acta oecologica 34 (2008) 202–213
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Original article
Invasion patterns of ground-dwelling arthropods in Canarian laurel forests Erik Arndta,*, Jo¨rg Pernerb a
Department LOEL, Anhalt University of Applied Sciences, Strenzfelder Allee 28, D-06406 Bernburg, Germany U.A.S. Umwelt- und Agrarstudien GmbH, Ilmstrasse 6, D-07743 Jena, Germany
b
article info
abstract
Article history:
Patterns of invasive species in four different functional groups of ground-dwelling
Received 12 September 2007
arthropods (Carnivorous ground dwelling beetles; Chilopoda; Diplopoda; Oniscoidea)
Accepted 15 May 2008
were examined in laurel forests of the Canary Islands. The following hypotheses were
Published online 30 June 2008
tested: (A) increasing species richness is connected with decreasing invasibility as predicted by the Diversity–invasibility hypothesis (DIH); (B) disturbed or anthropogenically
Keywords:
influenced habitats are more sensitive for invasions than natural and undisturbed habitats;
Canary islands
and (C) climatic differences between laurel forest sites do not affect the rate of invasibility.
Laurel forest
A large proportion of invasives (species and abundances) was observed in most of the
Biological invasions
studied arthropod groups. However, we did not find any support for the DIH based on
Extrinsic factors
the examined arthropod groups. Regarding the impact of the extrinsic factors ‘disturbance’
Carabidae
and ‘climate’ on invasion patterns, we found considerable differences between the studied
Staphylinidae
functional groups. Whereas the ‘disturbance parameters’ played a minor role and only
Oniscoidea
affected the relative abundances of invasive centipedes (positively) and millipedes (nega-
Myriapoda
tively), the ‘climate parameters’ were significantly linked with the pattern of invasive detritivores. Interactions between native and invading species have not been observed thus far, but cannot completely be excluded. ª 2008 Elsevier Masson SAS. All rights reserved.
1.
Introduction
Biological invasions have become a global phenomenon stimulating intense theoretical, experimental and observational research in ecology. One basic theoretical concept predicts that communities with high diversity are highly competitive and less invasible (‘Diversity–invasibility hypothesis’, supported by the resource-competition theory; Elton, 1958; MacArthur and Wilson, 1967; Pimm, 1991; Crawley et al., 1999; Tilman, 1999). However, studies exploring the validity of this hypothesis vary in their conclusions. Whereas the
results of several theoretical and experimental studies conducted at small spatial scales support this hypothesis (e.g. Case, 1990; Luh and Pimm, 1993; Tilman et al., 1996; Tilman, 1997; Hooper and Vitousek, 1998; Naeem et al., 2000; Dukes, 2001), other experimental and observational studies, mostly related to regional scales, did not find any evidence for its validity in nature (e.g. Robinson et al., 1995; Planty-Tabacchi et al., 1996; Palmer and Maurer, 1997; Stohlgren et al., 1999). This contradiction was attributed to the impact of extrinsic factors like disturbance, climate, or soil conditions, which co-vary with the diversity of native and invasive species and
* Corresponding author. Fax: þ49 3471 3559 1110. E-mail addresses:
[email protected] (E. Arndt),
[email protected] (J. Perner). 1146-609X/$ – see front matter ª 2008 Elsevier Masson SAS. All rights reserved. doi:10.1016/j.actao.2008.05.005
acta oecologica 34 (2008) 202–213
might mask effects of diversity (Levine and D’Antonio, 1999; Naeem et al., 2000; Kennedy et al., 2002). The basic idea of Elton (1958) ignores the possible effect of such extrinsic factors and predicts a negative association between the (native) species diversity and the success of invaders. However, most of these results concern plant communities, often in experimental studies. In this study we analysed data of ground-dwelling arthropods sampled in laurel forests of the Canary Islands. These forests are one of the species richest ecosystems of the Canarian Archipelago and only occur in the northern parts of the mountainous western islands as well as on Madeira and the Azores (Del-Arco et al., 1999; Walter and Breckle, 1999). Laurel forests are characterized by 20 mainly endemic tree species and a forb-rich flora. Currently non-native species are negligible at the floristic level. During a preliminary study of soil macro-invertebrates on the Canary Islands we recorded however an extremely high percentage of invasives in some of the laurel forest sites. Because ecological studies of invasive terrestrial invertebrates and in particular investigations based on ground-dwelling taxa are very rare thus far, the analysis of the soil macro-fauna sampled in laurel forests seemed to be an interesting approach to test the applicability of the theoretical framework of species invasions. We used two carnivorous (Carabidae/Staphylinidae, Chilopoda) and two detritivorous arthropod groups (Diplopoda, Isopoda) to test the following hypotheses: (A) increasing species richness is connected with decreasing invasibility; (B) disturbed or anthropogenically influenced habitats are more sensitive for invasions than natural and undisturbed habitats; and (C) climatic differences between laurel forest sites do not affect the rate of invasibility. Hypothesis A is based on the Diversity–invasibility hypothesis (Elton, 1958; Tilman, 1999) assuming that a low number of native species means ‘empty niches’ in the sense of food web models which may support an increasing number of invaders. The conceivable link between the degree of disturbance and the observed invasion success (hypothesis B) was already mentioned by Vitousek (1990) and Williamson (1996). We were interested to test if this hypothesis meets all invertebrate groups in the same way. Furthermore Williamson (1996: 71) pointed out that most invaders spread in new habitats with climatic conditions similar to their original area, but that ‘‘.there are plenty exceptions to climatic matching. All in all, climatic matching seems a fine example of a factor that ought to be of overriding importance and yet is on the whole a rather weak indicator or predictor [of invasions].’’. Therefore, we expected (hypothesis C) that the climatic differences between study sites should not affect the proportion of invasives.
2.
Material and methods
2.1.
Study area and sampling sites
The Canarian Archipelago is the central part of the Macaronesian subregion in the Mediterranean biogeographical region. Laurel forests occur there as cloud forests on the northern parts of the five mountainous western Canary Islands at elevations between 600 and 1400 m. Because of influence by
203
trade winds, these forests are characterized by comparatively cool and moist conditions. Four endemic tree species of the Lauraceae, together with 16 further trees and shrubs (e.g. Myrica faya, Erica arborea, and Viburnum rigidum) are the dominant plants. The study was confined to the three westernmost islands of the archipelago (Fig. 1) though laurel forests also occur on Gran Canaria and Tenerife. However, Gran Canaria was largely deforested in the past (more than 99% of laurel forest are disappeared) and therefore an examination of native laurel stands was impossible there. Also, Tenerife has lost about 90% of its native laurel forests; the remaining areas are isolated and of different historical origin (Ferna´ndez Lopez, 2001). Depending on the different extent of recent laurel forests 11 sites were selected on La Gomera, five on La Palma and four on El Hierro (Table 1). All sites were located in the socalled ‘‘bioclimatophilous region of laurel forest’’ (Del-Arco et al., 1999), indicating that the central laurel zone is characterized by comparable soil conditions and potential vegetation. The sites selected include stands with natural forest structure (not or hardly influenced by human activity), as well as sites influenced by forestry, forest fire or other human activities. The largest area of Canarian laurel forest remained on La Gomera, still representing a continuous ecosystem. The selected 11 study sites on La Gomera (last letter of site code ¼ G; see Table 1) mainly represent permanent monitoring sites of the Garajonay National Park. We received information on precipitation, temperature, and historical use of the sites from the National Park. The laurel forest of El Hierro is also represented by a continuous forest area, but it is much smaller than that on La Gomera and mainly situated on the steep north slope of the mountain ridge. Four sites were selected on El Hierro (last letter of site code ¼ H; see Table 1). The forest rudiments of La Palma are strongly fragmented (Delgado et al., 2001a). Five sites in three different forest rudiments were selected on La Palma (last letter of site code ¼ P, see Table 1). The three forest rudiments represent natural, slightly influenced and strongly influenced forest plots. All examinations were carried out in the forest interior to minimize edge effects or influences from roads or differing forest structure in the neighbourhood. The minimum distance to a forest edge was 50 m (E1P, M3H, E1H) and to a forest pad 20 m (M1H, M2H), in most cases the distances were about 100 m. Site effects and reciprocal influences of study sites may be important in field studies. With exception of the neighbouring L1G/E4G near El Cedro on La Gomera, and M3H/E1H near Tigaday on El Hierro all sites under study were more than 300 m apart from each other. The litter layer, and coverage of tree and forb strata were determined on each site. The depth of the litter layer is an average value of the study site in the environment of the pitfall traps/soil sample points (Table 1). It is widely accepted that a low cover of Laurus as well as a high cover of Erica and Myrica indicate degradation or anthropogenic disturbances in laurel forests (Pe´rez de Paz et al., 1990; Hohenester and Welss, 1993; Del-Arco et al., 1999: 285). Therefore we used the percentage cover of those trees as surrogates for the degree of disturbance
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acta oecologica 34 (2008) 202–213
Fig. 1 – Map of the Canary Archipelago with extent of laurel forests as well as percentage of original forest area according to Ferna´ndez Lopez (2001).
(Table 1; following called ‘disturbance parameters’) and tested if disturbance level affects the proportion of invasives (hypothesis B). The age of trees was also tested as indicator of ‘disturbance’, because natural forests represent sites with old trees dominated by Laurus and Persea (first letter of site code ¼ L; see Table 1). Secondary forests, where parts of the laurel trees were removed (‘fayal-brezal’ in Canarian terminology) are often dominated by Myrica (first letter of site code ¼ M; see Table 1). Sites dominated by Erica (first letter of site code ¼ E; see Table 1) and partly covered with Ilex describe young tertiary forests (15–40 years old) with pioneer vegetation. Therefore, Erica and Myrica dominated forest sites could be regarded as succession stages after removing the natural laurel forest. However, native Myrica-Erica stands also exist on La Gomera in the upper, cooler and drier region of the laurel zone on the mountain ridges (higher than 1300 m). These native Myrica-Erica forest sites are very rare and the only one examined here is site M1G (see Table 1). To test for climatic effects on the proportion of invasives we used the potential direct solar insolation (PDSI), altitude of sites and cover of tree and forb layer as well as the extent of litter layer; these are considered to be important parameters affecting the above-ground climate (‘climate parameters’). Mean daily PDSI values were calculated using mean aspect and inclination of sites (Homann, Schumacher and Perner, unpublished software program based on an algorithm by Volz, 1959).
2.2.
Data collection
Pit fall traps were used to collect surface active macroinvertebrate taxa. Five traps (plastic cups, 65 mm diameter,
containing a 5% acetic acid/salt mixture) were placed in one line about 3–4 m distant from each other. The traps were open for 6 weeks during March–August 2003 (3 weeks each in spring and in summer) and were checked weekly. The endogeic fauna was examined using soil samples. For this purpose, eight soil samples (25 25 15 cm; excluding the litter layer) were taken per site (five in March, three in August, same period as pitfall traps). The soil samples were brought into the laboratory and were examined for invertebrates 3 mm using shovel, tweezers and a stereo microscope. All specimens were transferred to 70% EOH, counted and identified on species level. The analyses are based on the following functional groups comprising two carnivorous and two detritivorous groups: Non-specialized ground dwelling predatory beetles (Carabidae and Staphylinidae; C1), Chilopoda (C2), Diplopoda (D1), and Isopoda (D2). Results of pitfall traps (PT) could be linked to all groups; data of soil samples (SS) allowed conclusions on Chilopoda (C2), Diplopoda (D1), and Isopoda (D2).
2.3.
Data analysis
The species were classified as ‘natives’ or ‘invasives’ according to the following sources: Carabidae: Machado (1992), Izquierdo et al. (2004); Staphylinidae: Schu¨lke, unpublished data (pers. commun.); Chilopoda: Zapparoli, unpublished checklist (pers. commun.); Diplopoda: Vicente and Enghoff (1999), Arndt et al. (in press); Isopoda: Arndt and Mattern (2005), Izquierdo et al. (2004). To test for (negative) correlations between the number as well as the log-transformed abundances of native and invasive species, simple Pearson’s correlations were used.
Table 1 – Vegetation and site characteristics of the 20 investigated laurel forest stands. Parameters used for further analysis are bold-faced Site
L2G
L3G
L4G
L5G
L6G
M1G
E1G
E2G
E3G
E4G
M1H
M2H
M3H
E1H
L1P
L2P
M1P
M2P
E1P
90
75
76
80
85
90
80
60
90
75
60
60
85
85
50
80
95
85
85
80
70
35
65
77
70
46
0
18
0
5
0
0
0
0
0
40
95
60
50
20
0 0 5 0 5 10 0 0 35
0 34 0 0 0 6 0 0 50
0 5 3 0 3 0 0 0 50
0 0 1 1 1 0 0 0 10
0 4 0 0 0 7 4 0 5
0 0 4 12 3 25 0 0 5
0 0 0 50 30 0 0 0 75
0 0 0 36 6 0 0 0 50
0 0 0 80 10 0 0 0 80
0 0 10 55 5 0 0 0 2
0 0 5 40 10 5 0 0 5
0 0 0 6 54 0 0 0 90
0 0 0 17 68 0 0 0 85
0 0 0 8.5 76.5 0 0 0 15
0 0 0 40 10 0 0 0 30
40 0 0 0 0 0 0 0 10
0 0 0 0 0 0 0 0 10
0 0 0 0 17 0 0 8 5
0 0 0 17 18 0 0 0 30
0 0 12 24 24 0 0 0 10
8 800 100 NE 30 288090 476
8 950 100 N 7 288090 607
8 950 100 N 15 288090 550
8 1300 100 N 30 288090 417
2 1000 100 S 15 288090 702
1 1350 100 N 35 288090 369
0 1200 50 S 10 288090 690
1 1000 16 0 0 288090 648
1 800 30 E 35 288090 572
3 1300 100 0 0 278430 652
7.5 860 15 N 20 278430 514
0 860 15 N 20 278430 514
1 600 100 W 30 288430 587
10 700 100 W 40 288430 551
3 1200 60 E 40 288430 551
3 1200 60 E 30 288430 587
3 1000 30 W 10 288430 637
1.5 1000 100 SW 15 288090 682
0.5 1000 40 NE 25 288090 510
1.5 1200 30 NE 10 278430 607
acta oecologica 34 (2008) 202–213
Tree layer in total (%) Laurus novocanariensis Ocotea foetens Persea indica Ilex canariensis Erica arborea Myrica faya Viburnum rigidum Rhamnus glandulosa Pinus canariensis Forb layer in total (%) Litter layer (cm) Altitude (m a.s.l.) Age Exposition Inclination (8) Latitude PDSI (cal/cm2/day)
L1G
205
206
acta oecologica 34 (2008) 202–213
The proportion of invasives (invasives/(invasives þ natives)) within each functional group was modelled using the abundances as well as species richness of natives and some selected parameters describing the physical environment (climate and disturbance parameters, see above and Table 1) as independent explanatory variables. Because of data structure (no homogeneity of variances, errors not normally distributed) weighted regression (modelling) models are needed, and therefore generalized linear models (GLMs) were used (Crawley, 2002; Quinn and Keough, 2002; Pysˇek et al., 2003). We used a logit-link function to model the proportional responses of invasive and native species within each functional group, which is appropriate for such data. In cases of overdispersion, the variance function was modelled proportional to mu (1 mu) (mu ¼ mean of the distribution; using the quasi option in S-PLUS; see Crawley, 2002). The following stepwise procedure was used in all analyses to derive the most parsimonious model that adequately describes the observed data (see suggestions of Crawley, 2002: 449). In a first step a fully saturated model was fitted including all three climate parameters or all three disturbance parameters each. The contribution of each parameter was examined in a second step by sequential parameter exclusion and re-fitting the model. The parameter set whose exclusion induced the highest and significant different increase in deviance compared with others was finally assumed as the minimal adequate model. The significance of explanation change in model deviance was tested using F-test.
3.
separately fitted to the parameters describing the disturbance status of sites (‘disturbance parameters’) and the climate affecting parameters (‘climate parameters’, see Section 2). The detailed model fits are presented in Table 2 and the significant results of the minimal adequate model for the abundance proportions of invasives are illustrated in Fig. 2. A significant increasing proportion of invasives with increasing disturbance level indicated by decreasing cover of Laurus and/or increasing cover of Erica or Myrica (see Section 2) could be found for the relative abundance of invasive centipedes only (C2, Fig. 2d). In contrast, the abundance proportion of invasive millipedes (D1, soil sample, Fig. 2e) was significant positively related to Laurus-cover and the relative occurrence of invasive isopod species (D2, pitfall traps) was significant negatively related to Erica-cover both corresponding with a decreasing disturbance level. ‘Climate parameters’ are significantly linked with the pattern of invasive detritivores as stepwise decreasing deviances from the minimal to the full model demonstrate (Table 2). With the exception of the relative species numbers of invasive isopods (D2, pitfall traps) the proportion of invasives (abundances and species numbers) increases significantly with decreasing altitude or/and increasing insolation (PDSI) for both detritivorous groups (Fig. 2a–c). The relative occurrence of invasive carnivores (C1, C2) does not show significant relationships to a certain tested ‘climate parameter’. Only the relative species number of invasive centipedes (C2) was significantly related to the cover of tree layer. No effects of forest age, litter layer, and herb stratum were found at all.
Results
In both pitfall traps and soil samples 6552 individuals (2911 natives, 3641 invasives) of 67 arthropod species (46 natives, 21 invasives) from the examined four functional groups were collected. Abundances and species numbers decreased in general from spring to summer catches in all taxa or functional groups. However, in total species number of Carabidae (adults) did not decrease in summer, but the species composition changed partly. Several Calathus species disappeared, whereas Laemostenus and Cymindis species occurred in the summer pit fall catches. Whereas nearly all species of ground and rove beetles were natives, we found a much higher proportion of invasive species in the other groups. In particular, conspicuously high abundances of invasive species were detected for isopods and millipedes (Appendices A and B). Laemostenus complanatus was the only invasive species with distinctly higher abundance in summer than in spring. We used a simple Pearson’s correlation analysis to test for a (negative) correlation between number or (log-transformed) abundances of natives and invasive species. However, no significant correlations were found in any of the studied groups except a positive (!) correlation between the number of native and invasive species of Chilopoda (r ¼ 0.467, P < 0.05) in soil samples. To search for parameters which affect significantly the proportion of invasive arthropods we used generalized linear model procedures (GLM). The proportions of invasives (abundances and species numbers) of all functional groups were
4.
Discussion
4.1. Effects of native diversity and extrinsic factors on the proportion of invasives Many authors found a negative correlation between species diversity and invasibility based on studies of plant communities (Pimm, 1991; Tilman et al., 1996; Tilman, 1997; Hooper and Vitousek, 1998; Dukes, 2001; see also reviews by Levine and D’Antonio, 1999; Prieur-Richard and Lavorel, 2000). This relation, known as Diversity–invasibility hypothesis (DIH, e.g. Elton, 1958; Tilman, 1999), was tested on a ground-dwelling animal community with different functional groups (hypothesis A). We did not find any support for this hypothesis based on the examined arthropod groups. A significant correlation between species numbers of native and invasive species only occurs in Chilopoda (C2). However, this is a positive correlation (r ¼ 0.467, P < 0.05) and therefore would contradict the DIH. Two reasons may be responsible for the failure of the DIH in our study. (i) The species numbers in the different ecological groups might be too low and this relation might be covered by stochastic effects. The observed species numbers per site and group vary between 1 and 9, with 4 on average. They are much lower than those of the plant communities examined in studies mentioned above. Sites with high species numbers (E1P, L1P, L2P) include more invasive than native species. (ii) The second reason could be a stronger influence of extrinsic factors which
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acta oecologica 34 (2008) 202–213
Table 2 – Results of fitting ‘climate parameters’ and ‘disturbance parameters’ to the proportion of invasive arthropods (abundances and species numbers) using GLM. The change in deviance (Ddev.) was tested using F statistics and significant (P < 0.05) minimal models are bold-faced. (D) and (L) behind the significant terms included in model indicate the sign of function coefficients Climate effects Term included in model
Ddev.
Disturbance effects F
P
Term included in model
Ddev.
F
P
Carabidae and Staphylinidae (C1), pitfall trapping, abundancesddeviance of null model ¼ 133.32 Altitude 117.93 0.17 0.6886 Erica Altitude þ Tree.layer 117.15 0.01 0.9282 Erica þ Laurus Altitude þ Tree.layer þ PDSI 116.03 0.01 0.9137 Erica þ Laurus þ Myrica
84.31 73.41 63.37
0.00 0.00 0.00
0.9839 0.9924 0.9927
Carabidae and Staphylinidae (C1), pitfall trapping, species numbersddeviance of null model ¼ 18.73 PDSI 18.48 0.10 0.7514 Erica PDSI þ Tree.layer 18.45 0.01 0.9147 Erica þ Laurus PDSI þ Tree.layer þ Altitude 18.44 0.01 0.9420 Erica þ Laurus þ Myrica
15.82 15.30 15.29
1.86 0.33 0.01
0.1912 0.5730 0.9356
Chilopoda (C2), soil samples, abundancesddeviance of null model ¼ 96.60 Tree.layer 92.10 0.76 0.3947 Tree.layer þ Altitude 91.67 0.07 0.7913 Tree.layer þ Altitude þ PDSI 91.60 0.01 0.9125
Laurus (L)* Laurus þ Erica Laurus þ Erica þ Myrica
70.93 65.40 65.40
6.48 1.39 0.00
0.0216 0.2550 0.9893
Chilopoda (C2), soil samples, species numbersddeviance of null model ¼ 5.69 Tree.layer (L)* 4.50 4.94 0.0409 Laurus Tree.layer þ PDSI 4.08 1.76 0.2038 Laurus þ Erica Tree.layer þ PDSI þ Altitude 3.87 0.90 0.3571 Laurus þ Erica þ Myrica
5.23 4.56 4.38
1.70 2.48 0.68
0.2109 0.1350 0.4217
Diplopoda (D1), soil samples, abundancesddeviance of null model ¼ 662.94 Altitude (L)** 358.73 12.71 0.0026 Laurus (D)* Altitude þ Tree.layer 253.43 4.40 0.0522 Laurus þ Erica 252.97 0.02 0.8917 Laurus þ Erica þ Myrica Altitude þ Tree.layer þ PDSI
501.87 486.27 481.37
5.31 0.51 0.16
0.0349 0.4836 0.6928
Diplopoda (D1), soil samples, species numbersddeviance of null model ¼ 16.06 Altitude (L)* 12.88 5.96 0.0266 Erica Altitude þ Tree.layer 11.74 2.14 0.1625 Erica þ Laurus Altitude þ Tree.layer þ PDSI 10.98 1.43 0.2485 Erica þ Laurus þ Myrica
13.61 13.40 13.23
3.62 0.30 0.25
0.0752 0.5901 0.6207
Laurus Laurus þ Erica Laurus þ Erica þ Myrica
687.31 660.64 655.49
1.36 0.53 0.10
0.2687 0.4820 0.7551
Isopoda (D2), soil samples, species numbersddeviance of null model ¼ 15.46 Altitude (L)* 10.62 7.44 0.0197 Myrica Altitude þ PDSI 10.59 0.04 0.8378 Myrica þ Erica Altitude þ PDSI þ Tree.layer 9.46 1.74 0.2139 Myrica þ Erica þ Laurus
13.44 12.97 11.89
2.50 0.58 1.34
0.1424 0.4617 0.2712
Isopoda (D2), pitfall trapping, abundancesddeviance of null model ¼ 2250.63 PDSI (D)* 1189.12 8.10 0.0129 Erica PDSI þ Tree.layer 894.62 2.25 0.1560 Erica þ Laurus 822.18 0.55 0.4694 Erica þ Laurus þ Myrica PDSI þ Tree.layer þ Altitude
2170.77 2156.36 2008.19
0.50 0.09 0.93
0.4903 0.7678 0.3510
Isopoda (D2), pitfall trapping, species numbersddeviance of null model ¼ 20.03 Altitude 17.39 3.02 0.1043 Erica (L)* Altitude þ Tree.layer 16.48 1.03 0.3263 Erica þ Laurus Altitude þ Tree.layer þ PDSI 16.42 0.07 0.7980 Erica þ Laurus þ Myrica
15.55 15.34 14.85
5.47 0.25 0.60
0.0348 0.6267 0.4522
Isopoda (D2), soil samples, abundancesddeviance of null model ¼ 755.67 Altitude (L)*** 128.31 92.55 < 0.0000 Altitude (L)D PDSI (D)* 74.07 8.00 0.0164 Altitude þ PDSI þ Tree.layer 65.05 1.33 0.2732
*P<0.05; **P<0.02; ***P<0.01.
co-vary with the diversity of native or invasive species. A positive relation between the species richness of native (endemic) and invasive species was already found by Borges et al. (2006) on arthropods of the Azorean islands. However, functional groups were not evaluated separately in this study. Borges et al. (2006) concluded that endemic carnivorous arthropod
species could be positively influenced by increasing abundances of introduced species (representing several functional groups) because of larger food resources. The DIH (hypothesis A in Section 1) would indicate the result of interspecific interactions like competition which play an important role in plant communities. We do not
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acta oecologica 34 (2008) 202–213
Fig. 2 – Relation between abundance proportions of invasives and significant ‘climate’ and ‘disturbance’ parameters included in minimal adequate model using GLM. Real proportion data are included as circles. The different bubble-sizes of circles illustrate the different weight of samples.
know if such interactions play the same role in the examined animal communities. Dangerfield (1989) and Hassall and Dangerfield (1989) described competition between Armadillidium vulgare and a Porcellio species. Competition was also repeatedly suggested for carabid communities (e.g. Loreau, 1989, 1990, 1994; Niemela¨ and Spence, 1991). In the present study invasive carabid and staphylinid beetles (functional group C1, see Appendices A and B) only occur in sites with few species (La Palma, El Hierro), but never in laurel sites with high diversity (La Gomera). They occupy vacant niches (Arndt, 2006) which would support the DIH. However, species interactions in some of the small and fragmented laurel study sites may also be superposed by other factors like frequent local extinctions, microclimate, and especially neighbourhood relations. We believe that the number of species observed per
site is too low to draw trustworthy conclusions from the present results. The second hypothesis (B) describes the link between the increasing disturbance level and the number of invasive species. There is much support for a positive correlation between disturbance and invasibility mainly based on plant community studies (Vitousek, 1990; Williamson, 1996; Kowarik, 2003). However, studies using invertebrate functional groups analysing this relationship are rare (Borges et al., 2006; Delgado et al., 2001b). We used vegetation parameters of the study site (Table 1) reflecting influence of forestry as characters of disturbance. A forest site with natural (laurel) vegetation and old trees tends to be ‘undisturbed’; in contrast secondary or tertiary sites with young trees reflect disturbed forests. The various examined arthropod taxa respond in
acta oecologica 34 (2008) 202–213
different ways to ‘disturbance’. Centipedes (C2) show a significant increasing proportion of invasives with increasing disturbance level (Fig. 2d). In contrast, invasive millipedes (D1) and isopods (D2, pitfall traps) are negatively related to disturbance factors. Isopods in soil samples and the carnivorous ground and rove beetles (C1) do not respond significantly. In fact, these results fall into the non-uniform findings in studies of plant communities and illustrate that also some other parameters than anthropogenic activities seem to affect the success of invasive species. Furthermore, these non-uniform findings may also underpin the different responses of different trophic levels concerning the partly linked relations between climate, disturbance and complexity (Voigt et al., 2007). Hypothesis (C), derived from the literature (Williamson, 1996), denies a strong correlation between occurrence of invasive animals and local climatic factors. This hypothesis can be confirmed for both carnivorous groups. However, it must be rejected for the detritivorous taxa. There is a significant correlation between both detritivorous groups and climatic factors. Abundances and species numbers of millipedes (D1) as well as abundances of isopods (D2) increase significantly with decreasing altitude or/and increasing insolation (PDSI). At least these results give some evidence that the invasion pattern of some taxa or functional groups may be also significantly controlled by climatic factors. The tree layer is the second important factor determining distribution pattern of millipedes (D1) and isopods (D2) (Table 2) in the examined habitat type. The tree layer itself is not significant but reduces the deviance of species and abundances in millipedes strongly. Isopods show a similar picture (species number, pitfall traps). A dense tree layer with Laurus novocanariensis and other natural broad-leaved trees can be interpreted as suitable trophic base for detritovorous groups. Therefore trophic and climatic factors seem to be more important for invasion or distribution of millipedes and isopods than parameters of disturbance. Borges et al. (2006) obtained similar results in Azorean arthropod assemblages, where disturbance parameters, habitat characters, but also climatic parameters significantly influenced the number of introduced species. In this Azorean study the group of introduced species is mainly represented by Araneae, carnivorous Coleoptera, Homoptera and Lepidoptera rather than by numerous detritovorous taxa.
4.2. Invasive ground-dwelling arthropods in an island forests ecosystem Small islands seem to be more fragile and prone to species invasions than continental areas because of the characteristics island species have evolved and the gregariousness of invasive species (Dulloo et al., 2002). Many island studies give evidence that the ecology of such former hotspots of biodiversity has been affected by invasive alien species which caused a large-scale degradation and impoverishment of the indigenous flora and fauna (e.g. Chown et al., 2002; Dulloo et al., 2002; Jones et al., 2002; Kelly and Samways, 2003). The Canary Islands are one such hotspot of biodiversity. Despite the small total area of 7447 km2 the seven main islands harbour 1147 families, 4520 genera, and more
209
than 12,680 species of terrestrial plants and animals (Izquierdo et al., 2004 supplemented by additional unpublished data). The percentage of endemic species is about 38% in insects, 70% in millipedes, 60% in terrestrial Malacostraca, 78% in molluscs, and 26% in vascular plants (Izquierdo et al., 2004). However, the biodiversity of this archipelago is endangered by devastation of the naturally small ecosystems through anthropogenic influences like settlements (especially tourism infrastructure), agricultural plantations and clear-cutting of forests as well as by invasions of alien species. The proportions of native and invasive ground-dwelling arthropods in laurel forests of the Canary Islands were studied the first time. Whereas up to now invasive species are nearly negligible at the producer level, at the consumer level a large proportion of aliens (species and abundances) was observed in most of the studied arthropod groups. Several invasives were found for first time either on a particular island or in a particular site. Although previously largely ignored, recently some studies have highlighted the potential impacts of invasive soil invertebrates on ecosystem structure and function. Aliens can change, for example, soil carbon, nitrogen and phosphorus pools and can considerably affect the distribution and function of roots and microbes (Kelly and Samways, 2003; Bohlen et al., 2004; Suarez et al., 2004). Crooks (2002) emphasized that those invaders will have the largest impacts on ecosystems which directly modify ecosystems and thus have cascading effects for resident biota. Furthermore he argued that such invasive ecosystem engineers can facilitate further invasions by directly changing habitat characteristics (extrinsic factors). From these points of view the locally very high dominance of some invading decomposers like Armadillidium vulgare, Cylindroiulus disjunctus or Ommatioulus moreletii, originally distributed in the Mediterranean region, seems to be critical. Therefore, further studies in the Canarian laurel forests should focus in particular on impacts of management and continued fragmentation of laurel forests on the diversity of natives and the changing extrinsic factors promoting the invasion of aliens.
Acknowledgements ´ ngel Ferna´ndez and The field work was supported by A co-workers (Garajonay National Park, La Gomera), Jose´-Marı´a Fernandez Palacios (University of La Laguna, Tenerife), and Antonio Machado (La Laguna, Tenerife) who are gratefully acknowledged. Birte Wisser, Michael Klemm, Pierre Angelo Cocco, and Stephan Fiedler (Anhalt University, Bernburg, Germany) carried out most of the field work spending much of their time in the project. Henrik Enghoff (Kopenhagn, Denmark), Norman Lindner (Leipzig, Germany), Dirk Mattern (Erfurt, Germany), Michael Schu¨lke (Berlin, Germany), and Marzio Zapparoli (Viterbo, Italy) kindly determined material or improved questionable specimens of Staphylinidae, Isopoda, Chilopoda and Diplopoda. We are indebted especially to Marzio Zapparoli and Michael Schu¨lke who provided us unpublished lists of introduced Chilopoda and Staphylinidae.
210
Appendix A Arthropods sampled with pitfall traps in the 20 studied laurel forest sites. Data of 5 traps per site are summarized. FG ¼ functional group, for code used see Table 2. Status indicates native (n) and invasive (i) species
FG
Status
L1G
L2G
L3G
Broscus crassimargo Wollaston Calathus gomerensis Colas Calathus laureticola Wollaston Calathus marcellae Colas Calathus pilosipennis Machado Calathus spretus Wollaston
C1 C1 C1 C1 C1 C1
n n n n n n
1 0 0 0 0 0
2 0 0 0 0 0
0 0 0 0 0 0
Cryptophonus schaumii (Wollaston) Cymindis simillima (Wollaston) Cymindis velata (Wollaston) Dicrodontus aptinoides (Wollaston) Gabrius canariensis (Fauvel) Gomerina calathiformis (Wollaston) Laemostenus complanatus (Dejean)
C1 C1 C1 C1 C1 C1 C1
n n n n n n i
0 2 0 1 0 0 0
0 2 0 0 0 0 0
Licinopsis angustula Machado Ocypus affinis (Wollaston) Ocypus olens (Mu¨ller) Ocypus subaenescens Wollaston Ocypus sylvaticus Wollaston Othius punctulatus (Goeze)
C1 C1 C1 C1 C1 C1
n n i n n n
0 0 0 0 1 0
Paraeutrichops pecoudi Mateu Quedius expectatus Israelson Quedius megalops Wollaston Trechus flavocinctus Geophilus carpophagus Leach Henia bicarinata (Meinert)
C1 C1 C1 C1 C2 C2
n n n n i I
Lithobius crassipes L. Koch Lithobius obscurus Meinert Lithobius pilicornis Newport Lithobius sp. nov. 1 Lithobius tenerifae Latzel Acipes franzi (Loksa) Brachydesmus proximus Latzel
C2 C2 C2 C2 C2 D1 D1
Brachydesmus superus Latzel Cylindroiulus disjunctus Read Glomeris canariensis Golovatch Glomeris gomerana Attems Glomeris hierroensis Enghoff and Golovatch Hirudicrypticus canariensis (Loksa)
L4G
E3G
E4G
L6G
E2G
M1G
E1G
L5G
M1H
M2H
M3H
E1H
0 0 1 0 0 0
0 0 0 0 0 0
0 0 0 1 0 0
0 0 0 0 0 0
1 5 0 0 5 0
0 13 1 0 5 0
11 6 0 2 2 0
5 1 0 0 0 0
0 0 0 0 0 18
0 0 0 0 0 29
0 0 0 0 0 22
0 0 0 0 0 14
0 2 0 0 0 0 0
0 2 21 0 0 0 0
0 5 42 0 0 0 0
0 1 13 0 0 0 0
0 9 91 1 0 0 0
0 17 37 0 0 2 0
0 1 17 0 0 0 0
0 3 89 0 0 0 0
0 3 32 0 0 0 0
0 0 0 0 2 0 0
0 0 0 0 8 0 0
0 0 0 0 4 0 0
0 0 0 0 1 0
0 0 0 0 0 0
0 0 0 0 15 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 2 0
0 0 0 0 9 0
0 0 0 0 35 0
0 0 0 0 21 0
0 0 4 1 0 0
0 0 0 19 0 0
1 0 0 0 0 0
2 0 0 1 1 0
3 0 0 0 0 0
0 3 0 0 0 0
0 0 1 0 0 0
0 0 0 0 1 0
0 0 1 1 1 0
1 0 1 0 0 0
0 0 0 1 0 0
66 0 0 0 0 0
18 0 1 0 0 0
0 0 3 1 0 0
i i n n n n i
0 0 0 0 0 0 4
0 0 0 0 0 0 3
0 0 0 0 0 0 2
0 0 0 0 0 0 0
0 0 0 0 0 0 0
0 0 0 2 1 0 0
0 0 0 0 0 0 0
0 0 1 0 0 0 0
0 0 0 1 0 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 0
D1 D1 D1 D1 D1 D1
i n n n n n
0 0 2 0 0 0
0 0 0 3 0 2
0 0 0 2 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
1 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
7 0 0 0 0 0
Nopoiulus kochii (Gervais) Ommatoiulus moreleti (Lucas) Polydesmus coriaceus Porat Agabiformis lentus Budde-Lund Armadillidium vulgare (Latreille) Ctenoscia minima (Dollfus) Eluma purpurascens Budde-Lund
D1 D1 D1 D2 D2 D2 D2
i i i i i n i
0 2 0 0 0 0 0
0 0 0 0 0 0 1
0 1 0 0 0 0 8
0 2 0 1 0 2 0
0 0 0 0 0 1 0
0 0 0 0 1 0 3
0 0 0 0 24 0 85
0 6 0 0 0 0 0
0 1 0 0 0 0 0
Porcellio meridionalis Vandel Porcellio septentrionalis Vandel Porcellio sp. nov. Soteriscus stricticauda (Dollfus) Venezillio ausseli (Dollfus)
D2 D2 D2 D2 D2
n n n n n
0 0 0 0 0
1 0 0 0 0
0 0 0 0 0
170 0 0 0 1
0 0 0 0 0
1 0 0 1 0
5 0 0 0 8
0 0 1 0 8
195 0 0 0 10
E1P
L1P
L2P
M1P
M2P
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 1 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 1
0 0 0 0 0 0 0
0 0 0 0 0 0 16
1 0 0 0 0 0 1
0 0 0 3 0 0
0 0 0 9 0 0
0 1 0 0 0 0
0 0 0 0 0 0
0 4 0 0 0 0
2 2 0 0 0 9
0 8 0 0 0 0
0 0 1 0 0 0
0 0 1 0 3 1
0 0 2 0 2 1
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 13 0
18 0 0 0 0 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 0
0 0 0 0 0 0 0
0 2 0 0 0 0 0
0 1 0 0 0 0 0
10 0 0 0 0 0
10 1 0 0 3 0
21 1 0 0 0 0
23 0 0 0 0 0
7 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
6 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0
0 0 0 0 0 0 3
0 0 0 0 0 0 0
0 6 0 0 0 0 0
0 21 0 0 0 0 54
0 5 0 0 0 0 1
0 8 0 0 0 0 0
0 10 24 0 855 0 0
0 1 8 0 33 0 3
0 2 8 0 56 0 39
0 1 4 0 166 0 0
2 5 0 0 347 0 0
1 0 0 0 0
0 0 0 0 0
67 0 0 0 0
78 0 0 0 0
23 0 0 0 0
8 0 0 0 1
0 0 0 0 0
0 0 0 0 0
0 7 0 0 0
0 0 0 0 0
0 0 0 0 0
acta oecologica 34 (2008) 202–213
Species
Appendix B Arthropods collected with soil samples in the 20 studied laurel forests. Data of 8 samples per site are summarized. FG ¼ functional group, for code used see Table 2. Status indicates native (n) and invasive (i) species FG
Status
L1G
L2G
L3G
L4G
E3G
E4G
L6G
E2G
M1G
E1G
L5G
M1H
M2H
M3H
E1H
E1P
L1P
L2P
M1P
M2P
Cryptops hortensis Leach Dignathon microcephalus Lucas Geophilus carpophagus Leach Geophilus insculptus Attems Henia bicarinata (Meinert) Lithobius crassipes L. Koch Lithobius gomerae Eason Lithobius obscurus Meinert Lithobius pilicornis Newport Lithobius sp. nov. 1 Lithobius sp. nov. 2 Lithobius tenerifae Nannophilus eximius (Meinert) Pachymerium ferrugineum (Koch) Acipes franzi (Loksa) Blaniulus guttulatus (Fabricius) Brachydesmus proximus Latzel Brachydesmus superus Latzel Choneiulus palmatus (Nemec) Cylindroiulus disjunctus Read Dolichoiulus rectangulus Enghoff Dolichoiulus senilis (Attems) Dolichoiulus silvapalma Enghoff Dolichoiulus typhlops (Ceuca) Fuhrmannodesmidae gen. sp. Glomeris gomerana Attems Glomeris hierroensis Enghoff and Golovatch Nopoiulus kochii (Gervais) Ommatoiulus moreleti Lucas Polydesmus coriaceus Porat Armadillidium vulgare (Latreille) Ctenoscia minima (Dollfus) Eluma purpurascens Budde-Lund Haplopththalmus sp. Porcellio meridionalis Vandel Porcellio septentrionalis Vandel Trichoniscus sp.
C2 C2 C2 C2 C2 C2 C2 C2 C2 C2 C2 C2 C2 C2 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D1 D2 D2 D2 D2 D2 D2 D2
n n i n i i n i i n n n i i n i i i n n n n n i i n n i i i i n i i n n n
3 1 2 28 0 0 0 0 0 0 0 1 6 0 0 17 0 13 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0
9 0 2 13 3 2 0 0 0 0 0 0 3 0 1 0 0 31 0 0 0 1 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0
16 0 0 12 1 0 0 0 0 1 0 1 4 0 0 0 0 34 0 0 0 3 0 0 0 0 0 0 3 0 0 0 6 0 0 0 0
3 0 0 42 1 4 16 0 0 0 0 0 2 2 22 0 0 3 0 0 0 1 0 0 0 0 0 0 1 0 0 3 0 0 80 0 0
0 0 0 41 1 0 0 0 0 0 0 0 2 0 1 0 0 36 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0
0 0 0 15 1 1 2 0 0 0 0 0 2 0 4 0 0 9 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0
5 0 3 21 0 0 0 0 0 0 0 0 3 0 15 0 0 73 0 0 0 0 0 0 0 0 0 0 0 0 0 0 46 0 2 0 0
1 0 0 4 2 1 0 0 0 4 0 0 0 0 0 0 0 0 0 0 3 0 0 0 0 1 0 0 7 0 0 0 0 0 1 0 0
1 2 17 13 2 4 0 0 0 22 1 4 1 0 13 0 0 1 0 0 0 1 0 0 0 0 0 0 2 0 0 0 0 0 167 0 0
0 0 2 16 5 0 0 0 0 0 0 0 0 0 35 0 0 37 0 0 0 0 0 0 0 0 0 0 0 0 0 0 6 0 0 0 0
0 0 5 32 0 0 0 0 0 0 0 0 1 0 6 0 0 115 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0
0 0 17 30 0 0 0 0 0 0 0 0 0 0 11 0 0 1 0 108 0 0 0 0 0 0 2 0 8 0 0 0 0 0 25 0 0
0 0 14 67 0 1 0 0 0 0 0 0 0 0 0 0 0 93 0 91 0 0 0 0 0 0 1 0 4 0 0 0 33 0 10 0 2
0 0 16 12 0 6 0 0 0 0 0 0 0 0 0 0 0 93 0 0 0 0 0 0 1 0 0 0 5 0 0 0 0 0 0 0 1
0 0 13 37 0 2 0 0 0 0 0 0 0 0 0 0 0 21 0 5 0 0 0 0 0 0 0 0 9 0 0 0 0 0 0 0 0
5 0 22 24 1 0 0 0 0 0 0 0 0 0 0 0 0 44 0 3 0 0 0 4 6 0 0 0 35 1 129 0 0 0 0 0 17
1 0 6 4 0 0 0 0 0 0 0 0 0 0 0 0 1 3 0 0 0 0 13 0 0 0 0 6 11 8 2 0 8 0 0 0 0
55 0 3 2 0 0 0 0 0 0 0 0 0 0 0 0 48 68 6 0 0 0 7 1 0 0 0 15 15 21 18 0 32 85 0 3 1
8 0 0 39 0 0 0 5 1 0 0 0 0 0 0 0 0 37 0 0 0 0 0 5 4 0 0 0 8 0 14 0 0 0 0 0 55
23 0 32 49 0 0 0 3 0 0 0 0 0 0 0 0 0 60 2 0 0 0 26 9 11 0 0 13 12 0 57 0 0 0 0 0 54
acta oecologica 34 (2008) 202–213
Species
211
212
acta oecologica 34 (2008) 202–213
references
Arndt, E., 2006. Niche occupation by invasive ground-dwelling predator species in Canarian laurel forests (Spain). Biol. Invasions 8, 893–902. Arndt, E., Mattern, D., 2005. Ecological data of isopods (Crustacea: Oniscidea) in laurel forests from the Western Canary Islands. Vieraea 33, 41–50. Arndt, E., Enghoff, H., Spelda, J. Millipedes (Diplopoda) of the Canarian Islands: Checklist and key. Vieraeain in press. Bohlen, P.J., Groffman, P.M., Fahey, T.J., Fisk, M.C., Suarez, E., Pelletier, D.M., Fahey, R.T., 2004. Ecosystem consequences of exotic earthworm invasion of north temperate forests. Ecosystems 7, 1–12. Borges, P.A.V., Lobo, J.M., de Azevedo, E.B., Gaspar, C.S., Melo, C., Nunes, L.V., 2006. Invasibility and species richness of island endemic arthropods: a general model of endemic vs. exotic species. J. Biogeogr 33, 169–187. Case, T.K., 1990. Invasion resistance arises in strongly interacting species-rich model competition communities. Proc. Natl. Acad. Sci.-Biol. 87, 9610–9614. Chown, S.L., McGeoch, M.A., Marshall, D.J., 2002. Diversity and conservation of invertebrates on the sub-Antarctic Prince Edward Islands. Afr. Entomol. 10, 67–82. Crawley, M.J., 2002. Statistical Computing: An Introduction to Data Analysis Using S-Plus. John Wiley and Sons, Chichester. Crawley, M.J., Brown, S.L., Heard, M.S., Edwards, G.R., 1999. Invasion-resistance in experimental grassland communities: species richness or species identity? Ecol. Lett. 2, 140–148. Crooks, J.A., 2002. Characterizing ecosystem-level consequences of biological invasions: the role of ecosystem engineers. Oikos 97, 153–166. Dangerfield, J.M., 1989. Competition and the effects of density on terrestrial isopods. Monitore zoologico italiano (N.S. Monografia 4, 411–423. Del-Arco, M., Acebes, J.-R., Pe´rez-de-Paz, P.-L., Marrero, M.C., 1999. Bioclimatology and climatophilous vegetation of Hierro (part 2) and La Palma (Canary Islands). Phytocoenologia 29, 253–290. Delgado, J.D., Are´valo, J.R., Ferna´ndez-Palacios, J.M., 2001. Fragmentacio´n de los ecosistemas forestales. In: Ferna´ndezPalacios, J.M., Martı´n Esquivel, J.L. (Eds.), Naturaleza de las Islas Canarias. Ecologı´a y Conservacio´n. Publicaciones Turquesa S.L., Tenerife, pp. 173–180. Delgado, J.D., Are´valo, J.R., Ferna´ndez-Palacios, J.M., 2001. Road and topography effects on invasion: edge effects in rat foraging patterns in two oceanic islands forests (Tenerife, Canary Islands). Ecography 24, 539–546. Dukes, J.S., 2001. Biodiversity and invasibility in grassland microcosms. Oecologia 126, 563–568. Dulloo, M.E., Kell, S.P., Jones, C.G., 2002. Conservation of endemic forest species and the threat of invasive species. Impact and control of invasive alien species on small islands. The International Forestry Review 4, 277–285. Elton, C.S., 1958. The Ecology of Invasion by Animals and Plants. Chapman and Hall, London. ´ ., 2001. Conservacio´n y restauracio´n ecolo´gica Ferna´ndez Lo´pez, A de los bosques. In: Ferna´ndez -Palacios, J.M., Martı´n Esquivel, J. (Eds.), Naturaleza de las Islas Canarias. Ecologı´a y Conservacio´n. Publicaciones Turquesa S.L., Tenerife, pp. 375–382. Hassall, M., Dangerfield, J.M., 1989. Inter-specific competition and the relative abundance of grassland isopods. Monit.. zool. ital. (N.S.) Monografia 4, 379–397. Hohenester, A., Welss, W., 1993. Exkursionsflora fu¨r die Kanarischen Inseln. Ulmer, Stuttgart. Hooper, D.U., Vitousek, P.M., 1998. Effects of plant composition and diversity on nutrient cycling. Ecol. Monogr 68, 121–149.
Izquierdo, I., Martı´n, J.L., Zurita, N., Arechavaleta, M. (Eds.), 2004. Lista de Especies Silvestres de Canarias (Hongos, Plantas y Animales Terrestres). Consejerı´a de Medio Ambiente y Ordenacio´n Territorial, Gobierno de Canarias. Jones, A.G., Chown, S.L., Gaston, K.J., 2002. Terrestrial invertebrates of Gough Island: an assemblage under threat? Afr. Entomol 10, 83–91. Levine, J.M., D’Antonio, C.M.D., 1999. Elton revisited: a review of evidence linking diversity and invasibility. Oikos 87, 15–27. Loreau, M., 1989. On testing temporal niche differentiation in carabid beetles. Oecologia 81, 89–96. Loreau, M., 1990. Competition in a carabid beetle community: a field experiment. Oikos 58, 25–38. Loreau, M., 1994. Chapter 5: Ground beetles in a changing environment: determinants of species diversity and community assembly. In: Boyle, T.J.B., Boyle, C.E.B. (Eds.), Biodiversity, Temporate Ecosystems, and Global Change. Springer, Berlin, Heidelberg. Luh, H.-K., Pimm, S.L., 1993. The assembly of ecological communities: a minimalist approach. J. Anim. Ecol 62, 749–765. Kelly, J.A., Samways, M.J., 2003. Diversity and conservation of forest-floor arthropods on a small Seychelles island. Biodivers. Conserv 12, 1793–1813. Kennedy, T.A., Naeem, S., Howe, K.M., Knops, J.M.H., Tilman, D., Reich, P., 2002. Biodiversity as a barrier to ecological invasion. Nature 417, 636–638. Kowarik, I., 2003. Biologische Invasionen: Neophyten und Neozoen in Mitteleuropa. Ulmer, Stuttgart. MacArthur, R.H., Wilson, E.O., 1967. The Theory of Island Biogeography. Princeton University Press, Princeton. Machado, A., 1992. Monografı´a de los cara´bidos de las Islas Canarias. IBCER, La Laguna, Tenerife. 734. Naeem, S., Knops, J.M.H., Tilman, D., Howe, K.M., Kennedy, T., Gale, S., 2000. Plant diversity increases resistance to invasion in the absence of covarying extrinsic factors. Oikos 91, 97–108. Niemela¨, J., Spence, J.R., 1991. Distribution and abundance of an exotic ground-beetles (Carabidae): a test of community impact. Oikos 62, 351–359. Palmer, M.W., Maurer, T., 1997. Does diversity beget diversity? A case study of crops and weeds. J. Veg. Sci. 8, 235–240. Pe´rez de Paz, P.L., Del-Arco, M., Acebes, J.R., Wildpret, W., 1990. La vegetacio´n cormofı´tica (vascular) del Parque Nacional de Garajonay. In: Pe´rez de Paz, P.L. (Ed.), Parque Nacional de Garajonay, pp. 137–172. Patrimonio Mundial. ICONA y Excmo. Cabildo insular de La Gomera. Pimm, S.L., 1991. The Balance of Nature? University of Chicago Press, Chicago. Planty-Tabacchi, A.-M., Tabacchi, E., Naiman, R.J., De Ferrari, C., Decamps, H., 1996. Invasibility of species-rich communities in riparian zones. Conserv. Biol. 10, 598–607. Prieur-Richard, A.-H., Lavorel, S., 2000. Invasions: the perspective of diverse plant communities. Aust. J. Ecol 25, 1–7. Pysˇek, A., Pysˇek, P., Jarosik, V., Hajek, M., Wild, J., 2003. Diversity of native and alien plant species on rubbish dumps: effects of dump age, environmental factors and toxicity. Divers. Distributions 9, 177–189. Quinn, G.P., Keough, M.J., 2002. Experimental Design and Data Analysis for Biologists. Cambridge University Press, Cambridge. Robinson, G.R., Quinn, J.F., Stanton, M.L., 1995. Invasibility of experimental habitat islands in a California winter annual grassland. Ecology 76, 786–794. Stohlgren, T.J., Schell, L.D., Bull, K.A., Otsuki, Y., Newman, G., Bashkin, M., Son, Y., Binkley, D., Chong, G.W., Kalkhan, M.A., 1999. Exotic plant species invade hot spots of native plant diversity. Ecol. Monogr 69, 25–46. Suarez, E.R., Pelletier, D.M., Fahey, T.J., Groffman, P.M., Bohlen, P.J., Fisk, M.C., 2004. Effects of exotic earthworms on soil
acta oecologica 34 (2008) 202–213
phosphorus cycling in two broadleaf temperate forests. Ecosystems 7, 28–44. Tilman, D., 1997. Community invasibility, recruitment limitation, and grassland biodiversity. Ecology 78, 81–92. Tilman, D., 1999. The ecological consequences of changes in biodiversity: A search for general principles. Ecology 80, 1455–1474. Tilman, D., Wedin, D., Knops, J., 1996. Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature 379, 718–720. Vitousek, P.M., 1990. Biological invasions and ecosystem processes: towards an integration of population biology and ecosystem studies. Oikos 57, 7–13.
213
Vicente, M.C., Enghoff, H., 1999. The millipedes of the Canary Islands (Myriapoda: Diplopoda). Vieraea 27, 183–204. Voigt, W., Perner, J., Jones, T.H., 2007. Using functional groups to investigate community response to environmental changes: two grassland case studies. Global Change Biology 13, 1710–1721. Volz, F., 1959. Globalstrahlung auf geneigten Ha¨ngen. Meteorol. Rundsch 11, 132–135. Walter, H., Breckle, S.-W., 1999. Vegetation und Klimazonen. Ulmer, Stuttgart. Williamson, M., 1996. Biological Invasions. Chapman and Hall, London.