Springer 2006
Hydrobiologia (2006) 559:113–122 DOI 10.1007/s10750-005-1024-5
Primary Research Paper
The decline of the endangered populations of the native freshwater crayfish (Austropotamobius pallipes) in southern Spain: it is possible to avoid extinction? Jose Marı´ a Gil-Sa´nchez & Javier Alba-Tercedor* Department of Animal Biology and Ecology, Faculty of Sciences, University of Granada, 18071-Granada, Spain (*Author for correspondence: E-mail:
[email protected]) Received 18 October 2004; in revised form 29 June 2005; accepted 31 July 2005
Key words: Austropotamobius, conservation, extinction, Southern Europe, Spain
Abstract The southeastern mountains of Spain represent the southernmost limit of the genus Austropotamobius and the species A. pallipes (Lereboullet). The taxonomic position of this isolated crayfish in southern Spain is not clear, being genetically close to A. italicus, but morphologically distinct. A severe decline occurred during the 1980s, especially due to expansion of the alien species Procambarus clarkii, a North American freshwater crayfish and a vector of the aphanomycosis disease. In order to design a strategy for native crayfish conservation, recent trends in native crayfish populations, influence of isolation and habitat variables on their survival and possibilities for their recovery through restocking were studied. A decline in populations was observed between 1991 and 2002 (from 26 to 14 populations), and a total extinction could be predicted within the next 13 years. Two well-defined periods of rapid extinction rates were detected, 1991–1994 (1.25 populations/year) and 1999–2002 (1.50 populations/year). Main causes of extinction for these two periods were invasion by P. clarkii and mortality by unknown causes, respectively. The isolation variables had some positive effects on survival of populations but these cannot offer a sufficient guarantee, since several cases of extinction can be affected by a large component of stochastic factors, including random catastrophes. On the other hand, survival trials and restocking experiments showed that it was possible to recover lost habitats, when P. clarkii was absent and environmental conditions were good. It is concluded that it is possible to avoid extinction of the native crayfish populations; however, the conservation strategies must be based on an urgent stocking/restocking program.
Introduction The white-clawed crayfish, Austropotamobius pallipes (Lereboullet), is a European freshwater crayfish placed on the Red List of Threatened Animals of the IUCN (International Union for Conservation of Nature) as a vulnerable species (Baillie & Groombridge, 1996). It is also cited in annexes II and IV of European Community Directives for the Conservation of Natural Habitats and Wild Flora and Fauna, as a species requiring special conservation measures. Recent
studies considered the Iberian freshwater crayfish as a subspecies (Austropotamobius pallipes italicus), but it is considered a separate species (A. italicus) by some authors (Santucci et al., 1997; Grandjean et al., 2000, 2002; Holdich, 2002). During the recent CRAYNET meeting held in Kilkenny, Ireland, in June 2003, it was agreed to consider A. pallipes as a supraspecific taxon with two separate species: A. pallipes (Lereboullet) and A. italicus (Faxon), considering that populations of southern Spain belong to A. italicus (Alonso, person. comm.). However, for the
114 moment we decided to maintain it as A. pallipes, because the morphological data of taxonomic value reveal that the crayfish of the southern Iberian Peninsula are well distinguished from the Italian populations, especially by the ratio length of apex/length of rostral morphology, with a mean value of 0.29 for A. italicus and 0.22 for A. pallipes (Grandjean et al., 2000), but less than 0.15 for crayfish of the mountains of southeast Spain (Gil-Sa´nchez, 1999). On the other hand, there are contradictory opinions about the origin of the Spanish populations. Grandjean et al. (2001), based on the low genetic diversity, considered that it might be an introduced species. Trontelj et al. (2005), comparing mitochondrial DNA sequences concluded that the Iberian crayfish belong to the NW Italy A. pallipes mitochodrial DNA haplotype group, and consider that the Iberian Peninsula was stocked artificially from Northern Italy. However, Santucci et al. (1997), based on the high allozyme diversity, considered it as a glacial relict. Moreover, one population of southeastern Spain had the highest allozyme diversity among 31 populations studied from across the entire distribution area in Europe, and this population together with another from northeastern Spain showed a number of alleles not detected elsewhere (Santucci et al., 1997). The native crayfish is a critically endangered species within the southeast Spanish mountains, especially due to introduction and expansion of the North-American red-swamp crayfish, Procambarus clarkii (Girard) (Gil-Sa´nchez & Alba-Tercedor, 2002; Gil-Sa´nchez et al., 2002). This alien species is an important vector of the aphanomycosis (Die´guez-Uribeondo & So¨derha¨ll, 1993; Die´guez-Uribeondo et al., 1995), a lethal disease caused by the oomycete Aphanomyces astaci Schikora and named ‘crayfish plague’ that has decimated most populations of the European freshwater crayfish (Alderman & Polglase, 1988; Vogt, 1999; Gherardi & Holdich, 1999; Barbaresi & Gherardi, 2000). In fact during the 80s of the 20th century, 90% loss of the rivers occupied by native crayfish in southeast Spain was observed, and nowadays they only inhabit isolated and remote headwater areas where P. clarkii is absent because of ecological barriers (Gil-Sa´nchez & Alba-Tercedor, 2002; Gil-Sa´nchez et al., 2002). On the other hand, southeast Spain represents the southern limit for the genus Austropotamobius. It
has a dry Mediterranean climate, therefore, it represents one of the most extreme habitats for the European freshwater crayfish. In this sense, negative environmental factors unusual for other northern habitats, such as large drought periods, may bring important negative impacts on these marginal populations. Thus the extinction of these populations corresponds to an irreparable loss of the populations of the southern limit distribution of the Austropotamobius crayfish in Europe. The aim of this study was firstly to investigate the recent tendencies and causes of extinction of the relict populations of the native crayfish in southeast Spain. Secondly, importance of habitat isolation for conservation was analyzed, considering that the main local risk was the expansion of aphanomycosis from lower river stretches occupied by P. clarkii. Thirdly, resulting from the previous research objectives, a restocking management strategy was successfully demonstrated in the field, in order to investigate the chances for conservation of the native freshwater crayfish of the southeast Iberian Peninsula.
Study area The study area was located in Granada Province (12 351 km2), within the Baetic Mountains in southern Spain. Elevations ranged from sea level at the Mediterranean Sea, to 3,492 m a.s.l. in the Sierra Nevada. Calcareous zones are well represented. The climate is Mediterranean, with mean annual rainfall of 600 mm but ranging from 350 mm in some eastern depressions up to 1,000 mm in some mountains (Sierra Nevada). Most watercourses belong to the Guadalquivir River basin. The watercourses have a high seasonality, typical of Mediterranean rivers (Gasith & Resh, 1999). Combined with the influence of water extraction for agriculture, these watercourses drastically diminish in water flow during summer, sometimes even drying up. On the other hand, winter and spring floods are present with high rain levels. In general, water quality, as well as river habitats, are very good at the headwaters, but deteriorate downstream as a result of urban sewage and agriculture (Zamora-Mun˜oz, 1992; Picazo-Mun˜oz, 1995; Zamora-Mun˜oz & Alba-Tercedor, 1996; Alba-Tercedor et al., 2004).
115 Material and methods
Analysis of isolation variables
Monitoring of crayfish populations
Importance of the environmental variables describing both geographical and biological isolation of populations was analyzed using three categorical and seven quantitative variables. The categorical variables are: presence of barriers, type of barriers (dry stretches as long barriers, and waterfalls or anti-erosion dams as short barriers, and fish presence (a risk factor since fish can be a vector transporting A. astaci spores, Alderman et al., 1987). The quantitative variables are: altitude as a biological barrier for P. clarkii (GilSa´nchez & Alba-Tercedor, 2002); distances to: the nearest population of P. clarkii, the nearest village, the nearest paved road and the nearest unpaved road; and hectares of cultivation within 1 km of radius as a human presence indicator (another risk factor by transporting P. clarkii or A. astaci spores or by habitat alterations and/or pollution), and size (habitat length occupied by crayfish). The habitat variables were compared between extinct populations and surviving populations in 2002 (Table 1). For five populations, a decline of population size but not complete extinction was observed. In these cases, the crayfish in the downstream stretch disappeared and so the population in the downstream was treated as ‘extinct’ whereas the population in the upstream survivalstretch was treated as ‘surviving.’ The relationship between the number of survival years from 1991 to 2002 (values from 0 up to 11) of each population and the quantitative variables was also studied by non-parametric correlations.
Headwaters areas with springs, brooks and streams in calcareous zones were selected and an intensive sampling program has been carried out since 1991 by hand-catching and trapping every summer (for details see Gil-Sa´nchez & AlbaTercedor, 2002), the optimal season for looking for local crayfish (Gil-Sa´nchez, 1999). For this purpose, hand sampling and/or baited fish-traps were used following the normal procedures used by others (Brown & Bowler, 1977; Arrignon, 1983; Reynolds & Matthews, 1993; Holdich & Domaniewski, 1995; Lappalainen & Pursiainen, 1995; Gherardi et al., 1996). Cylindrical fish traps of 50 20 cm were made with plastic net of 1 1 cm mesh size. These traps had two free opposite openings to allow entry (Fja¨lling, 1995) and one blocked central opening to allow extraction. In summer (July and August), three traps (baited with frozen squid) were set during the evenings (between 17:00 and 21:00 h) at each sampling site and collected the following day (between 08:00 and 12:00 h). Absence of crayfish was assumed when no crayfish were detected with high sampling effort (i.e., more than 1 h of hand sampling for little brooks). Moreover, annual monitoring of all known populations was conducted through 1996– 2002, and some populations were also visited even earlier between 1992 and 1995. The populations were visited at least once each summer and the occupied habitat length was estimated by multiple sampling of sites, for cases of large populations (between three and five sites/km). During the monitoring of populations, cause of extinction was registered whenever possible (such as P. clarkii invasion, drought, mortality events, etc.). Laboratory analyses of dead crayfish were carried out in order to look for hyphae of A. astaci, following Cerenius et al. (1987) indications for optical microscopy observation. However, this identification method is based only on morphology of the hyphae, so results were carefully evaluated since more research is needed for the fungus identification (Cerenius et al., 1987; Die´guez-Uribeondo & So¨derha¨ll, 1993).
Survival trials and restocking experiments Restocking has been proposed as an important management strategy for conservation of the European crayfish (e.g., Die´guez-Uribeondo et al., 1997b; Holdich & Rogers, 1997). Possibilities for developing a restocking program were firstly studied by looking for potential adequate brooks and streams. Following the IUCN (1987) recommendations for re-introduction projects, favourable conditions for restocking were determined by: water quality, i.e., no pollution and calcareous areas (Laurent, 1988), isolation from P. clarkii,
116 Table 1. Quantitative habitat variables compared between extinct and surviving populations (U-test) and relationship of each variable with the number of survival years during the study period (Rs) Variable
Extinct populations
Survival populations
Mean
SD
Range
Mean
916.1
244.0
510–1450
1046.9 228.7
12.3
11.2
0–30
Dist. to village (km)
1.9
2.4
0–5.75
Dist. to paved road (km)
1.0
1.5
0–5.75
0.03 401.6
0–0.1 0–1514
Altitude (m a.s.l.) Dist. to P. clarkii pop. (km)
Dist. to unpaved road (km) 0.01 cultivations 1 km radius (ha) 345.2 Size (km)
2.0
2.9
0.01–8.5
i.e., presence of biological and physical barriers (Gil-Sa´nchez & Alba-Tercedor, 2002) and isolation from human presence, i.e., more than 2 km to the nearest road and less than 100 ha of cultivation within 1 km around (present study). Twenty five freshwater stretches were identified and 11 of them were selected as experimental stretches for survival trials. The survival trials were carried out using plastic mesh cages (Geddes et al., 1993) of 50 cm length. One to six crayfish were used per cage (n = 48) and one to three cages per experimental freshwater stretch. Cages were set at each experimental site into pools about 60 cm of depth, and controlled once per week. Food (fish or vegetable material) was supplied when necessary. Experiments were terminated after 9 months. Also, three other survival trials were carried out in areas occupied by P. clarkii populations, using three cages each containing two or three native crayfish. Three streams with successful survival trials were selected for restocking experiments, mainly by translocation within the same river basin. For the restocking experiments, all crayfish were collected by cylindrical traps. Table 2 shows the data on number of crayfish, Table 2. Summary of translocations Restocking Year Number Sex-ratio Size of Density point
of
(males/
crayfish females)
pool
(cray
(m2)
fish/m2)
A
1997
52 0.79
16
3.33
B
1998
43 0.79
10
4.30
C
1998
109 1.09
20
5.45
SD
U
Z
P
Rs survival years
Range 86.5
)1.29 0.19
0.32 n.s.
8.0 0.25–31
82.0
)1.49 0.14
0.40 p < 0.05
3.8
2.7
0.5–9.8
56.5
)2.48 0.13
0.55 p < 0.05
2.7
1.8
0.5–6.7
45.5
)2.91 0.003
0.53 p < 0.05
83.0 43.0
)1.42 0.073 )3.61 0.002
0.40 p < 0.05 )0.51 p < 0.05
114.0
)0.19 0.84
17.9
0.1 0.1 68.2 145.9 1.4
500–1430
0–0.65 0–510
1.3 0.02–4.5
0.02 n.s.
density and year of translocation. Densities and sex-ratio were similar to those local natural populations, with crayfish sizes ‡5 cm of total length (Gil-Sa´nchez, 1999). The restocked populations were monitored monthly by hand sampling during the first year and then annually monitored like other wild populations. Data analyses Chi-square test on contingency tables was used for studying patterns with the categorical variables. Owing to the small sample sizes, non-parametric tests were employed for quantitative variables, both Mann–Whitney U-test and Spearman rank correlation. All data were analyzed using the Statistica software package and probability levels were set at p = 0.05.
Results Population trends and extinction causes A total of 26 populations were present in 1991, but only 14 in 2002. Therefore, a strong decline of native crayfish was observed during the study period, )46.1% for the number of populations and )64.4% for the length of occupied habitat (Fig. 1). The extinction rate was 1.09 populations/year. Thus, without any conservation program, total extinction would occur in 13 years or in 2015. Moreover, observed trends were well explained by
117 rates. The first was related to P. clarkii invasions and drought and the second was defined by mortality events with unknown causes (Table 3). There were statistical differences for freshwater habitat type between extinct and surviving populations (v2 = 12.82, p = 0.0016; f.d. = 2), so springs were the most vulnerable habitat, whereas brooks were the best habitat (Fig. 2). Influence of isolation variables Barriers were present in 100% of the surviving populations and in 82.3% of the extinct ones (v2 = 19.35, p = 0.0001, f.d. = 1). There were no differences (v2 = 4.17, p = 0.057; f.d. = 1) between the types of barriers: large barriers (dry stretches) were present in 50% of the surviving populations and in 35.7% of the extinct ones. For fish presence (Salmo trutta Linnaeus, Barbus sclateri Gu¨nther, Squalius pyrenaicus (Gu¨nther), Chondostroma polylepis Steindachner and Cobitis paludica (de Buen), there were no differences (v2 = 4.42, p = 0.05, f.d. = 1) between the surviving populations (50% with fishes) and the extinct populations (35.29% with fish). The only two quantitative variables that differed between the extinct and surviving populations were distance to the nearest paved road and surface area of cultivation: the extinct populations had lower distances to the nearest road and higher amounts of cultivations (Table 1). Analysis of relationships between the number of survival years and
60
30
50
25
40
20
30
15
20
10
10
5
0
number of populations
km occupied
two regression equations (number of populations: y = )0.808 x + 25.66, R = 0.93, p < 0.0001; occupied habitat: y = )8.08 + e(4.2)0.077x), R = 0.96, p < 0.0001) predicting total extinction in 15–19 years or between 2017 and 2021. Extinction rate was not constant over time, with three well-defined periods: 1.25 populations/year between 1991 and 1994, 0.25 population/year between 1995 and 1998 and 1.50 populations/year between 1999 and 2002 (Fig. 1). Causes of total extinction of population or reduction of crayfish stretches (i.e., reduction of the area occupied by a crayfish population) were known for 8 cases, the most important being the fast substitution of the native species by P. clarkii, followed by the total drying up of the occupied stretches (Table 3). One small population living in a little spring of only 30 m2 became extinct due to sediment saturation by sand. Only one fungal disease was detected, in a massive mortality event associated with a Saprolegnia sp., reducing greatly the stretch occupied by one crayfish population (from 3.5 km to 500 m.). For the other 9 cases, extinction or stretch reduction was usually related to mortalities from unknown causes. Laboratory analyses of 13 crayfish from 4 populations affected by massive mortality did not reveal hyphae of Aphanomyces astaci. One of the 4 populations was affected by Saprolegnia sp., these results were confirmed by Die´guez-Uribeondo et al. (1997a). Different causes were observed during the two detected periods of high extinction
0 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002
Figure 1. Decline of the native crayfish populations between 1991 and 2002. Line represents the number of populations and bars represent occupied habitat size.
118 Table 3. Causes of extinction of crayfish populations and reductions in crayfish river stretches Extintion/contraction causes
Number of
Stretches
1991–1996
1997–2002
Total number (%)
populations P. clarkii invasion
4
0
4
0
4–23.52
Drought Saturation by fine sediments
0 1
2 0
2 0
0 1
2–11.76 1–5.88
Diseases
0
1
0
1
1–5.88
Massive mortality by unknown causes
5
1
2
4
6–35.29
Unknown causes (no dead animals were detected)
2
1
0
3
3–17.64
12
5
8
9
Total (N)
quantitative variables showed significant results for all variables but two, altitude and size. Survival rates were positively related to the distances to the nearest P. clarkii population, village and paved and unpaved roads and negatively related to the cultivations (Table 1). Survival trials and restocking experiments For the three survival trials located in the streams occupied by P. clarkii, all the crayfish (n = 8) died between 1 and 3 months (Fig. 3). All of them showed many hyphae morphologically similar to these of A. astaci. Six survival trials from the 11 experimental stretches were frustrated by factors independent of habitat suitability (human disturbance or floods). Thus, only seven trials were useful for the first 4 months, six for the five first months and five trials for the rest of the experiments. Figure 3 shows the percentage of survival related to time.
80
N=23
EXTINTC PRESENT
70 60 50
30
One trial that reached 9 months was carried out using three cages, with only one of them reached 9 months and the other two 1.25 months (frustrated by human) and 2.25 months (by mortality associated to moulting), respectively. The two trials that only reached 4 months failed due to unknown causes, but only one cage was used for each one. Among the mortality causes detected (18 crayfish), the most important was related to death during moulting (10). The other ones died from unknown causes (6), entangling the chelae in the cage netting (1) and trampling by cows (1). No hyphae of A. astaci were observed in the dead crayfish (n = 18) according to our analysis. Two restocking experiments were successful (A and B of Table 2), with not only presence detected during the monitoring (1997–2002) but also with maximum habitat expansion by 1 km and a minimum by 300 m. Expansion rates were 200 m/year and 75 m/year, respectively.
Discussion
90
40
17
N=5
N=3
20 10 0 SPRINGS
BROOKS
STREAMS
Figure 2. Percentage of habitat occupied by extinct populations and by surviving populations of native crayfish.
The present data show that the native crayfish of the south-eastern mountains of Spain have a high probability of becoming extinct soon. The negative impact of P. clarkii introduction was important (Gil-Sa´nchez & Alba-Tercedor, 2002), although this alien species did not appear to be a direct problem during the last extinction period observed since 1996. Expansion of P. clarkii to the headwater areas is ecologically limited and, within our study area, no stable populations have been detected at altitudes higher than 820 m a.s.l. (Gil-Sa´nchez & Alba-Tercedor, 2002). However,
119 120
% of survival
100 80 60 40 20 0 1
2
3
4
5
6
7
8
9
month Figure 3. Survival curves for the survival trials; circles: survival trials carried out within water courses with P. clarkii populations (n = 3); squares: survival trials carried out within potential habitats for recovering native crayfish (n = 7).
there is no doubt that lower river stretches are an important reservoir for aphanomycosis, since P. clarkii is widely distributed there. Spreading of spores of A. astaci by human or even aquatic animals is an important risk (Alderman & Polglase, 1988; Lodge et al., 2000), and it may be the reason for the relative importance of the isolation variables. Although Aphanomyces astaci could not be detected during mortality events, sometimes it is very difficult to see hyphae by optical analysis (Die´guez-Uribeondo, pers. comm.). But not only was A. astaci infection a potential disease affecting crayfish, also the presence of Saprolegnia sp. was revealed. This fungal species is sometimes known to decimate crayfish populations (Die´guez-Uribeondo et al., 1994, 1997a). Moreover, one extinct population was affected by the porcelain disease caused by Thelohania contejeani Henneguy and one was infected by Psorospermiun haeckeli Hilgendorf (Gil-Sa´nchez, 1999). Both the pathogenics could have a synergic lethal effect in conjunction with other stressing conditions such as large floods which destroyed a part of the channel-bank refuges for crayfish; important features for bank protection and crayfish refuge like riparian trees and shrubs (Smith et al., 1995). For drought, the second important known cause of the observed extinctions, it is probable that the present populations will be not affected by this, because these populations are located within
springs or brooks with good water levels and flow regimes, taking the Mediterranean climate context into account. In fact, no new extinction cases of these populations were observed since 1995. The possible effects of fish presence do not appear to be important. However, for all cases of fish presence, both brown trout (S. trutta) and barbel (B. sclateri) were isolated in the downstream area by barriers (weirs and anti-erosion dams), so it could not be a crayfish plague vector from the downstream areas occupied by P. clarkii. The presence of barriers, but not the type of barriers, was a limitation to extinction, although it was not enough by itself. In this sense, the two quantitative variables with significant differences between the lost populations and the stable ones were related with human presence, and humans can be an important vector for transmission of diseases independent of the studied barriers. Thus, the distance to paved roads could be related to expansion of diseases by man, as through infected fishing materials. It is known that zoospores of A. astaci can be transported on damp material (Alderman & Polglasse, 1988). Furthermore, humans can contribute to illegal stockings of P. clarkii even in inadequate habitats for this American species (e.g., mountain brooks), as was confirmed for at least three cases. Although these P. clarkii introductions fortunately failed, it could have been sufficient for the translocation of A. astaci from infected areas. On
120 the other hand, cultivations can affect crayfish through chemical contamination (Hogger, 1988; Holdich & Lowery, 1988a). The isolation variables showed some importance for survival of the crayfish populations, but it was clear that isolation was not a guarantee sufficient to avoid extinction. It appears that the last period which registered the highest extinction rates (since 1998) was a typical case of environmental stochasticity with random catastrophes (Lande, 1993). So, unfortunately new extinctions may be expected, as the regression curves predicted. There are very few published data on crayfish restocking within the Mediterranean areas (Die´guez-Uribeondo et al., 1997b). In northern Portugal 91.6% of 24 restocking experiments with native A. pallipes were without success because of P. clarkii presence (Correia et al., 1996). In the case of survival trials, one experiment carried out in the Murray River basin (Australia) with Euastacus armatus (von Martens) offered similar results to ours, with a 50% survival rate (n = 6 cages) over the first year (Geddes et al., 1993). Mortality observed for our trials was usually related to moulting, a critical event when individuals are exposed to several negative factors, such as stress from imprisonment in cages. Cannibalism is a frequent phenomenon among crayfish during moulting (Lowery, 1988). While it could be a cause of death since three individuals were found eaten, it was not possible to determine the real cause of death, because these individuals might be dead prior to the cannibalism. For the Australian experiment, low water flow and fine sediments were suggested as probable causes of mortality (Geddes et al., 1993). Nevertheless for our study area, the survival trials, together with restocking experiments, showed that it was possible to recover the lost habitats, where P. clarkii was still absent through ecological barriers. Restocked crayfish showed good adaptation and fast expansion resulting in good breeding success related to habitat quality. In the failed restocking experiment, the crayfish lived for a while as they were easily captured and the monthly positive observations continued, but they suddenly disappeared and the observations have become negative since. No dead individuals were observed and thus poaching was suspected because it was also not rare for brown trout (Salmo trutta) within the
same area. At least, minimum survival time was 7.5 months and breeding was confirmed, but berried females disappeared during March without any chance for breeding success since very small juveniles (L2) are not present until June in that area (Gil-Sa´nchez, 1999). Conservation implications The present data showed that it was possible to avoid extinction of the native crayfish, within a short time scale. However, it must be based on an urgent management strategy, since isolation as a strategy is not a guarantee sufficient for their conservation, considering the fast extinction rate. The best conservation strategy should be to obtain as high a number of populations as possible, since the risk of extinction of each population is mainly affected by stochastic factors, with almost no chance of prevention through management. However, loss of populations from habitats in headwater areas can easily be recovered. It is logistically easy to restock by translocations within suitable habitats previously tested by survival trials. The two new restocked populations diminished the last extinction rate (since 1998) from 1.5 populations/year to 1.0 population/year. Therefore, annual extinction rates can be buffered by this type of management within up to 30 brooks or streams. These new populations obviously will be affected by the observed stochasticity of extinction probability, but the stocking/restocking program represents the best chance for the future of the crayfish. Quick decision and application are also necessary in order to avoid the continuous loss of genetic material. Thus, the two restocked populations were founded by individuals from two populations nowadays extinct. Thus, a part of the genetic variation was saved and similar actions must be carried out at least for the main populations of native crayfish within the region.
Acknowledgements We thank the ‘Consejerı´ a de Medio Ambiente, Junta de Andalucı´ a’ for financial support during 1996–1997, Dr Fernando Alonso and Dr Javier Die´guez for having provided us valuable information, David Nesbitt for early language help and
121 Dr Julian Reynolds (Trinity Collegue, Univ. of Dublin) for his valuable help with the final linguistic corrections of the manuscript. Cristina Sa´nchez helped much in the monitoring. We are especially indebted to two anonymous referees who greatly improved the manuscript with their suggestions and thorough corrections. This study was benefited by the project GUADALMED I: HID98-0323-C05-05 and GUADALMED II: REN 2001-3438-C07-06/HID (‘Direccio´n General de Ensen˜anza Superior e Investigacio´n Cientı´ fica, subdireccio´n General de Proyectos de Investigacio´n Cientı´ fica y Te´cnica’, Ministerio de Educacio´n y Cultura, Madrid, Spain).
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