Invasion patterns of ground-dwelling arthropods in Canarian laurel forests

Invasion patterns of ground-dwelling arthropods in Canarian laurel forests

acta oecologica 34 (2008) 202–213 available at www.sciencedirect.com journal homepage: www.elsevier.com/locate/actoec Original article Invasion pa...

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acta oecologica 34 (2008) 202–213

available at www.sciencedirect.com

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

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

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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

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