeopardize the evolutionary and adaptive potential of
populations and species (Frankham et al. 2002).
The lesser kestrel Falco naumanni was one of the most
abundant raptors in Europe before a sharp population
decline which began in the late 1960 s (Bijleveld 1974). As
a result, this small migratory and facultatively colonial
falcon totally or partially disappeared from several loca-
tions of its former breeding range (Biber 1990), and is now
patchily distributed from Portugal to China (Cramp and
Simmons 1980). To date, numerous captive breeding pro-
grams have successfully contributed to the reinforcement
and re-establishment of decimated or extinct populations in
Western Europe (e.g. Pomarol 1993) by using the method
of hacking (Sherrod et al. 1981).
In this study, we have performed the first genetic
assessment of ongoing captive breeding and reintroduction
programs of the globally vulnerable lesser kestrel (BirdLife
International 2004). Firstly, we investigated levels of
genetic diversity in captive populations. Hatching failure,
one of the most cited fitness consequences of inbreeding in
birds (e.g. Keller 1998; Morrow et al. 2002), has been
occasionally high in captivity in lesser kestrels ([50% of
fertile eggs; Cola
´
s et al. 2002), contrasting with the normal
values of this parameter in the wild (\10% of fertile eggs,
e.g. Serrano et al. 2005). In fact, hatching success in cap-
tivity is the only parameter which has not exceeded the
performance of the species in the wild (Pomarol et al.
2004a).
Secondly, we compared single-pairing (one male and
one female) versus colonial captive breeding (multiple
males and females) strategies. We focused on variations in
breeding success as primary determinants of genetically
effective population size (e.g. Nunney and Elam 1994;
Hedrick 2005). To this aim, we calculated the minimum
number of breeding pairs that contributed to a high pro-
portion of fledglings at two captive centres working on
single-pairing into individual pens. Paternity of fledglings
within colonial enclosures can only be confirmed through
genetic inference, and therefore, we employed polymor-
phic microsatellites to infer kinship. Two hypotheses can
be made in this respect. The first hypothesis would predict
an increase of the variance in male breeding success
because of mixed reproductive strategies such as those
observed, although at low rates, in wild colonies [see
exceptional polygynous mating systems in Tella et al.
(1996) and low extra-pair paternity rates \7.5% in Alcaide
et al. (2005)]. Alternatively, the simulation of colonial
environments may stimulate the breeding behaviour of
individuals which could otherwise remain sexually inactive
(see for instance Waas et al. 2005), with the subsequent
increase in overall productivity compared to single-breed-
ing pairs.
Finally, we evaluated the extent of genetic variation that
has been successfully transmitted from captive stocks to
reintroduced populations to help optimize the main genetic
goal of a reintroduction program. In this respect, it is
widely assumed that high levels of genetic diversity max-
imize the possibilities of re-establishing a self-sustaining
population in the long term (e.g. Ballou and Lacy 1995;
Frankham et al. 2002).
Materials and methods
Captive, reintroduced and wild populations
In Spain, three captive populations kept by non-government
organizations for Native fauna and its Habitat Rehabilitation
‘‘ GR E FA ’’ ( www.grefa.org), Defence and Study of Natural
Environment ‘DEMA’ (www.demaprimilla.org)andthewild-
life recovery center of Torreferrusa attached to the Cata-
lonian government ‘TORREF’ (http://mediambient.
gencat.cat/cat/el_medi/fauna/fauna_auctoctona/centres/torrefe
rrussa.jsp) were investigated (see Fig. 1). Founder indi-
viduals of captive demes, usually injured birds which
could not be rehabilitated and returned into the wild, were
derived from different locations belonging to the main
Spanish population or translocated from other captive
populations. Management actions of breeders encompass
the registration of crosses between individuals in proper
stud books and the introduction of new individuals into
Fig. 1 Breeding distribution of the lesser kestrel in Western Europe.
Reintroduced (black asterisks) and captive (white asterisks) popula-
tions investigated in this study are indicated. See Table 1 for codes
332 Conserv Genet (2010) 11:331–338
123
the genetic pools to avoid inbreeding. The proportion of
birds which annually die (about 5%) is easily replaced
given that this option is not constrained by the number of
lesser kestrels available in the study area (see Table 1;
Pomaroletal.2004a for more details). To date, different
captive stocks have contributed to several reintroduc-
tion and reinforcement programs in Spain and France
(e.g. Pomarol et al. 2004a, http://crecerellette.lpo.fr/life/
life.html).
Three reintroduced populations of lesser kestrels (Lleida
and Gerona in Catalonia plus La Rioja, Fig. 1) were also
investigated. The lesser kestrel disappeared from Catalonia
(North Eastern Spain) as a breeding species in 1986. A
reintroduction program beginning in 1989 has led to a
population distributed in two main nuclei (Gerona and
Lleida) which was estimated at 94 breeding pairs in 2003
(see Pomarol et al. 2004b). The lesser kestrel also disap-
peared from La Rioja (Central Northern Spain) around the
second half of the XX century. After an evaluation of
habitat suitability for the reintroduction of the species, the
first colony was founded in 1997 by the release by hacking
and subsequent return after migration of captive-born birds
(Lopo et al. 2004 for more details). The population size of
this colony was estimated at 13 breeding pairs by 2003.
Finally, four geographically distinct natural populations
(Southern France, Ebro Valley, Spanish core area and
Portugal) were analysed to provide comparative data (see
Table 1; Fig. 1).
Sampling and DNA extraction
Biological samples for genetic analyses were obtained
from wild and reintroduced populations during the 2002
and 2003 breeding seasons. Only one nestling per brood
was analysed to minimize the sampling of related indi-
viduals. In 2004, we sampled the breeding stocks of DEMA
and GREFA (see Table 1) as well as the captive-born
progeny produced at the two largest colonial pens of
DEMA (N = 96 nestlings).
The DNA extraction protocol we used follows that
described by Gemmell and Akiyama (1996). Blood and
feathers tips were digested by incubating with proteinase K
in 300 ll of a buffered solution for at least 3 h. Proteins were
selectively discarded by adding 1 volume of a 5 M LiCl
solution and two volumes of chloroform–isoamylic alcohol
(24:1). After centrifugation at maximum speed, DNA was
precipitated using two volumes of absolute ethanol. Pellets
thus obtained were dried and washed twice with 70% etha-
nol, and later stored at -20°C in 0.1 ml of TE buffer.
Microsatellite genotyping
We amplified nine microsatellite markers originally iso-
lated from the peregrine falcon Falco peregrinus (Fp5,
Fp13, Fp31, Fp46-1, Fp79-4, Fp89, Fp107, see Nesje et al.
2000; Cl347 and Cl58, see Alcaide et al. 2008a). For each
locus, the polymerase chain reaction (PCR) was carried out
in a PTC-100 Programmable Thermal Controller (MJ
Research Inc., Waltham, MA, USA) using the following
PCR profile: 35 cycles of 40 s at 94°C, 40 s at 55°C, 40 s
at 72°C and finally, 4 min at 72°C. Each 11 ll reaction
contained 0.2 units of Taq polymerase (Bioline, London,
UK), 19 manufacturer-supplied buffer, 1.5 mM MgCl
2
,
0.02% gelatine, 0.12 mM of each dNTP, 5 pmol of each
primer and, approximately, 10 ng of genomic DNA. For-
ward primers were 5
0
-end labelled with HEX, NED or
6-FAM fluorocroms. Amplified fragments were resolved
on an ABI Prism 3100 Genetic Analyser and later scored
Table 1 Polymorphism statistics of wild (W), captive (C) and reintroduced (R) populations of lesser kestrels across 8 microsatellites
Population size Code N Number of alleles per locus HeHo Rs F
IS
Fp5 Fp13 Fp31 Fp46 Fp79 Fp89 Cl347 Cl58
Southern France (W) \100 BP FRA 26 5 3 6 6 17 3 6 3 0.60–0.60 4.59 0.04
Ebro Valley (W) \1,000 BP EBV 174 6 4 7 10 33 4 10 5 0.65–0.64 4.92 0.026
Spanish core area (W)
12,000–19,000 BP
SCA 207 6 4 7 9 38 4 11 5 0.65–0.65 5.12 0.014
Portugal (W) \300 BP POR 25 6 3 6 7 19 3 8 3 0.66–0.65 5.06 0.016
GREFA (C) \100 BP 32 6 3 7 9 25 4 9 3 0.68–0.67 5.33 0.028
DEMA (C) \100 BP 59 6 4 7 7 28 4 8 4 0.67–0.68 5.04 -0.007
Gerona (R) \50 BP GER 14 5 4 6 4 16 3 5 3 0.64–0.62 4.93 0.078
Lleida (R) \100 BP LLE 25 5 3 4 7 21 4 8 4 0.63–0.61 4.95 0.060
La Rioja (R) \50 BP LRI 16 4 4 5 7 14 3 8 3 0.63–0.64 5.02 0.011
The number of alleles detected at each marker in each population is indicated in its corresponding column. The number of individuals sampled at
each population (N), expected heterozygosities (He), average observed heterozygosities (Ho) and allelic richness (Rs) estimates are showed.
Allelic richness estimates were based on a minimum number of 14 individuals. Estimated population sizes in breeding pairs (BP) when the
samples were taken are also given. See Fig. 1 for geographical locations
Conserv Genet (2010) 11:331–338 333
123
using the GenMapper software version 3.5 (Applied Bio-
systems, Foster City, CA, USA).
Genetic analyses
We excluded locus Fp107 from our analyses since previous
paternity and population genetic studies conducted for lesser
kestrels have shown the occurrence of null alleles and sig-
nificant heterozygosity deficits at this locus (see Alcaide
et al. 2005, 2008a, b, 2009). No mismatches in the segre-
gation of alleles from parents to offspring, significant
deviations from Hardy–Weinberg expectations or evidence
of linkage disequilibrium between any pair of loci have been
detected in previous studies after using the same molecular
methods. We therefore employed the permutation test
(N = 10,000) implemented in the program FSTAT ver 2.9.3
(Goudet 2001) to test for significant differences in genetic
diversity among captive, wild and reintroduced populations.
In order to avoid putative biases caused by uneven sampling,
the software FSTAT calculates a standardised estimate of
allelic richness (R
S
) independent of sample size. Average
observed heterozygosity (Ho) and the inbreeding coefficient
F
IS
were also calculated and compared using FSTAT. The
extent of population differentiation was calculated accord-
ing to the traditional F
ST
estimate using the software
GENETIX 4.04 (Belkhir et al. 1996). The significance of
pairwise F
ST
estimates was given by a P-value calculated
using 10,000 random permutations tests that were further
adjusted according to sequential Bonferroni corrections for
multiple tests (Rice 1989).
Paternity inference within colonial enclosures
We inferred paternity at the two largest colonial breeding
pens that were kept at DEMA facilities during the 2004
breeding season. Such colonial enclosures contained 36
and 16 adult kestrels, respectively, supplied with ad libitum
feeding. All individuals were identifiable through PVC
rings. Colonial enclosures consisted of several labelled
nest-boxes which could be manipulated from the exterior
of the building. Thus, eggs could be easily removed
without disturbing the whole colony. All eggs were label-
led according to where the nests they were laid to control
for the origin of the artificially reared nestlings. Nests
boxes were also provided with devices to observe the
inside of the nest. Incubating females could therefore be
identified. This fact, jointly with the registration of copu-
lation events between marked birds, allowed us to elucidate
what breeding pairs were attending each particular nest.
All adult birds and nestlings were genotyped at six out
of the nine microsatellite markers mentioned above (Fp5,
Fp31, Fp46-1, Fp79-4, Fp89 and Cl347). Locus Fp107 was
excluded because of mismatches, probably due to the
amplification of null alleles, in the segregation of alleles
from parents to offspring (see Alcaide et al. 2005 for
details). There was no special reason for excluding Loci
Fp13 and Cl58 except for their comparably low polymor-
phism and because of the aim of accelerating the data
collection process without compromising the resolution
power of the molecular approach. Parentage exclusion for
first and second parents, as well as the probability of two
individuals sharing the same genotype was calculated with
CERVUS 2.0 (Marshall et al. 1998) and IDENTITY 1.0
(Wagner and Sefc 1999), respectively. Mendelian inheri-
tance was checked at every locus in each particular case.
Those nestlings sharing alleles from their putative parents
at all loci were considered actual offspring of the couple.
The genotypes of the remaining males in the colony were
also checked to assure unequivocal paternity assignments.
Those cases in which nestlings would fail to match any of
the two alleles of the putative father at two or more loci
were considered as the result of extra-pair paternity.
Calculation of variances in breeding success of captive
kestrels during single-pairing breeding strategies
From 1996 to 2007, the number of fledglings produced
by 35 reproductive lesser kestrels kept in TORREF was
registered. The number of fledglings produced by 70
reproductive adults kept in GREFA was available from
2005 to 2007 breeding seasons. In both cases, we focused
exclusively on those kestrels that raised offspring, so these
numbers did not include non-breeding birds. We calculated
the minimum number of breeding pairs that contributed to
a high proportion of fledglings during the period of time
investigated in each particular case.
Results
Genetic diversity in captive populations
The permutation test performed in FSTAT did not report
statistically significant differences in allelic richness (5.04 vs.
5.18), average observed heterozygosities (0.64 vs. 0.68) or the
inbreeding coefficient F
IS
(0.021 vs. 0.006) between wild and
captive populations after analysing eight polymorphic
microsatellite markers (all two-tailed P-values [ 0.05, see
Table 1).
Both captive populations analysed (DEMA and
GREFA) were genetically differentiated from the Ebro
Valley and the French populations, but pair-wise F
ST
estimates did not significantly differ from 0 when com-
pared to wild populations from southwestern Iberia (SCA
and POR, Table 2).
334 Conserv Genet (2010) 11:331–338
123
Single pairing versus colonial breeding strategies
The analysis of the breeding performance data set from the
captive stocks of GREFA and TORREF revealed that, in
both cases, only a low proportion of breeding pairs (28.6%)
contributed to at least one half of the total number of
fledglings produced (50 and 56%, respectively). Paternity
inference within the colonial enclosures kept at DEMA
facilities revealed similar results, with only seven breeding
pairs (26.9% of the reproductive birds) contributing to 56%
of the fledglings produced during the 2004 breeding
season. Concerning mixed-reproductive strategies, we
detected two cases of sequential polygyny, i.e. males
raising two broods with successive females, in the largest
colonial pen in DEMA. On the contrary, no genetic evi-
dence of extra-pair paternity was found. All paternity
assignments were assigned unequivocally, especially due
to the highly polymorphic locus Fp79-4 (Table 1). The
combined probability of exclusion for the microsatellite
marker set that we used was estimated at 0.95. The like-
lihood of two individuals carrying an identical genotype
was estimated at 6.21 9 10
-6
.
Genetic diversity in reintroduced populations
We did not find statistically significant differences in allelic
richness (5.04 vs. 4.97), average observed heterozygosities
(0.64 vs. 0.62) or the inbreeding coefficient F
IS
(0.021 vs.
0.049) between wild and reintroduced populations (all two-
tailed P-values [ 0.05). However, reintroduced popula-
tions showed statistically significant lower average
heterozygosities (0.62 vs. 0.68) and higher inbreeding
coefficients F
IS
(0.049 vs. 0.006) in relation to the captive
demes from which released birds came (two-tailed P-val-
ues = 0.012 and 0.031, respectively).
Reintroduced populations only showed statistically sig-
nificant evidence of genetic differentiation when compared
to the geographically isolated population from Southern
France (Fig. 1; Table 2). Genetic divergence in relation to
the French population is comparably high in spite of
the geographic proximity of reintroduced populations.
Thus, reintroduced populations somewhat depart from the
isolation-by-distance patterns documented for natural
populations of lesser kestrels in Eurasia (see Alcaide et al.
2008a, b, 2009 for details).
Discussion
This study supports the utility of several management rec-
ommendations, such as the registration of crosses between
individuals in proper stud books and the frequent introduc-
tion of new individuals into the genetic pools, to maintain
high levels of genetic diversity in captive populations of
lesser kestrels without previous genetic monitoring. Poly-
morphisms statistics at 8 microsatellite markers in lesser
kestrels argue against low genetic variation as a primary
cause of the comparably low and occasionally very low
hatching rates documented in captivity (see Cola
´
s et al.
2002; Pomarol et al. 2004a). Rather, high rates of hatching
failure could be linked to other factors such as the feeding
conditions of the breeding stock and/or the management of
the eggs (e.g. Pomarol et al. 2004a). F
ST
-pairwise estimates
also revealed that both captive demes analysed did not differ
significantly from their natural source population, a fact that
reinforces the absence of strong fluctuations in the distri-
bution of allele frequencies.
Genetic diversity in reintroduced populations did not
differ significantly from natural populations in the absence
of previous genetic monitoring either. From the perspective
of population structuring, the departure of reintroduced
populations from naturally occurring isolation-by-distance
patterns (see Alcaide et al. 2008a, b) can be attributed to
the lack of migration-drift equilibrium in recently founded
populations (see for instance Leberg and Ellsworth 1999;
DeYoung et al. 2003). However, our results suggest that
Table 2 F
ST
-pairwise values (above diagonal) between four geographically distinct natural populations of lesser kestrels (W), captive (C) and
reintroduced populations (R)
EBV (W) SCA (W) POR (W) FRA (W) GER (R) LLE (R) LRI (R) GREFA (C) DEMA (C)
EBV (W) 0.003* 0.005 0.012* 0.008 0.006 0.013 0.010* 0.008*
SCA (W) 0.004 0.016* 0.010 0.006 0.009 0.006 0.006
POR (W) 0.027* 0.007 0 0.010 0.003 0.008
FRA (W) 0.019* 0.028* 0.032* 0.033* 0.025*
GER (R) 0.001 0.030* 0.017 0.013
LLE (R) 0.012 0.010 0.010
LRI (R) 0.013 0.014
GREFA (C) 0.005
Significant values after Bonferroni corrections for multiple tests are indicated by asterisks. See Fig. 1 for geographic locations
Conserv Genet (2010) 11:331–338 335
123
uneven contributions of reproductive birds to the captive-
born progenies may be responsible for a non-optimal
transmission of genetic diversity from captive stocks to
reintroduced populations. This fact, which has been already
documented in the literature for other captive flocks (e.g.
McLean et al. 2008), is particularly important in lesser
kestrels since many of the most prolific breeding pairs are
forced to produce a second and even a third clutch during
the same breeding season (Pomarol et al. 2004a;J.L.
Antolı
´
n et al., personal communication). As this study
shows, large variations in reproductive success of indi-
viduals are similarly occurring for both single-pairing and
colonial breeding facilities, with only about one-fourth of
the reproductive birds producing 50–56% of fledglings.
Hence, the occurrence of polygynous behaviours at low
rates does not seem to significantly decrease the effective
population size. The lack of extra-pair fertilizations, on the
other hand, suggests that an increase in mate guarding
might have overridden the effects of large breeding den-
sities or female promiscuity in colonial breeding systems
with ad libitum feeding. Our results do not seem to support
smaller variances in individual breeding success linked to
social stimulation of breeding and a broader availability of
potential mates either.
Founder effects during both captive breeding and set-
tlement stages can be counteracted by minimizing the
release of related birds into the same location. A recent
study by Lenz et al. (2007) also suggests the utility of
manipulating sex-ratios to increase the effective population
size during captive breeding of this species. Immigration is
particularly important to diminish average genetic simi-
larity and increase overall heterozygosity, as it has been
already demonstrated by Ortego et al. (2007) in natural
population of lesser kestrels. The effect of conspecific
attraction in this respect is particularly well documented
(Serrano and Tella 2003; Serrano et al. 2004, but see
Calabuig et al. 2008), and thus, birds kept in pens or even
decoys can be regularly used in newly established colonies
to promote both settlement of released individuals and
recruitment of wild birds. As Pomarol et al. (2004b) have
previously indicated, immigration from the close Ebro
Valley population may have decisively contributed to
population growth in the reintroduced populations in
Catalonia (GER and LLE). Such gene flow events may also
explain the lack of significant patterns of genetic differ-
entiation between natural and reintroduced populations
(Table 2). Although immigration may involve individuals
dispersing long distances, as exemplified by one bird from
the Ebro Valley (North Eastern Spain) recruited as a
breeder 300 km away in the reintroduced population of
Villena (Middle Eastern Spain, M Alberdi, personal com-
munication), dispersal probabilities between populations
sharply decrease with geographic distance in this species
(Serrano and Tella 2003; Alcaide et al. 2008a, 2009).
Given that reintroduction programs may be necessary in
highly isolated areas where natural colonization and
immigration are highly improbable, reintroducing geneti-
cally diverse birds may be of importance to guarantee
population persistence.
In conclusion, this study revealed high levels of genetic
variation for ongoing but non-genetically monitored captive
breeding and reintroduction programs of the lesser kestrel.
However, we found a significant loss of genetic variation
from captive flocks to reintroduced populations because of
large variances in breeding performance of individuals.
Although the lesser kestrel program does not seem to be
seriously compromised by this finding, this information
could be crucial for highly endangered species in which the
number of founders remains below the recommended min-
imum (20–30 individuals), and where the incorporation rates
of new birds to refresh the genetic pools and natural gene
flow is comparably low. Undoubtedly, genetic monitoring is
a desirable practice to maximize reproductive success and
genetic variation in captive-born individuals which will be
subsequently released into the wild or used to supplement the
captive stocks (Frankham et al. 2002; see examples in
Gautschi et al. 2003; Ralls and Ballou 2004; Hedrick and
Fredrickson 2008). Genetic monitoring can however become
costly and time-consuming, especially if molecular markers
for the target species are not available. Since some conser-
vation initiatives cannot simply afford it, the experiences
summed from other captive breeding and reintroduction
programs can become of high assistance.
Acknowledgments We are indebted to all the people who kindly
helped to collect kestrel samples Therefore, we are thankful to J. L.
Tella, E. Ursu
´
a, A. Gajo
´
n, J. Blas, G. Lo
´
pez, C. Rodrı
´
guez, J. Bus-
tamante, R. Alca
´
zar, J. D Morenilla, P. Prieto, I. Sa
´
nchez, A. Garcı
´
a,
I. Ga
´
mez, F. Carbonell, G. Gonza
´
lez, R. Bonal, J. M. Aparicio, A. de
Frutos, P. Olea, E. Banda, C. Gutie
´
rrez, P. Pilard and L. Brun. We
especially thank people from the captive breeding centers of DEMA,
GREFA and TORREFERRUSA (M. Martı
´
n, F. Carbonell and others).
Daniel Janes and Tobias Lenz definitely contributed to improve this
manuscript. We are also indebt to the Associate Editor Dr. Vicki
Friesen and several anonymous reviewers for their kind and helpful
assistance during the peer-review process.This study was supported
by the MCyT (project REN2001-2310 and CGL2004-04120), which
also provided a research grant to M. Alcaide.
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... Among the chicks born in 2017, there were one male and four females. Since the microsatellite loci used in our work have autosomal localization [20,21,[29][30][31], the sex ratio in the samples does not affect estimates of genetic variation. ...
... Microsatellite loci have been used for over 20 years to study genetic variation and analyze breeding and relatedness in populations of different falcon species [20,21,[29][30][31][32][33][34][35][36][37][38][39][40][41][42][43]. Nevertheless, the genetic variation of natural populations of falcons that inhabit the territory of the Russian Federation has still not been investigated well. ...
... A study of the nest structure for Finnish peregrine falcon using ten microsatellite loci, nine of which were the same as in our study, showed even lower estimates of PI = 6.68 × 10 -8 and PIsibs = 1.37 × 10 -3 [40]. In the study of the population structure of the lesser kestrel, Falco naumanii, using eight heterologous loci of peregrine falcon and gyrfalcon, it was found that PI = 6.21 × 10 -6 [29]. The estimates using five highly polymorphic loci for the merlin, Falco columbarius, were PI = 5.69 × 10 -7 and PIsibs = 1.67 × 10 -4 [49]. ...
... Genetic assessments that do exist usually compare reintroduced and captive populations such as those in the lesser kestrel (Falco naumanni) (Alcaide et al. 2010); the blackfooted ferret (Mustela nigripes) (Wisely et al. 2008); American martens (Martes americana) (Hillman et al. 2017) and Griffon vulture (Gyps fulvus) (Le Gouar et al. 2008). Such studies have also been used to infer historic genetic patterns, including bottleneck events from modern populations (e.g., Alpine ibex (Capra ibex ibex) Biebach and Keller 2009) and to assess gene flow between reintroduced and natural populations (e.g., in eastern wild turkeys (Meleagris gallopavo silvestris), Latch and Rhodes 2005). ...
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... population and the chances of successful population establishment are often lower if captive animals are used compared to directly moving wild sourced individuals(Fischer and Lindenmayer 2000;Williams and Hoffman 2009). Unless all individuals are taken from the wild(Dobson and Lyles 2000) the source population for the captive stock will be a subset of the remaining wild population, potentially resulting in an additional bottleneck and negating some captive breeding objectives(Alcaide et al., 2010).Populations can adapt to captivity when the selective pressures faced in the wild are removed, relaxed and/or replaced due to disease management, predator removal, food and husbandry protocols, resulting in traits that hinder establishment and persistence and recruitment in the wild(Britt et al., 2004;Håkansson and Jensen 2005;Williams and Hoffman 2009). Examples of adaptation to captivity have been demonstrated in the golden lion tamarin (Leontopithecus rosalia)(Britt et al., 2004) and the jungle fowl (Gallus gallus)(Håkansson & Jensen 2005). ...
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Reintroductions are becoming an increasingly popular tool for threatened species management and broader scale restoration projects. Reintroductions require a series of important decisions to be made from planning and implementation through to postrelease establishment and persistence of populations. Decision making in reintroduction is frequently impeded by high levels of uncertainty. Linguistic, epistemic and aleatory uncertainties often lead to a failure to meet project objectives. This has led to repeated calls for setting clear objectives and using these to focus monitoring in a way that allows applied science to support management. // Viewed in this way, applied science can naturally assist the decision making process. It is important to reduce only the uncertainties that will help inform the choice between two or more possible actions. These can be reduced through targeted monitoring and research. The failure of applied science to approach research in this way is one possible explanation for the ‘research –implementation gap’ that persists in conservation biology. Throughout this thesis I use decision analytic tools to evaluate and inform the discipline of reintroduction biology. Decision analytic tools are increasingly being utilised in diverse fields of resource management. The benefits for more formally incorporating decision science into conservation biology are obvious and repeatedly lauded, yet it remains unclear how much the approach is used to ensure applied science is truly informing management, particularly in the growing discipline of reintroduction biology. // Overall, my PhD intends to promote the application of formal decision tools to threatened species management and showcase how it can reduce uncertainty and support decision making specifically in reintroductions. In using the Regent Honeyeater recovery actions as a case study, I will evaluate whether management actions to recover the species are working, as well as highlighting areas where resources can be targeted to reduce the uncertainties that influence management decisions, rather than wasting it on those that are not relevant.
... Indeed, conservation efforts have saved several raptor species from the brink of extinction (e.g., Butchart et al., 2006;Cade and Burnham, 2003;Jones et al., 1995;Snyder and Snyder, 2000). Translocation programs are restoring some populations of threatened raptors (e.g., Alcaide et al., 2010;Ferrer et al., 2014;McClure et al., 2017). Technology is being developed to mitigate mortality at wind power facilities (e.g., Foss et al., 2017;Marques et al., 2014;McClure et al., 2018). ...
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... The lesser kestrel (Falco naumanni) is a small (120-145 g), long-lived (maximum live-span 10 years) colonial falcon, being females large than males [23] and references therein. The species feeds mainly on invertebrates, and has experienced a marked decline in some areas of its breeding range during the last 30 years, being the target species for several reintroduction programs [24]. The lesser kestrel data on demographic parameters was taken from literature [23] coming from a color-ringing and monitoring of breeding performances in 12 colonies in the Seville province (Spain) during 6-year period (1988-1993). ...
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... Unfortunately, post-release genetic monitoring is rarely done (Schwartz et al., 2007). We found few examples of studies examining the genetics of released cohorts (Drauch and Rhodes, 2007;Gonzalez et al., 2008;Karlsson et al., 2008;Alcaide et al., 2010; this study) and these examined only a single year class or provided a single "snapshot" of the genetic composition of multiple year classes rather than examining trends in genetic composition over time. Likely constraints on post-release monitoring include time/labor investment, financial limitations, low recapture probability, and desire to avoid imposing handling stress on valuable captive-released individuals. ...
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... Due to the lesser kestrel decline and also for research purposes, several breeding programs have been put in place in Spain in recent years (Pomarol 1993;Negro et al. 2007;Alcaide et al. 2010). One of these reintroductions was carried out in (2008)(2009)(2010)(2011) at an external cage (6x2x2 m) to facilitate conspecific attraction at the colony. ...
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Unmanned Aerial Systems (UAS) have been used for decades in the military field, mainly in dangerous or tedious missions where it is preferable to send a vehicle equipped with sensors than to use human piloted conventional aircrafts for information gathering. In recent years technology has advanced, the market has grown exponentially, prices have descended and the use of the systems is simpler, which has led to the incorporation of the UAS to the civilian world. UAS have proven useful in ecology related tasks, such as animals monitoring and habitats characterization, and their potential for spatial ecology has been pointed out, but to date there are just a few studies addressing their specific use in conservation biology. This Ph.D. thesis attempts to fill the gap of knowledge in practical functions of small UAS in conservation biology. It describes for the first time the use of these systems in an immediately applicable way for impact assessment of infrastructures and protection of endangered species. It also presents UAS as a tool for obtaining high- resolution spatiotemporal information, which helps to understand habitat use in rapidly changing landscapes. Furthermore, it demonstrates that these systems can provide information as valid as the obtained by conventional techniques on the spatial distribution of species in protected areas. The experiments performed in the frame of this thesis show that low cost small UAS equipped with embarked cameras that provide high-resolution images offer the possibility of monitoring the environment at the researcher’s desired frequency and revisiting sites to perform systematic studies, which is valuable for ecological research. The results also revealed that UAS use in conservation biology has some constraints, mainly related with the scope of the missions, the limiting costs of the systems, operating restrictions associated to weather, legal limitations and the need of specialized personnel for operating the systems, as well as some difficulties for data analysis related with image processing. Overall, given the novelty of the subject and the importance it is expected to have in the near future, I consider that providing information on the capabilities and limitations of UAS, based on practical experiments in conservation biology, is not only of scientific interest but combines environmental and industry interests, which brings added value and usefulness of this thesis to society.
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Oryx (2006), 40:4:411-418 Cambridge University Press Copyright © 2006 Fauna & Flora International doi:10.1017/S0030605306001360 Species reintroduction programmes, in prioritizing areas for reintroductions, have traditionally used tools that include measures of habitat suitability and evaluations of area requirements for viable populations. Here we add two tools to this approach: evaluation of ecological requirements of species and evaluation of future suitability for species facing changing climates. We demonstrate this approach with two species for which reintroduction programmes are in the planning stages in Mexico: California condor Gymnogyps californianus and Mexican wolf Canis lupus baileyi. For the condor, we identify three areas clustered in the Sierra San Pedro Ma´rtir, Baja California; for the wolf, we identify a string of suitable sites along the Sierra Madre Occidental of western Mexico. We discuss the limitations of this approach, identifying ways in which the models illustrated could be made more realistic and directly useful to reintroduction programmes.
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The ratio of the effective population size to adult (or census) population size (N-e/N) is an indicator of the extent of genetic variation expected in a population. It has been suggested that this ratio may be quite low for highly fecund species in which there is a sweepstakes-like chance of reproductive success, known as the Hedgecock effect. Here I show theoretically how the ratio may be quite small when there are only a few successful breeders (N-b) and that in this case, the N-e/N ratio is approximately N-b/N. In other words, high variance in reproductive success within a generation can result in a very low effective population size in an organism with large numbers of adults and consequently a very low N-e/N ratio. This finding appears robust when there is a large proportion of families with exactly two progeny or when there is random variation in progeny numbers among these families.
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Inbreeding depression is thought to be a major factor affecting the evolution of mating systems and dispersal. While there is ample evidence for inbreeding depression in captivity, it has rarely been documented in natural populations. In this study, I examine data from a long-term demographic study of an insular population of song sparrows (Melospiza melodia) and present evidence for inbreeding depression. Forty-four percent of all matings on Mandarte Island, British Columbia, were among known relatives. Offspring of a full-sib mating (f = 0.25) experienced a reduction in annual survival rate of 17.5% on average. Over their lifetime, females with f = 0.25 produced 48% fewer young that reached independence from parental care. In contrast, male lifetime reproductive success was not affected by inbreeding. Reduced female lifetime reproductive success was mostly due to reduced hatching rates of the eggs of inbred females. Relatedness among the parents did not affect their reproductive success. Using data on survival from egg stage to breeding age, I estimated the average song sparrow egg on Mandarte Island to carry a minimum of 5.38 lethal equivalents (the number of deleterious genes whose cumulative effect is equivalent to one lethal); 2.88 of these lethal equivalents were expressed from egg stage to independence of parental care. This estimate is higher than most estimates reported for laboratory populations and lower than those reported for zoo populations. Hence, the costs of inbreeding in this population were substantial and slightly above those expected from laboratory studies. Variability in estimates of lethal equivalents among years showed that costs of inbreeding were not constant across years.
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Restocking programs were significant components of initiatives that restored white-tailed deer (Odocoileus virginianus) populations to the southeastern United States. However, past studies have reached conflicting conclusions regarding the effectiveness and consequences of translocations on the genetic structure of southeastern deer. We conducted further analysis of published datasets via matrix comparison methods to resolve these differences. Our analysis suggests translocations have had substantial and persistent effects on the genetic composition of deer populations into which translocated individuals were released. Regional and long-distance translocations influenced local populations by reducing the relationship between genetic differentiation and geographic distance among populations,and concordance between patterns of mitochondrial DNA (mtDNA) and allozyme variation. Strong associations of geographic and genetic distance among populations not directly receiving restocked deer indicate the genetic contributions of translocations are localized due to limited dispersal. Coastal island populations may warrant additional protective measures because they retain much of the historic genetic structure of southeastern white-tailed deer and may represent reservoirs of unique genetic material.
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The last wild California Condor (Gymnogyps californianus) was brought into captivity in 1987. Captive breeding was successful and reintroduction efforts began in 1992. The current population is descended from 14 individuals belonging to three genetic "clans." This population bottleneck led to the loss of genetic variation and changes in allele frequencies, including a probable increase in the frequency of the putative allele for chondrodystrophy, a lethal form of dwarfism. We use studbook data to analyze the current genetic and demographic status of the population and explain how it is managed to meet specific goals. In August 2002 the population consisted of 206 individuals distributed among three captive-breeding facilities and three reintroduction sites. The population is managed to preserve genetic diversity using the concept of mean kinship. Growth of the total population has been between 10% and 15% per year since 1987, but the growth of the captive population has been only about 5% per year since 1992 due to the removal of chicks for reintroduction. Assuming that founding birds within clans were half-siblings, the birds used to found the captive population theoretically contained 92% of the heterozygosity present in the hypothetical wild base population. About 99.5% of this heterozygosity has been retained in the current population. Alleles from most founders are well represented across captive-breeding facilities and reintroduction sites. The genetic status of this population compares favorably with other species that have been rescued from extinction by captive breeding.