Genetic consequences of animal translocations: A case study using the field cricket, Gryllus campestris L.

B I O L O G I C A L C O N S E RVAT I O N

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

available at www.sciencedirect.com

journal homepage: www.elsevier.com/locate/biocon

Genetic consequences of animal translocations: A case study using the field cricket, Gryllus campestris L. K.A. Witzenbergera, A. Hochkircha,b,* a

University of Osnabru¨ck, Department of Biology/Chemistry, Division of Ecology, Barbarastraße 13, D-49076 Osnabru¨ck, Germany University of Trier, Department VI, Biogeography Group, Am Wissenschaftspark 25-27, D-54296 Trier, Germany

b

A R T I C L E I N F O

A B S T R A C T

Article history:

Animal relocations have become a common tool in nature conservation, but the genetic

Received 23 April 2008

consequences of such projects have rarely been studied in insects. As both natural and

Received in revised form

artificial formation of new populations may lead to genetic drift (founder effect),

2 September 2008

decreased genetic diversity and increased rates of inbreeding, genetic analyses can pro-

Accepted 14 September 2008

vide valuable information to evaluate the success of a relocation project. The field

Available online 26 October 2008

cricket (Gryllus campestris) has been subjected to reintroduction and translocation projects in England and northern Germany. Here, we present a microsatellite study on

Keywords:

the population genetics of one recently established population of this species in compar-

Conservation genetics

ison with several older populations and some recently colonized sites. Our results show

Founder effect

that the translocation did not result in a significant loss of genetic diversity, when com-

Insect conservation

pared to source and other natural populations suggesting that translocation of a high

Microsatellite

number of nymphs from different subpopulations may be a suitable method to decrease

Population genetics

the loss of genetic diversity and reduce the risk of inbreeding. Furthermore, the trans-

Reintroduction

location had no negative effect on the source population, which reached a new maximum population size in 2006. An assignment test showed that individuals from the translocated population (F4 generation) were still assigned to the source populations, whereas two young subpopulations that originated by natural colonization from the central population about ten years ago already formed separate genetic clusters. As the strong fragmentation of G. campestris populations in northern Germany hampers natural colonization of newly created potential habitats, translocation projects seem to be an appropriate method to preserve this species. Ó 2008 Elsevier Ltd. All rights reserved.

1.

Introduction

Habitat loss, degradation and fragmentation are believed to be the key threats to biodiversity (Primack, 2002; Gro¨ning et al., 2007). One consequence of habitat fragmentation is a reduced genetic exchange among populations inhabiting different habitat patches, leading to a genetic isolation of popu-

lations (Saunders et al., 1991; Fischer and Lindenmayer, 2007). Subsequently, the genetic diversity within populations might decrease, whereas population differentiation might increase over time (Frankham et al., 2005). Reduced gene flow and the loss of genetic diversity can also enhance the rate of inbreeding, which is considered to reduce a population’s adaptability to a changing environment. Moreover,

* Corresponding author: Address: University of Trier, Department VI, Biogeography Group, Am Wissenschaftspark 25-27, D-54296 Trier, Germany. Tel.: +49 651 201 4692; fax: +49 651 201 3851. E-mail addresses: [email protected] (K.A. Witzenberger), [email protected] (A. Hochkirch). 0006-3207/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.biocon.2008.09.017

3060

B I O L O G I C A L C O N S E RVAT I O N

homozygous recombination of deleterious recessive alleles become more likely, which may lead to inbreeding depression and increase the probability of extinction (Reed et al., 2002). Habitat management, such as the restoration of degraded habitats and the creation of corridors between remaining habitats are known to be suitable measures to counteract such processes (Hochkirch et al., 2008), but in many cases, the conditions for habitat restoration are poor or natural populations are too fragmented and not capable of (re)colonizing existing suitable habitat (Cheyne, 2006). During recent decades, reintroduction and translocation projects have become widespread tools to accelerate colonization processes of endangered species. However, such projects are costly and require continuous supervision and monitoring (IUCN, 1998; Fischer and Lindenmayer, 2000). An evaluation of the translocation success is often impossible without adequate monitoring programs. Monitoring of translocation projects should at least include an evaluation of population persistence, growth and dispersal over some generations, as well as an analysis of habitat improvement (IUCN, 1998; Cheyne, 2006). As translocated populations are often based on a low number of founders, inbreeding and loss of genetic diversity may hamper the success of such projects (reviewed in Mock et al., 2004). Population genetics can therefore provide valuable information on the newly established population, such as the genetic diversity, inbreeding effects or information on which founders contributed most to the newly established population. Despite the increasing number of translocation projects, studies on the genetic consequences of translocations are comparatively sparse. Most studies on translocation genetics deal with highly endangered vertebrates (e.g. Houlden et al., 1996; Wisely et al., 2003), fishes (e.g. Fumagalli et al., 2002; Eldridge and Naish, 2007) or game (e.g. Mock et al., 2004), whereas insects have rarely been studied. In most cases, the influence of translocation projects on the genetic diversity of populations remains unknown. The field cricket (Gryllus campestris L., 1758) is endangered in large parts of Central and Northern Europe (Hochkirch et al., 2007). It is confined to dry, oligotrophic habitats, such as heathland and dry grassland, which have become rare and fragmented at the northern edge of its range (Kleukers et al., 1997; Lu¨tkepohl, 2001; Fottner et al., 2004). The species represents one of the few insect species, which has been subjected to reintroduction and translocation programs. It has been successfully reintroduced in England, based on a captive breeding program (Pearce-Kelly et al., 1998). In northern Germany, a new population has recently been established following a translocation project (Hochkirch et al., 2007). Here, we analyse the genetic consequences of this project by comparing the genetic diversity and levels of inbreeding of the translocated population with some of the remaining populations in northern Germany and some sites that have been recently colonized by natural dispersal. Our aim is to examine, whether the genetic variability in the translocated population was comparable to the diversity in natural populations and whether inbreeding might represent a problem in the new populations. For better comparison, we examined the genetic diversity in the few remaining wild populations in northern Germany.

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

2.

Methods

2.1.

Study object

The field cricket (G. campestris) is a flightless and comparatively large cricket species (17–26 mm), which is characterized by a shiny black body coloration (Marshall and Haes, 1990). The species mainly inhabits dry grasslands and is restricted to heathlands and oligotrophic grasslands at the northern edge of its range (Kleukers et al., 1997). The reproductive season of the univoltine species lasts from May to July. Nymphs hatch in mid July and hibernate during their tenth or eleventh instar (Ko¨hler and Reinhardt, 1992). The final moult takes place at the end of April or at the beginning of May. While males are territorial and defend their burrows fiercely, females are vagrant and are attracted by singing males. They lay their eggs in bare ground either close to a burrow or inside the burrow. Populations of G. campestris are known to undergo extreme fluctuations and are strongly affected by weather conditions (Remmert, 1992). G. campestris is declining and red-listed in large parts of Central and Northern Europe, such as the UK (Pearce-Kelly et al., 1998), Germany (Ingrisch and Ko¨hler, 1997), the Netherlands (Kleukers et al., 1997), Belgium (Decleer et al., 2000), Luxembourg (Proess and Meyer, 2003), Denmark (http://redlist.dmu.dk, 2006) and Lithuania (Budrys and Pakalnisˇkis, 2007).

2.2.

Study area and translocation project

The study area is located in the central part of the natural region ‘‘Diepholzer Moorniederung’’(DHM) between the towns Hanover, Bremen and Osnabru¨ck (Lower Saxony, northern Germany; Fig. 1). The area is characterized by large peat bogs, wetlands and dry sand ridges (Hochkirch and Adorf, 2007). In Lower Saxony, the field cricket is listed as critically endangered (Grein, 2005), with about ten populations left (Grein, 2000; Hochkirch et al., 2007). The source population is located at the eastern edge of the nature reserve, ‘‘Neusta¨dter Moor’’ and was the only natural population remaining west of the river Weser (Hochkirch, 1996; Teerling and Hochkirch, 2002). This population has been monitored and managed intensely since 1990 by the non-governmental organization BUND (‘‘Friends of the Earth – Germany’’). In 1991, the population size decreased to 32 singing males, most of which were located on one heathland site (GH, Fig. 2). Due to careful habitat management, the population increased during the following years and spread until it reached the peat bog in the west and north (Fig. 2). In 2001, the population reached 949 singing males and a translocation project was started, intending to establish a second self-sustaining population in a nearby nature reserve (‘‘Renzeler Moor’’). It was decided to use only individuals from the Neusta¨dter Moor population, which had been monitored for a long period and appeared to be very healthy. Other populations were not considered, as there was no exact data available on their population sizes, population trends and their genetic identities. The Neusta¨dter Moor population consisted of several subpopulations, some of which originated recently by natural colonization (Fig. 2). In order to avoid harm to any of these subpopulations (particularly to the oldest subpopulation GH) and to increase the

B I O L O G I C A L C O N S E RVAT I O N

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

3061

Fig. 1 – Location of the study sites (black circles) in Lower Saxony. The detail shows the sampling sites in the main study area ‘‘Diepholzer Moorniederung’’ (DHM). Sampled populations are; Posthausen (PO), Hanover (HA), Senne (SE), Renzeler Moor (RM), Hahnenberg (HB), Grillenheide (GH) and Moorrand (MR).

genetic diversity as far as possible, nymphs were collected from different sites spanning the whole population (Hochkirch et al., 2007). In July 2001, a total of 213 nymphs were collected, including 75 specimens from GH and 99 specimens from HB as well as another subpopulation between these two sites. The nymphs were then released at the ‘‘Renzeler Moor’’ (RM). This artificial population increased in size and reached a maximum of 335 males in 2005 (see Hochkirch et al., 2007 for more detailed information).

2.3.

Sampling

Sampling for genetic analyses was conducted in 2006 (F4 generation), at the end of the reproductive period (June and July). We used a non-lethal sampling method to avoid negative effects on the (sub-) populations (we use this term when referring to both populations and subpopulations). As Orthoptera readily autotomize a hind leg when caught, we sampled single hind legs of about 20 specimens per (sub-) population. In addition to the translocated population (Renzeler Moor, RM), we sampled three subpopulations of the source population (Neusta¨dter Moor). These included the old core subpopulation (Grillenheide, GH) and two subpopulations, which originated from natural colonization during the last decade (HB and MR, see Figs. 1 and 2). For comparison we also included two other populations from Lower Saxony (Posthausen, PO and Hanover, HA) as well as a large population from North Rhine-Westphalia (Senne, SE). Further populations occur near Mu¨nster and in the east of Lower Saxony. However, these proved to be too small to include them in the analysis. As field crickets produce a loud and distinct calling song, it is unlikely that any population in Lower Saxony remain undetected.

2.4.

DNA extraction and amplification

Genomic DNA was extracted from muscle tissue using the DNeasy Tissue Kit (Qiagen), following the manufacturers protocol. Each sample was typed at four microsatellite loci, Gbim04 and Gbim15 developed by Dawson et al. (2003) as well as Gbim29 and Gbim66 developed by Bretman et al. (2008). Initially three more primers (Gbim03, Gbim06 and Gbim26 from the same authors) were used, but these had to be excluded, as the number of null alleles was too high. Amplification was performed using the HotMasterMix (Eppendorf). The 5 0 -end of each forward primer set was labelled with a fluorescent marker, either 5-FAM or JOE. The products were genotyped on an ABI PRISM 377 automated DNA sequencer (Applied Biosystems Inc.). Fragment lengths were determined using GENESCAN and GENOTYPER 2.5 (Applied Biosystems Inc.).

2.5.

Statistical analysis

The mean number of alleles, allelic richness and inbreeding coefficients (FIS) for each population or subpopulation was estimated using FSTAT 2.9.3.2 (updated from Goudet, 1995). The measure ‘‘number of effective alleles’’ (Table 1) differs from the number of observed alleles in so far, as alleles with very low frequencies are not considered, since they are very likely to be lost in a bottleneck. The measure of allelic richness (El Mousadik and Petit, 1996) was used, as it is independent of sample size. GenAlEx 6.0 (Peakall and Smouse, 2006) was used to determine the expected (HE) and observed (HO) heterozygosity for each locus and each population (or subpopulation). GenAlEx 6.0 also performs a Chi-square test to assess the significance of a departure from Hardy–Weinberg

3062

B I O L O G I C A L C O N S E RVAT I O N

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

Fig. 2 – Number and dispersion of singing males in the Neusta¨dter Moor in four selected years: (1991) Minimum population size (map includes the location of the main study sites in the Neusta¨dter Moor: the core subpopulation ‘‘GH’’ and the subpopulations that originated from natural colonization ‘‘HB’’ and ‘‘MR’’. (1995) Colonization of HB. (1996) Colonization of MR. (2001) Year of the translocation of cricket nymphs to the Renzeler Moor. The line indicates the border of the bog. The size of the black dots corresponds to the number of singing males recorded on the site. For explanation of the abbreviations, see Fig. 1.

Table 1 – Number of alleles averaged across all loci (n = sample size), allelic richness (based on the lowest sample size: 19) and mean effective population size (NE) for the populations and subpopulations studied Population

n

Average number of alleles Observed

GH MR HB RM PO SE HA

24 23 21 21 20 21 19

4.25 3.75 3.75 4.25 5.75 4.75 4.50

Average allelic richness

Mean NE (range 2–100)

Effective 2.85 2.07 2.22 2.71 4.22 3.08 2.93

4.10 3.58 3.60 4.17 5.67 4.65 4.46

25.08 25.76 22.85 31.86 27.79 19.68 25.45

(95% (95% (95% (95% (95% (95% (95%

CL = 18.76–38.31) CL = 16.78–48.14) CL = 15.69–38.91) CL = 22.46–49.37) CL = 21.41–40.85) CL = 14.36–32.13) CL = 18.83–42.52)

The calculations of NE are based on an estimated minimum population size ranging between 2 and 100 individuals.

equilibrium (HWE). The data was inspected for the presence of null alleles using Micro-Checker 2.2.3 (Van Oosterhout et al., 2004). A test for linkage disequilibrium was performed with FSTAT 2.9.3.2 using a log-likelihood ratio G-statistic with Bonferroni corrections. The effective population size (NE) was calculated for each sampled population and for the combined

populations in the Neusta¨dter Moor using ONeSAMP (Tallmon et al., 2008). To examine the genetic structure within and between populations or subpopulations, an analysis of molecular variance (AMOVA) with 999 permutations was performed in GenAlEx 6.0 based upon RST (Slatkin, 1995) and FST. Similar to FST val-

B I O L O G I C A L C O N S E RVAT I O N

ues in Wright’s F-statistics, RST values are based upon the variance in allele frequencies among the populations or subpopulations, but RST is based on a stepwise mutation model (SMM) and more likely to fit the mutations of microsatellite loci than FST (Balloux and Lugon-Moulin, 2002). In addition, STRUCTURE 2.1 (Pritchard et al., 2000) was used to conduct an individual-based cluster analysis. The following run parameters were used: correlated allele frequency model, a burn-in period of 100,000 simulations followed by a run length of one million Markov chain Monte Carlo simulations and five iterations for each K. There should be no exchange between the populations HA, SE, PO and GH due to the great distance between them (>50 km). However, the subpopulations MR and HB originated by natural colonization from GH and there might still be dispersal between these sites. The translocated population RM was mainly founded with individuals caught at GH and HB. Therefore, we analysed the data using the admixture model. The assignment to a population was only considered certain, if the respecting probability was q P 0.8.

3.

Results

3.1.

Genetic diversity

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

3.2.

Genetic differentiation

Nearly all (sub-) populations examined showed significant differentiation from other populations (Table 3), except for the Posthausen population, which did not differ significantly from the (sub-) populations SE, HB and RM. All (sub-) populations in the Diepholzer Moorniederung (DHM) differed significantly from one another except for GH and MR. The translocated population RM was significantly differentiated from all examined subpopulations in the Neusta¨dter Moor. The analysis of the optimal number of populations (K) with STRUCTURE (Pritchard et al., 2000) showed the best value for six populations (Table 4). STRUCTURE found clear differences between the populations and between the three subpopulations at the Neusta¨dter Moor (GH, MR and HB). However, it was not possible to distinguish the translocated population (RM) from these three subpopulations. Four of the 21 individuals sampled at RM were assigned to HB, three to GH and one to MR (q P 0.8), whereas the majority of specimens (12) showed a mixed origin of different subpopulations in the Neusta¨dter Moor. Only one cricket showed a mixed assignment between GH (q = 0.58) and one of the other populations (SE; q = 0.32).

4. The population in Posthausen (PO) had the highest number of effective alleles, followed by the largest population (SE). The number of effective alleles in the population near Hanover (HA) was somewhat higher than in the Diepholzer Moorniederung (DHM). In this region, we found slightly more effective alleles in the old core population ‘‘Grillenheide’’ (GH) than in the subpopulations which originated from natural colonization during the last decade (MR and HB), but in the translocated population (RM) the value was rather similar to GH. Private alleles were found at all loci and in all populations except for the recently established subpopulations MR, HB and the translocated population RM. ONeSAMP revealed that the effective population sizes of all sampled populations were quite low, ranging from 19.7 to 31.9 (Table 1). The largest population (SE) had the lowest effective population size, whereas the highest NE was found at the translocated population (RM). The mean NE for the combined subpopulations at the Neusta¨dter Moor (i.e. GH, MR and HB) was 44.87 (95% CL = 31.87– 66.98). Four (sub-) populations showed significant deviation (heterozygote deficiency) from Hardy–Weinberg-Equilibrium at one or two loci (Table 2). This was true for SE as well as the recently established subpopulations in the DHM (MR, HB and RM). Inbreeding coefficients (FIS) ranged from 0.101 (HA) to 0.168 (PO). The old core population at the Neusta¨dter Moor (GH) had a rather low inbreeding coefficient, whereas the two naturally colonized subpopulations (HB and MR) showed much higher inbreeding coefficients (Table 2). In the translocated population (RM), FIS was intermediate. FIS values differed substantially among the four microsatellite loci (Table 2). Potential null alleles were only detected at one locus (Gbim15) in the population PO. However, the locus amplified in all individuals. The global test for genotypic linkage disequilibrium across all populations revealed no significant departure for any combination of microsatellite loci.

3063

Discussion

4.1. Translocation effects and natural colonization processes Our results show that the translocation did not result in a significant loss of genetic diversity, when compared to source and other natural populations although the source population underwent a severe bottleneck and some subpopulations have been established only recently. This suggests that the translocation of a high number of individuals from different subpopulations may be a suitable method to decrease the loss of genetic diversity and reduce the risk of inbreeding. In the source population, the young subpopulations MR and HB had comparatively high inbreeding coefficients and deviation from HWE, whereas the old core subpopulation (GH) was characterized by a low level of inbreeding and a number of private alleles. In the artificial population (RM), FIS was intermediate and allelic richness was higher than in any of the source subpopulations. The genetic pattern in the Neusta¨dter Moor (GH, MR, HB) matches the history of this population very well, which is known from a continuous monitoring programme. The population has suffered a strong decline at the beginning of the 1990s and survived only at GH from where it has expanded substantially over the course of the last two decades (Fig. 2; Hochkirch, 1996; Teerling and Hochkirch, 2002; Hochkirch et al., 2007). The decline at the beginning of the 1990s resulted in a reduction of the effective population size to approximately 45 individuals in 2006. This value fits well with the minimum population size estimated in 1991 (32 singing males, Hochkirch, 1996). Even though both young subpopulations in the Neusta¨dter Moor originated approximately during the same time by natural colonization (HB: 1995, MR: 1996), subpopulation HB was significantly differentiated from GH whereas MR was not. This difference might be explained by the different temporal courses of

3064

B I O L O G I C A L C O N S E RVAT I O N

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

Table 2 – Observed (Ho) and expected heterozygosity (He) at all loci Population

Locus

n

Ho

He

P

GH

Gbim04 Gbim15 Gbim29 Gbim66

24 24 21 24

0.667 0.542 0.857 0.625

0.675 0.614 0.721 0.531

0.220 0.100 0.898 0.802

0.013 0.117 0.189 0.176

0.037

MR

Gbim04 Gbim15 Gbim29 Gbim66

23 23 23 22

0.522 0.478 0.391 0.409

0.542 0.617 0.481 0.349

0.001*** 0.000*** 0.413 0.962

0.037 0.225 0.187 0.172

0.117

HB

Gbim04 Gbim15 Gbim29 Gbim66

20 21 21 21

0.500 0.333 0.429 0.667

0.664 0.417 0.543 0.499

0.002** 0.000*** 0.569 0.123

0.247 0.201 0.211 0.336

0.116

RM

Gbim04 Gbim15 Gbim29 Gbim66

21 21 21 21

0.571 0.429 0.762 0.619

0.658 0.595 0.681 0.563

0.350 0.245 0.741 0.001***

0.131 0.280 0.118 0.099

0.071

PO

Gbim04 Gbim15 Gbim29 Gbim66

20 20 20 20

0.800 0.550 0.650 0.600

0.748 0.803 0.710 0.774

0.294 0.149 0.575 0.575

0.070 0.315 0.085 0.225

0.168

SE

Gbim04 Gbim15 Gbim29 Gbim66

21 21 20 20

0.619 0.810 0.200 0.750

0.723 0.730 0.340 0.714

0.002** 0.209 0.000*** 0.692

0.144 0.109 0.412 0.051

0.076

HA

Gbim04 Gbim15 Gbim29 Gbim66

19 19 19 18

0.737 0.789 0.579 0.778

0.561 0.763 0.622 0.610

0.165 0.231 0.644 0.471

0.314 0.034 0.069 0.276

0.101

P gives the significance of a Chi-square test for Hardy–Weinberg equilibrium (*P < 0.05, **P < 0.01, shown for each locus and as a mean value for all loci (overall FIS).

***P

FIS

Overall FIS

< 0.001). The inbreeding coefficient (FIS) is

Table 3 – Genetic differentiation between the examined populations Population GH MR HB RM PO SE HA

GH

MR

– 0.000 0.162 0.096 0.055 0.116 0.195

ns – 0.253 0.111 0.074 0.168 0.219

HB

RM

PO

SE

HA

***

**

*

**

***

**

**

***

***

– 0.063 0.024 0.094 0.103

**

ns ns – 0.026 0.049

**

*

ns ns – 0.113

*

– 0.000 0.044 0.101

***

ns *

–

Pairwise RST values are given in the below diagonal and the corresponding significance of pair-wise population differentiation in the above diagonal (*P < 0.05, **P < 0.01, ***P < 0.001).

population growth on both sites. In 2001, subpopulation MR was still small (38 singing males), while subpopulation HB already reached comparatively high numbers (167 singing males). Nevertheless, the differentiation of subpopulation HB was rather surprising, as HB is only 1.3 km away from GH and connected by a number of subpopulations in the area between them. Therefore, we expected ongoing genetic exchange between these sites. However, our results suggest that dispersal between these sites is either low or has little influence on the genetic structure of subpopulation HB. These results show that a rapid population growth following the bottleneck after a natural colonization might promote differ-

entiation of populations, as predicted by simulations (Austerlitz et al., 1997; Excoffier, 2004). A similar pattern has been revealed during the recent range expansion process of Metrioptera roeselii in northern Germany (Hochkirch and Damerau, unpublished data) or malaria mosquitos (Anopheles arabiensis) in Nigeria (Onyabe and Conn, 2001). The population RM originated from a translocation of 213 cricket nymphs from different subpopulations at the Neusta¨dter Moor (mainly HB and GH) in 2001. This was reflected by the STRUCTURE output, in which the individuals from RM were assigned to these source populations and did not form a separate genetic cluster. Only few individuals from

B I O L O G I C A L C O N S E RVAT I O N

Table 4 – Estimated probability of the number of populations (K) for values of K = 1–9 and the corresponding standard deviation using STRUCTURE (Pritchard et al., 2000) K

lnPr(x/K)

1 2 3 4 5 6 7 8 9

1902.0 1742.3 1635.4 1571.6 1542.4 1513.7 1525.9 1571.9 1615.6

SD ±0.75 ±3.03 ±0.80 ±0.65 ±1.53 ±1.12 ±1.30 ±18.72 ±7.72

LnPr was estimated under the assumption that there was admixture between the examined populations. The highest likelihood was found for the model with six populations (K = 6).

RM were partly assigned to subpopulation MR (Fig. 3), which could be explained by the recent origin and rapid population growth of MR and its low differentiation from GH. Nevertheless, population RM was significantly differentiated from these source populations based upon RST values (Table 3). Translocated populations are known to retain the genetic signatures of their source population for quite some time (Mock et al., 2004). This makes assignment tests powerful tools to identify the source of immigrated or translocated individuals. At the same time, genetic drift may cause translocated populations to diverge from their source populations as a result of the loss of allelic diversity and shift in allele frequencies. Hence, translocated populations can show significant differences in allele frequencies, while the individuals within these populations are still assignable to a source population. The new population is separated from the Neusta¨dter Moor by wetlands and a river. Therefore, genetic exchange between these populations is unlikely in the near future (Hochkirch et al., 2007). Similar to the young natural subpopulations, RM showed a significant deviation from HWE, but only at one locus. Nevertheless, the translocated population had a higher effective population size than each of the subpopulations at the Neusta¨dter Moor. Even if all source subpopulations are pooled into one population, the effective population size of the translocated population remained relatively high. These results suggest that the mixture of founders

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

3065

from various subpopulations was an effective approach to minimize the loss of genetic diversity.

4.2.

Genetic diversity and genetic differentiation

Genetic variability is known to be significantly correlated with population fitness (Reed and Frankham, 2003). As habitat fragmentation decreases the genetic exchange among populations, it can lead to a loss of genetic diversity. Together with genetic drift and inbreeding, this can reduce the ability of a population to adapt to changes in its environment (Storfer, 1999; Frankham et al., 2005). Even if populations once had an identical genetic background, changes in allele frequencies may lead to differentiation over time (Frankham et al., 2005). Especially populations of small, flightless species, such as the field cricket, which have a low ability for dispersal, tend to get isolated rapidly. Keller et al. (2004) found a marked differentiation between subpopulations of the flightless ground beetle Abax parallelepipedus, which were only separated by a road. The field cricket is specialized on oligotrophic grassland and heathland habitats (Hochkirch, 1996; Kleukers et al., 1997), which have been common during the 19th century in northern Germany, but meanwhile have become rare and fragmented (Lu¨tkepohl, 2001; Fottner et al., 2004). It is likely that the remaining populations of the field cricket are relicts of a formerly more widespread species. Indeed, most of the examined cricket populations showed a significant genetic differentiation, which might also be influenced by the great geographic distances between the few remaining populations. The few similarities between some pairs of populations (HA and PO; PO and SE) cannot be explained by recent genetic exchange, which seems to be common in more mobile insects (Exeler et al., 2008). It is more likely, that this pattern is caused by the common origin of the populations in northern Germany and the former exchange among them. Dispersal between heathland patches probably decreased with increasing fragmentation. This process was accelerated by some general changes in land use. The loss of sheep herding at the end of the 19th century reduced the possibility of dispersal along herding routes (Lu¨tkepohl, 2001). In larger populations, some shared alleles might have been conserved. It is known that even in declining populations, a high degree of genetic diversity can be preserved, so that values for genetic diversity can exceed expectations by far (Streif et al., 2005). This may lead to combinations of a high number of alleles

Fig. 3 – Population assignment with STRUCTURE (performed assuming k = 6 and admixture between the sampled populations). Each individual is represented by a bar that is coloured responding to the possibility of origin within a certain population.

3066

B I O L O G I C A L C O N S E RVAT I O N

and high inbreeding coefficients, as we found in population PO. This population is currently strongly endangered by habitat loss. In 2006, only about 50 stridulating males were recorded on four sites (own observations). Many of the surrounding sites would benefit from habitat management. The inhabited sites are managed by a shepherd and population survival mainly depends on the continuity of the current land use. The strong differentiation of population HA to all other populations might be explained by a different colonization background. Postglacial colonization from different refugia has been documented for several European species and contact zones in Central Germany are common (reviewed by Hewitt, 2000). However, more cricket populations from other parts of Europe need to be sampled to test this hypothesis.

4.3.

Implications for translocation projects

Genetic analyses are a useful tool to evaluate the success of translocation projects. Up to now, the evaluation of most translocation projects is limited to the documentation of the survival and reproduction of the translocated individuals. Studies that investigate the genetic effects of translocations are still relatively sparse compared to the high number of relocation projects (Latch and Rhodes, 2005). So far, genetic analyses of translocated populations have mainly focused on fish and game species (e.g. Turkey: Mock et al., 2004; Latch and Rhodes, 2005; fishes: Fumagalli et al., 2002; Eldridge and Naish, 2007) or critically endangered vertebrates (e.g. Koala (Phascolarctos cinereus): Houlden et al., 1996; Takahe (Porphyrio hochstetteri): Jamieson and Ryan, 2000; Black footed ferret (Mustela nigripes): Wisely et al., 2003; Nailtail Wallaby (Onychogalea fraenata): Sigg, 2006). The existing studies indicate that the majority of translocations lead to a loss of genetic diversity (Houlden et al., 1996; Mock et al., 2004). There are examples were the subsequent additions of individuals from the source population helped to increase the genetic diversity in the translocated population (e.g. Maudet et al., 2002; Eldridge and Killebrew, 2008). However, supplementations can also further decrease the genetic diversity if individuals are supplemented that do not introduce new genetic diversity into the population (Sigg, 2006; Eldridge and Killebrew, 2008), e.g. if serial translocations are carried out (Mock et al., 2004). In some cases, the release of additional individuals or natural migration between populations can make it difficult to assess the effect of the initial translocation on the genetic diversity of the newly established population (Mock et al., 2004). Although inbreeding does not always lead to inbreeding depression, it increases the risk of homozygous recombination of deleterious recessive alleles (Reed et al., 2002). Inbreeding has been recognized as a potential threat to insect species (e.g. Duchateau et al., 1994; Cook and Crozier, 1995; Witkowski and Adamski, 1996; Saccheri et al., 1998). Several studies on inbreeding in crickets revealed negative effects, such as a reduced embryo viability, hatching success, nymphal survival, early fecundity or adult life span (Roff and DeRose, 2001; Roff, 2002; Jennions et al., 2004; Simmons et al., 2006; Drayton et al., 2007). Hence, inbreeding could influence population growth and survival of newly established populations (Roff, 1998). Females of some cricket species exhibit mechanisms of inbreeding avoidance, e.g. females of Telegryllus oceanicus and Gryllus

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

bimaculatus bias paternity in favour of unrelated males (Bretman et al., 2004; Simmons et al., 2006). However, such a strategy is not possible if a genetic bottleneck or founder event affected the gene pool within a population. To avoid negative effects of inbreeding or outbreeding during translocation projects, it is advisable to assess the genetic diversity of possible source populations prior to the translocation project. Genetic analyses allow to assess populations or individuals that should be excluded for a translocation project. If possible, candidates for a translocation or reintroduction should be taken from multiple source populations in order to enhance genetic diversity in the relocated population (Frankham et al., 2005). On the other hand, it is advisable to use individuals from a nearby area to avoid negative effects caused by local adaptations (Storfer, 1999). Genetic analyses should also be performed after the translocation has been carried out, in order to examine the genetic diversity of the newly established population and to avoid inbreeding depression. This might support decisions whether it is necessary to further augment the translocated population or not.

5.

Conclusions

As the strong fragmentation of G. campestris populations in northern Germany probably inhibits any natural colonization of newly created potential habitats, further translocation projects for this species are currently planned (Niemeyer, Brandt and Grein, pers. comm.). Our study shows that such translocation projects can be successful and that the genetic variability in a relocated population may be as high as in naturally derived populations. To minimize the risk of inbreeding, it is advisable to collect individuals from several subpopulations of a large population. Even though animal translocations have become a popular tool in nature conservation, translocations of insects are still sparse (Sarrazin and Barbault, 1996; Fischer and Lindenmayer, 2000; Dunn, 2005; Crone et al., 2007). Our results show that the field cricket project at the DHM was successful, as the translocated population reproduced successfully during the last five years, increased in population size considerably (Hochkirch et al., 2007) and exhibits a relatively high genetic diversity. In addition, we found no detectable negative effects on the source population, which increased to a new maximum of 2473 singing males in 2006.

Acknowledgements We are grateful to Nina Exeler for assistance in the laboratory, statistical advice and valuable comments on a previous version of the manuscript. Amanda Bretman and Tom Tregenza kindly provided information on microsatellite markers before they had been published. We would also like to thank Jana Deppermann, Julia Gro¨ning, Martin Husemann, Cordula Ju¨lch, Bernd Kaltwaßer and Isgard Lemke for helping us to collect samples. We are also grateful to Friedhelm Niemeyer, Anje Teerling and the staff of the BUND ‘‘Diepholzer Moorniederung’’ for the fruitful cooperation. Gu¨nter Grein provided us with detailed information on the existing cricket populations in northern Germany. Ulrike Coja helped us to perform the

B I O L O G I C A L C O N S E RVAT I O N

genetic analysis on the ABI Sequencer. Further, we are much obliged to Anselm Kratochwil for his constant support and encouragement throughout this project. The local and regional administrations (Land of Lower Saxony, NLWKN, Landkreis Diepholz) provided us with permissions to collect samples and access nature reserves. Finally, we owe great thanks to our financial supporters (Orthoperists’ Society, University of Osnabru¨ck).

R E F E R E N C E S

Austerlitz, F., Jung-Muller, B., Godelle, B., Gouyon, P.-H., 1997. Evolution of coalescence times, genetic diversity and structure during colonization. Theoretical Population Biology 51, 148– 164. Balloux, F., Lugon-Moulin, N., 2002. The estimation of population differentiation with microsatellite markers. Molecular Ecology 11, 155–165. Bretman, A., Wedell, N., Tregenza, T., 2004. Molecular evidence of post-copulatory inbreeding avoidance in the field cricket Gryllus bimaculatus. In: Proceedings of the Royal Society of London Series B-Biological Sciences. 271, 159–164. Bretman, A., Dawson, D.A., Hornsburgh, G.J., Tregenza, T., 2008. New microsatellite loci isolated from the field cricket Gryllus bimaculatus characterized in two cricket species Gryllus bimaculatus and Gryllus campestris. Molecular Ecology Resources 8, 1015–1019. Budrys, E., Pakalnisˇkis, S., 2007. The Orthoptera (Insecta) of Lithuania. Acta Zoologica Lithuania 17, 105–115. Cheyne, S.M., 2006. Wildlife reintroduction: considerations of habitat quality at the release site. BMC Ecology 6, 5. Cook, J.M., Crozier, R.H., 1995. Sex determination and population biology in the hymenoptera. Trends in Ecology and Evolution 10, 281–286. Crone, E.E., Pickering, D., Schultz, C.B., 2007. Can captive rearing promote recovery of endangered butterflies? An assessment in the face of uncertainty. Biological Conservation 139, 103– 112. Dawson, D.A., Bretman, A.J., Tregenza, T., 2003. Microsatellite loci for the field cricket, Gryllus bimaculatus and their cross-utility in other species of Orthoptera. Molecular Ecology Notes 3, 191– 195. Decleer, K., Devriese, H., Hofmans, K., Lock, K., Barenbrug, B. Maes, D., 2000. Voorlopige atlas en ‘‘rode lijst’’ van de sprinkhanen en krekels van Belgie¨ (Insecta, Orthoptera). Werkgroep Saltabel i.s.m. I.N. en K.B.I.N., Rapport Instituut voor Natuurbehoud 10, Brussel. Drayton, J.M., Hunt, J., Brooks, R., Jennions, M.D., 2007. Sounds different: inbreeding depression in sexually selected traits in the cricket Teleogryllus commodus. Journal of Evolutionary Biology 20, 1138–1147. Duchateau, M.J., Hoshiba, H., Velthuis, H.H.W., 1994. Diploid males in the bumble bee Bombus terrestris. Entomologia experimentalis et applicata 71, 263–269. Dunn, R.R., 2005. Modern insect extinctions, the neglected majority. Conservation Biology 19, 1030–1036. El Mousadik, A., Petit, R.J., 1996. High level of genetic differentiation for allelic richness among populations of the argan tree (Argania spinosa (L) Skeels) endemic to Morocco. Theoretical and Applied Genetics 92, 832–839. Eldridge, W.H., Killebrew, K., 2008. Genetic diversity over multiple generations of supplementation: an example from Chinook salmon using microsatellite and demographic data. Conservation Genetics 9, 13–28.

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

3067

Eldridge, W.H., Naish, K.A., 2007. Long-term effects of translocation and release numbers on fine-scale population structure among coho salmon (Oncorhynchus kisutch). Molecular Ecology 16, 2407–2421. Excoffier, L., 2004. Patterns of DNA sequence diversity and genetic structure after a range expansion: lessons from the infiniteisland. Molecular Ecology 13, 853–864. Exeler, N., Kratochwil, A., Hochkirch, Al., 2008. Strong genetic exchange among populations of a specialist bee, Andrena vaga (Hymenoptera: Andrenidae). Conservation Genetics 9, 1233– 1241. Fischer, J., Lindenmayer, D.B., 2000. An assessment of the published results of animal relocations. Biological Conservation 96, 1–11. Fischer, J., Lindenmayer, D.B., 2007. Landscape modification and habitat fragmentation: a synthesis. Global Ecology and Biogeography 16, 265–280. Fottner, S., Niemeyer, T., Sieber, M., Ha¨rdtle, W., 2004. Einfluss der Beweidung auf die Na¨hrstoffdynamik von Sandheiden. In: Keienburg, T., Pru¨ter, J., (Eds.), Feuer und Beweidung als Instrument zur Erhaltung magerer Offenlandschaften in ¨ kologische und sozioo¨konomische Nordwestdeutschland – O Grundlagen des Heidemanagements auf Sand- und Hochmoorstandorten. – NNA-Berichte. vol. 17, pp. 34– 43. Frankham, R., Ballou, J.D., Briscoe, D.A., 2005. Introduction to Conservation Genetics. Cambridge University Press, Cambridge. Fumagalli, L., Snoj, A., Jesensˇek, D., Balloux, F., Jug, T., Duron, O., Brossier, F., Crivelli, A.J., Berrebi, P., 2002. Extreme genetic differentiation among the remnant populations of marble trout (Salmo marmoratus) in Slovenia. Molecular Ecology 11, 2711–2716. Goudet, J., 1995. FSTAT: a computer program to calculate Fstatistics. Journal of Heredity 86, 485–486 (Version 1.2). Grein, G., 2000. Zur Verbreitung der Heuschrecken (Saltatoria) in Niedersachsen und Bremen. Informationsdienst Naturschutz Niedersachsen 20, 74–112. Grein, G., 2005. Rote Liste der in Niedersachsen und Bremen gefa¨hrdeten Heuschrecken mit Gesamtartenverzeichnis, 3 Fassung. Informationsdienst Naturschutz Niedersachsen 25, 1–20. Gro¨ning, J., Krause, S., Hochkirch, A., 2007. Habitat preferences of an endangered insect species, Cepero’s ground-hopper (Tetrix ceperoi). Ecological Research 22, 767–773. Hewitt, G.M., 2000. The genetic legacy of the Quaternary ice ages. Nature 405, 907–913. Hochkirch, A., 1996. Die Feldgrille (Gryllus campestris L., 1758) als Zielart fu¨r die Entwicklung eines Sandheidereliktes in Nordwestdeutschland. Articulata 11, 11–27. Hochkirch, A., Adorf, F., 2007. Effects of prescribed burning and wildfires on Orthoptera in Central European peat bogs. Environmental Conservation 34, 225–235. Hochkirch, A., Witzenberger, K.A., Teerling, A., Niemeyer, F., 2007. Translocation of an endangered insect species, the field cricket (Gryllus campestris Linnaeus, 1758) in northern Germany. Biodiversity and Conservation 16, 3597–3607. Hochkirch, A., Ga¨rtner, A.C., Brandt, D.T., 2008. Effects of forestdune ecotone management on the endangered heath grasshopper Chorthippus vagans (Orthoptera: Acrididae). Bulletin of Entomological Research 98, 449–456. Houlden, B.A., England, P.R., Taylor, A.C., Greville, W.D., Sherwin, W.B., 1996. Low genetic variability of the koala Phascolarctos cinereus in south-eastern Australia following a severe population bottleneck. Molecular Ecology 5, 269–281. Ingrisch, S., Ko¨hler, G., 1997. Rote Liste der Geradflu¨gler (Orthoptera s. l.). In: Binot, M., Bless, R., Boye, P., Gruttke, H., Pretscher, P., (Eds.), Rote Liste gefa¨hrdeter Tiere Deutschlands.

3068

B I O L O G I C A L C O N S E RVAT I O N

Schriftenreihe fu¨r Landschaftspflege und Naturschutz. vol. 55, pp. 252–254. IUCN, 1998. IUCN guidelines for re-introductions. IUCN, Gland & Cambridge. Jamieson, I.G., Ryan, C.J., 2000. Increased egg infertility associated with translocating inbred takahe (Porphyrio hochstetteri) to island refuges in New Zealand. Biological Conservation 94, 107–114. Jennions, M.D., Hunt, J., Graham, R., Brooks, R., 2004. No evidence for inbreeding avoidance through postcopulatory mechanisms in the black field cricket, Teleogryllus commodus. Evolution 58, 2472–2477. Keller, I., Nentwig, W., Largiade`r, C.R., 2004. Recent habitat fragmentation due to roads can lead to significant genetic differentiation in an abundant flightless ground beetle. Molecular Ecology 13, 2983–2994. Kleukers, R., Nieukerken, E.V., Ode´, B., Willemse, L., Wingerden, W.V., 1997. De Sprinkhanen en Krekels van Nederland (Orthoptera). Nederlandse Fauna I. KNNV Uitgeverij & EISNederland, Leiden. Ko¨hler, G., Reinhardt, K., 1992. Beitrag zur Kenntnis der Feldgrille (Gryllus campestris L.) in Thu¨ringen. Articulata 7, 63–76. Latch, E.K., Rhodes Jr, O.E., 2005. The effects of gene flow and population isolation on the genetic structure of reintroduced wild turkey populations: are genetic signatures of source populations retained? Conservation Genetics 6, 981–997. Lu¨tkepohl, M., 2001. Die Entwicklung von Sandheiden, Moorheiden und Ackerbrachen unter dem Einfluß der Beweidung durch Heidschnucken. Natur- und Kulturlandschaft 4, 217–223. Marshall, J.A., Haes, E.C.M., 1990. Grasshoppers and allied insects of Great Britain and Ireland. Harley Books. Colchester, Essex. Maudet, C., Miller, C., Bassano, B., Breitenmoser-Wursten, C., Gauthier, D., Obexer-Ruff, G., Michallet, J., Taberlet, P., Luikart, G., 2002. Microsatellite DNA and recent statistical methods in wildlife conservation management: applications in Alpine ibex (Capra ibex (ibex)). Molecular Ecology 11, 421–436. Mock, K.E., Latch, E.K., Rhodes Jr, O.E., 2004. Assessing losses of genetic diversity due to translocation: long-term case histories in Merriam’s turkey (Meleagris gallopavo merriami). Conservation Genetics 5, 631–645. Onyabe, D.Y., Conn, J.E., 2001. Population genetic structure of the malaria mosquito Anopheles arabiensis across Nigeria suggests range expansion. Molecular Ecology 10, 2577–2591. Peakall, R., Smouse, P.E., 2006. GENALEX 6: genetic analysis in Excel. Population genetic software for teaching and research. Molecular Ecology Notes 6, 288–295. Pearce-Kelly, P., Jones, R., Clarke, C., Atkin, P., Cunningham, A.A., 1998. The captive rearing of threatened Orthoptera: a comparison of the conservation potential and practical considerations of two species’ breeding programmes at the Zoological Society of London. Journal of Insect Conservation 2, 201–210. Primack, R.B., 2002. Essentials of Conservation Biology. Sinauer Associates, Sunderland. Pritchard, J.K., Stephens, M., Donnelly, P., 2000. Inference of population structure using multilocus genotype data. Genetics 155, 945–959. Proess, R., Meyer, M., 2003. Rote Liste der Heuschrecken Luxemburgs. Bulletin de la Socie´te´ des Naturalistes Luxembourgeois 104, 57–66. Reed, D.H., Frankham, R., 2003. Correlation between fitness and genetic diversity. Conservation Biology 17, 230–237.

1 4 1 ( 2 0 0 8 ) 3 0 5 9 –3 0 6 8

Reed, D., Briscoe, D.A., Frankham, R., 2002. Inbreeding and extinction: the effect of environmental stress and lineage. Conservation Genetics 3, 301–307. Remmert, H., 1992. Die Populationsdynamik von Feldgrillen und ¨ kologie: ein Lehrbuch. ihre Ursachen. In: Remmert, H. (Ed.), O Springer Verlag, Berlin, pp. 196–200. Roff, D.A., 1998. Effects of inbreeding on morphological and life history traits of the sand cricket, Gryllus firmus. Heredity 81, 28–37. Roff, D.A., 2002. Inbreeding depression: tests of the overdominance and partial dominance hypotheses. Evolution 56, 768–775. Roff, D.A., DeRose, M.A., 2001. The evolution of trade-offs: effects of inbreeding on fecundity relationships in the cricket Gryllus firmus. Evolution 55, 111–121. Saccheri, I., Kuussaari, M., Kankare, M., Vikman, P., Fortelius, W., Hanski, I., 1998. Inbreeding and extinction in a butterfly metapopulation. Nature 392, 491–494. Sarrazin, F., Barbault, R., 1996. Reintroduction: challenges and lessons for basic ecology. Trends in Ecology and Evolution 11, 474–478. Saunders, D.A., Hobbs, R.J., Margules, C.R., 1991. Biological consequences of ecosystem fragmentation: a review. Conservation Biology 5, 18–32. Sigg, D.P., 2006. Reduced genetic diversity and significant genetic differentiation after translocation: comparison of the remnant and translocated populations of bridled nailtail wallabies (Onychogalea fraenata). Conservation Genetics 7, 577–589. Simmons, L.W., Beveridge, M., Wedell, N., Tregenza, T., 2006. Postcopulatory inbreeding avoidance by female crickets only revealed by molecular markers. Molecular Ecology 15, 3817– 3824. Slatkin, M., 1995. A measure of population subdivision based on microsatellite allele frequencies. Genetics 139, 457–462. Storfer, A., 1999. Gene flow and endangered species translocations: a topic revisited. Biological Conservation 87, 173–180. Streif, R., Audiot, P., Foucart, A., Lecoq, M., Rasplus, J-.Y., 2005. Genetic survey of two endangered grasshopper subspecies, Prionotropis hystrix rhodanica and Prionotropis hystrix azami (Orthoptera, Pamphagidae): within- and between-population dynamics at the regional scale. Conservation Genetics 7, 331– 344. Tallmon, D.A., Koyuk, A., Luikart, G., Beaumont, M.A., 2008. ONeSAMP: a program to estimate effective population size using approximate Bayesian computation. Molecular Ecology Resources 8, 299–301. Teerling, A., Hochkirch, A., 2002. 10 Jahre Schutz der Feldgrille in der Diepholzer Moorniederung – Ziele, Erfolge und ¨ kologie 32, Perspektiven. Verhandlungen der Gesellschaft fu¨r O 407. Van Oosterhout, C., Hutchinson, W.F., Wills, D.P.M., Shipley, P., 2004. Micro-Checker: software for identifying and correcting errors in microsatellite data. Molecular Ecology Notes 4, 535–538. Wisely, S.M., McDonald, D.B., Buskirk, S.W., 2003. Evaluation of the genetic management of the endangered black-footed ferret (Mustela nigripes). Zoo Biology 22, 287–298. Witkowski, Z., Adamski, P., 1996. Decline and rehabilitation of the apollo butterfly Parnassius apollo (Linnaeus, 1758) in the Pieniny National Park (Polish Carpathians). In: Settele, J., Margules, C., Poschol, P., Henle, K. (Eds.), Species Survival in Fragmented Landscapes. Kluwer Academic Publishers, Dordrecht, pp. 7–14.