Species Richness and Composition Patterns Across Trophic Levels of True Bugs (Heteroptera) in the Agricultural Landscape of the Lower Reach of the Tisza River Basin

J Insect Conserv DOI 10.1007/s10841-012-9484-1

ORIGINAL PAPER

Species richness and composition patterns across trophic levels of true bugs (Heteroptera) in the agricultural landscape of the lower reach of the Tisza River Basin

Attila Torma • Pe´ter Csa´sza´r

Received: 30 September 2011 / Accepted: 25 March 2012 Ó Springer Science+Business Media B.V. 2012

Abstract River basins are among the most threatened the remaining grasslands of the lower reach of the Tisza ecosystems. The species diversity of several European river River Basin. basins decreased seriously during the last decade due to loss of habitats and increasing land use pressure on the Keywords Predaceous Herbivorous Insect diversity remaining habitats. We studied true bug assemblages in Land use Dike Riparian grasslands various land use types of grassland fragments and dikes as linear grassland habitats in the agricultural landscape of the lower reach of the Tisza River Basin. We tested the effects Introduction of the recorded variables of habitat quality, surrounding landscape and land use type on the abundance, species In their natural states, riverine landscapes are characterised richness and composition of true bugs. Altogether, 5,389 as a mosaic habitat-complex with considerable longitudinal adult Heteroptera individuals representing 149 species in extension. Due to the high connectivity and the heteroge- 13 families were collected. The factors which influenced neity of riverine landscapes (Naiman et al. 2005), flood- significantly the species richness of different trophic levels plains can support a diverse flora and fauna (Gregory et al. (i.e. herbivors, predators) and degrees of food specializa- 1991; Zwick 1992; Ward et al. 1999). However, river tion (i.e. generalist and specialist herbivors) were concor- controls have caused dramatic changes in rivers and their dant. Contrary to this, the factors which influenced the riparian zones (Dynesius and Nilsson 1994; Nilsson et al. abundance of the different feeding groups varied strongly. 2005). Nowadays, many European rivers are restricted to We emphasise the vegetation and land use types as pri- narrow riverbeds bordered by rigid dikes, riverbeds are marily influential factors for insects. Excluding the grass- often straightened and the majority of riparian habitats are feeding species, the number of both generalist, specialist transformed into intensive agricultural areas (Tockner et al. herbivorous and predaceous species were lower in agri- 2009). In the last decade Central European floodplains have cultural swards, i.e. hay-meadows and pastures than in old experienced a decline in biodiversity, due to a loss of field and dike habitats and their number increased with habitats and increased land use pressure on the remaining increasing vegetation diversity. Due to the high species habitats (Godreau et al. 1999; McCollin et al. 2000). The richness and abundance observed in dike and old field intensification of agricultural land use caused a decrease in habitats compared to agricultural swards, we emphasise species richness in most European countries (Marshall and their importance for conservation of insect diversity and we Moonen 2002) due to the destruction, fragmentation and stress the negative effects of agricultural intensification on isolation of the natural habitats (Tscharntke and Kruess 1999; Steffan-Dewenter and Tscharntke 2000). In the present study we focus on the effect of land use type, landscape structure and habitat quality on the species & A. Torma ( ) P. Csa´sza´r diversity and composition patterns of true bugs in the Department of Ecology, University of Szeged, Ko¨ze´p fasor 52, Szeged 6726, Hungary agricultural landscape of the lower reach of the Tisza River e-mail: [email protected] Basin. 123 J Insect Conserv

The River Tisza is the second largest river in Hungary they represent an ecologically diverse group including both and presumably the most characteristic river in the Carpa- predaceous and herbivorous species with different degrees thian Basin which is a separated biogeographical region of food specialization (Dolling 1991; Schuh and Slater (Pannon region) in Europe. The Tisza is 965 km long and its 1995; Schaefer and Panizzi 2000) and the various trophic catchment includes most of the Carpathian Mountains levels are supposed to be influenced by different factors covering approximately 157,000 km2 (Alfo¨ldi and (Zurbru¨gg and Frank 2006; Hines et al. 2005); moreover, Schweitzer 2003; Sommerwerk et al. 2009). The Tisza River influential factors could differ between the degrees of Basin is regarded as an important core area and also a green feeding specialization (Jonsen and Fahrig 1997). Finally, corridor (Galle´ et al. 1995 Galle´ 2005;IUCN1995), con- Heteroptera is a good indicator group of total insect taining highly endangered habitats (Decamps 1993; Frank- diversity in agricultural ecosystems (Duelli and Obrist lin 1993) and contributing to the conservation of 1998). biodiversity in Europe (Sze´p 1997; Sommerwerk et al. Here we address the following: 2009). The Tisza River Basin has high plant and animal 1. To reveal the effects of environmental controls (i.e. diversities (Galle´ 2005; Sommerwerk et al. 2009) and land use type, surrounding landscape structure and unique insect species as the endemic Tisza mayfly (Palin- vegetation) on the structure, species composition and genia longicauda Olivier, 1791) and the European stalk- distribution of true bugs. eyed fly (Sphyracephala europaea Papp et Fo¨ldva´ri 1997). 2. To test the significance of the environmental controls Although the major works of the so-called ‘‘regulation of on the species richness and abundance of true bugs Tisza’’ in the nineteenth century divided the historical belonging to different trophic levels (i.e. herbivores floodplain by dikes into flooded and nonflooded parts, the and predators) and with different ranges of host-plant extensively used, mosaic landscape preserved several nat- specialization (i.e. generalist and specialist ural habitats, i.e. pastures, woody pastures and hay-mead- herbivores). ows until the 1950s (Dea´k 2007; Sendzimir et al. 2008), and 3. To test whether the species composition of Heteroptera the traditional, Hungarian land use types (see e.g. Hortoba´gy assemblages differs significantly within a dike as a a site of the UNESCO World Heritage) maintained high linear landscape element at the landscape scale. biodiversity (Aradi and Lengyel 2003). Intensification of agriculture during the socialist era, reduced the area of grasslands (Dea´k 2007). Nowadays, the remaining grasslands are enclosed between highly modified landscape ele- Materials and methods ments, i.e. arable fields and forest plantations of mainly non- native tree species. In this agricultural landscape mosaic, the Study sites and sampling dikes form grassland-strip habitats. Dikes function as grassland habitats and corridors for insects (Galle´ et al. Two habitat complexes of the lower Tisza-valley were 1995; Krausz et al. 1995) and are of considerable length: selected for sampling in Csongra´d county (Fig. 1). The approximately 4,500 km of primary and secondary dikes structure of the landscape and the land use differed exist along the River Tisza and its tributaries in Hungary between the two habitat complexes. The Ta´pe´ habitat (Sendzimir et al. 2008). In general, linear grassy elements complex is situated near Szeged, where the landscape are supposed to enlarge insect habitats and to reduce frag- consists of arable fields with small patches of meadows, old mentation (Nicholls et al. 2001; Pywell et al. 2005;O¨ ckin- fields and forest plantations embedded in the matrix of ger and Smith 2007;So¨derstro¨m and Hedblom 2007), thus arable fields. The Doć habitat complex lies approximately the conservation, restoration and management of linear 30 km north of Szeged in a more natural landscape. Higher grassy elements of agricultural landscape is of concern in proportion of semi-natural grasslands and forests occur in many parts of Europe (e.g. Noordijk et al. 2009; Tikka et al. the latter. In both of the two habitat complexes we studied 2001; Saarinen et al. 2005; de la Penã et al. 2003). typical land use types, i.e. hay-meadows, pastures, old Heteroptera is an ideal group to study diversity patterns fields and dikes. Six and four grassland habitats were in fragmented and managed grasslands as firstly they are selected for the study in Doć and Ta´pe´ habitat complexes, known to be influenced by various vegetation properties respectively (for further details, see Table 1). In each (Sanderson et al. 1995; Schwab et al. 2002; Littlewood habitat, Heteroptera assemblages were sampled in three et al. 2009; Bro¨ring and Wiegleb 2005; Frank and Ku¨nzle plots. The distance between the plots within a grassland 2006; Torma and Ko¨rmo¨czi 2009; Torma et al. 2010; Galle´ habitat was approximately 250 m. True bugs were col- et al. 2010) and respond quickly to grassland disturbances lected by sweep netting, which is an adequate technique for such as mowing, grazing, burning and polluting (Morris sampling the structure and species diversity of assemblages 1973, 1975, 1979, 1990; Bra¨ndle et al. 2001). Secondly, (Remane 1958; Standen 2000; Coscaron et al. 2009). In 123 J Insect Conserv

radius, to reveal the scale which had the major influence on the assemblages. Finally, according to land use type; hay- meadow, pasture, old field and dike habitats were distinguished and they were defined as factors in the analyses.

Data analysis

To show the natural grouping of samples according to the species composition of true bugs, Non-metric Multidi- mensional Scaling (NMDS) with Bray-Curtis dissimilarity was used (Legendre and Legendre 1998; Oksanen 2011). We used permutational MANOVA (Permanova) based on Bray-Curtis similarity (Vegan R package, Oksanen et al. Fig. 1 The location of the studied habitat-complexes in Hungary. The above and below black circles mark Doć and Ta´pe´ habitat- 2010) to reveal the variation within the dike habitat in the complexes, respectively species composition of true bugs. The differences were tested between the samples situated in the different habitat- each plot 3 9 50 sweeps were applied two times in the complexes as well as between the samples situated in the summer of 2009. First sampling (30 May–1 June) was floodplain and on the protected side of the dike. applied before cutting, and the second period (20–22 July) To determine the variables which influenced the struc- was when the vegetation had already started regrowth after ture and species composition of Heteroptera assemblages, cutting. For data analyses, the data of sweeps in each plot constrained ordination (CCA—Canonical Correspondence were pooled resulting in 30 samples altogether. Analysis) was used (Lepsˇ and Sˇmilauer 2003; Oksanen 2011). The marginal and partial effects of the variables Assessments of variables were calculated and were tested for significance by a Monte Carlo procedure with 1,000 permutations. To visu- In each plot the habitat quality, the landscape structure and alise the effects of the significant variables, they were fitted the land-use type were assessed. To characterize the habitat to a scatterplot of NMDS using the envfit function (Ok- quality we used the properties of the vegetation, which sanen 2011). We used the Vegan R package (Oksanen et al. were sampled in three 1 9 1 m quadrates in each plot. The 2010) to carry out the ordinations. mean data of the three quadrats was used to define vari- Mantel test based on Bray-Curtis similarity was used to ables. To characterize the structure of the vegetation we analyse the relationship between the species distribution of recorded the average height of the vegetation (plantheight), Heteroptera assemblages and vegetation. the total cover of the vegetation (totcov), the cover of the To detect species which were associated with land use vegetation at 10 and 40 cm above the ground (cov10 and types, we applied the indicator value approach (Dufrene cov40). The total cover of vegetation was calculated as the and Legendre 1997). Indicator values were calculated using sum of the percentage cover of all plant species. The cover the labdsv package (Roberts 2010) of R statistic and the of vegetation at 10 and 40 cm above ground based on the statistical significance of the indicator values tested for percentage cover of those plants which reached at least the significance with a Monte Carlo procedure with 1,000 height of 10 and 40 cm, respectively. We also estimated permutations. the percentage cover of dead vegetation (deadveg). To We used the univariate regression tree method (URT) to characterize the richness of the vegetation we used the identify the influential explanatory variables (Crawley number of plant species (plants) and we calculated the 2007). The method performs a binary recursive partitioning diversity of vegetation (vegdiv) based on the percentage of the dataset. It also offers the opportunity to gain insight cover of plant species in the quadrats using the Shannon- into the structure of the data and identify the univariate Wiener index. To assess the features of the landscape we interactions between the variables (Tree Package, Ripley measured the proportion of habitat types—i.e. grasslands, 2009). The explanatory variables suggested by the URT forests, arable fields (grassland%, forest%, arable method were subjected to a linear mixed-effect model field%)—in a radius of 50, 100, 250, 500, 750 m around (GLMM) with nested random effects to test the signifi- each plot based on aerial photographs using Arcview 3.11 cance of their influence on species richness and abundance. GIS software. The remaining landscape elements (e.g. The effects of the different habitat complexes and grass- surface of the river, roads and buildings) were omitted lands were used as random effects and the explanatory because of their relative low proportion. We calculated the variables were used as fixed effects (Crawley 2007). The effects of the proportion of habitat types in each different species richness and abundance data were modelled 123 123 Table 1 Features of the study sites Habitat complex Habi-tat Land use type Coordinates and notes Disturbances and processes Vegetation

Doć D1 Hay meadow N46° 260 39,1300 Regular mowing Exclusively dominated by Alopecurus Arable fields %: 12.97 ± 1.79 E20° 070 43,4600 pratensis and Agropyron repens Forests %: 44.55 ± 4.07 D2 Pasture N46° 260 51,1300 Extensive grazing by cattle Dominated by grasses, few forbs occurred; Grasslands %: 36.39 ± 6.27 E20° 080 55,2100 dominant grasses: Festuca pseudovina, Botriochloa ischaemum; forbs: Potentilla collina, Lotus corniculatus, Artemisia santonicum D3 Pasture N46° 260 40,5200 Extensive gra-zing by sheep Dominated by grasses, few forbs occurred; E20° 090 36,9200 dominant grasses: F. pseudovina, Poa angustifolia; forbs: Podospermum canum D4 Dike meadow N46° 260 19,5400 Irregular flooding in the Dominated by grasses, reach in forbs; E20° 100 15,8900 floodplain side of the dominant grasses: F. pseudovina, A. dike; annual maintenance pratensis, Arrhenatherum elatius; dominant Secondary dike of the dikes is directed by forbs: Salvia nemorosa and other the law LVII/1995 e.g. Lamiaceae, Lathyrus tuberosus, Verbena mowing, uprooting officinalis D5 Dike meadow N46° 260 46,5200 bushes, etc. Dominated by A. elatius, dominant grasses: E20° 110 40,0900 A. elatius, F. pseudovina; dominant forbs: Galium rubioides, Aristolochia clematitis, Floodplain side S. nemorosa D6 Dike meadow N46° 260 47,9300 E20° 110 41,2000 Dominated by A. elatius; dominant grasses: Nonflooded side A. elatius, F. pseudovina; dominant forbs: G. rubioides, S. nemorosa Ta´pe´ T3 Dike meadow N46° 170 16,5300 Dominated by A. elatius; dominant grasses: Arable fields %: 51.34 ± 3.41 E20° 140 20,3800 A. elatius; dominant forbs: G. rubioides, A. clematitis Forests%: 21.00 ± 6.33 Floodplain side Grasslands %:22.97 ± 3.22 T4 Dike meadow N46° 170 15,8700 High density of A. elatius; dominant grasses: E20° 140 21,1200 A. elatius, Setaria viridis; dominant forbs: G. rubioides, S. nemorosa, Nonflooded side T1 Pasture N46° 170 11,1800 Extensive grazing by cattle High density of Limonium gmellini, rich in E20° 120 07,1700 species; dominant grasses: P. angustifolia, F. pseudovina; dominant forbs: L. gmellini, Achillea millefolium, A. santonicum, Matricaria imodora T2 Old field N46° 170 34,8700 Succession Dominated by Lamiaceae; low density of

E20° 130 02,2400 grasses: A. repens, P. angustifolia; Conserv Insect J dominant forbs: Lamiaceae (e.g. Lamium, (ca. 4 years old) Salvia), L. tuberosus The habitat complexes are characterized by the proportion of habitats (mean ± SD) around the plots J Insect Conserv following a Poisson error distribution using the library (e.g. Nabis spp., Orius minutus (Linnaeus, 1758), Dicyphus lme4 (Bates et al. 2011). The analyses were carried out geniculatus Fieber, 1858) species were associated with old with R statistic (R Development Core Team 2007). fields and dikes but only four common grass-feeding spe- As heteropteran bugs are diverse in terms of trophic cies were associated with the agricultural swards. Several level, we analysed species richness and abundance species were found in only one of the studied habitats, belonging to different trophic levels and different degrees especially those which were associated with alkaline of food specialization. We distinguished four groups: (1) grasslands, e.g. Henestaris halophilus (Burmeister, 1835), specialist herbivore group containing monophagous and Solenoxyphus fuscovenosus (Fieber, 1864), Polymerus strictly oligophagous species sucking various herbs but not cognatus (Fieber, 1858). Furthermore, the species collected grasses; (2) grass-feeding bugs containing mono- and exclusively in dikes were vagrant forest species, e.g. Hi- mainly oligofagous species feeding exclusively on grasses. macerus apterus (Fabricius, 1798), Heterotoma merioptera We separated the grass-feeding species because of their (Scopoli, 1763), Agnocoris rubicundus Fallen, 1807. high dominance and relevance in grassland habitats. (3) According to the Canonical Correspondence Analysis, The group of generalist herbivores containing broadly Heteroptera assemblages were mostly influenced by the oligophagous and polyphagous species and (4) the predator proportion of the surrounding habitat types in a radius of group containing the zoophagous and zoo-phytophagous 500 m (Fig. 2). Therefore, we used the proportion of the species. We used the appropriate volumes of Fauna Hun- surrounding grasslands, forests, arable fields in a radius of gariae (Hala´szfy 1959; Benedek 1969;Va´sa´rhelyi 1978, 500 m around the plots for further analyses. 1983), Fauna Roma˘niei (Kis 1984, 2001), Die Tierwelt The scatterplot of NMDS (Fig. 3) showed that the sam- Deutschland (Wagner 1952, 1966, 1967) and Faune de ples aggregated into three groups mainly in accordance with France (Pe´ricart 1983, 1984, 1998a, b, c) to assess the land use types. Samples of dikes, old fields and agricultural ecological requirements of species. swards, i.e. hay-meadows and pastures, separated well from each other. The samples of agricultural swards, however, were scattered, and this suggested a higher variability in the Results assemblage structure between and within the agricultural grasslands. The samples belonging to the different habitat Characteristics of true bug assemblages complexes did not separate, which suggests an influence of habitat type but not geographical distance on Heteroptera We collected 5,389 adult individuals representing 149 assemblages at the landscape scale. The grouping of samples species in 15 families (‘‘Appendix’’). The list of true bugs also indicates little variation in the species composition of contained 17 grass-feeding (1,574 specimens), 58 gener- Heteroptera assemblages within the dike habitat. The results alist (1,650 specimens) and 49 specialist (1,253 specimens) of Permanova confirmed this assumption. We did not find phytophagous species as well as 834 individuals of 16 significant differences in the species composition of the predaceous species. We could not sort nine species (78 Heteroptera assemblages of dikes between the two different specimens) into the groups, because of their little-known habitat complexes (pseudo-F = 1.83; P = 0.068). Further- ecological requirements. more, the species composition of samples situated on the The most frequent Heteroptera species were Orius niger protected and floodplain sides of the dikes did not differ Wolff, 1804 (11.2 % of all collected individuals), Lasia- either (pseudo-F = 1.54; P = 0.115). cantha capucina (Germar, 1837) (11.1%), Halticus apterus (Linnaeus, 1761) (7.3 %) and Acetropis longirostris Factors influencing the composition, species number (Puton, 1875) (5.2 %). We collected several species with a and abundance of true bug assemblages frequency of 1–5 % which were mostly common grass- feeding species e.g. Stenodema calcaratum (Falleń, 1807), According to CCA, most of the variables had significant Notostira elongata (Geoffroy 1785), Myrmus miriformis marginal effects on true bug assemblages except for the (Falleń, 1807), Aelia acuminata (Linnaeus, 1758). Beside total cover of vegetation and the cover of vegetation at a O. niger, further predaceous species e.g. Nabis pseudoferus height of 40 cm above ground (Table 3, Fig. 3). Habitat Remane 1949 and N. punctatus Costa, 1847 occurred with quality variables had no significant partial effects, which a considerable abundance. According to the indicator value showed the overlapping effect of the variables. Total approach, several species were related to different land use inertia was 5.032, of which our model explained 66.5 % types, especially to old field and dike habitats (Table 2). (pseudo-F(13,16) = 2.74; P \ 0.001). Both surrounding Both generalist (e.g. Chlamydatus spp. Lygus rugulipennis landscape (pseudo-F(3,26) = 3.15; P = 0.005), habitat Poppius, 1911, H. apterus) and specialist phytophagous quality (pseudo-F(7,22) = 1.86; P = 0.005) and land-use (e.g. L. capucina, Adelphocoris seticornis) and predaceous type (pseudo-F(3,26) = 4.63; P = 0.005) had a significant 123 J Insect Conserv

Table 2 List of species Species Land use type Indicator value associated with land use types, according to the Dufrene- Acetropis longirostris (Puton, 1875) Hey meadow 99*** Legendre indicator value approach Capsus ater (Linnaeus, 1758) Hey meadow 80** Aelia rostrata Boheman, 1852 Pasture 51* Chorosoma schillingi (Schummel, 1829) Pasture 44* Lasiacantha capucina (Germar, 1836) Old field 98*** Sciocoris cursitans (Fabricius, 1794) Old field 98*** Chlamydatus pullus Reuter, 1870 Old field 96*** Systellonotus triguttatus (Linnaeus, 1767) Old field 95*** Adelphocoris lineolatus (Goeze, 1778) Old field 87*** Chlamydatus pulicarius (Falleń, 1807) Old field 85*** Megalocoleus molliculus (Falleń, 1829) Old field 84*** Adelphocoris seticornis (Fabricius, 1775) Old field 72** Nabis pseudoferus Remane 1949 Old field 72** Trigonotylus caelestialium (Kirkaldy, 1902) Old field 63* Pterotmetus staphyliniformis (Schilling, 1829) Old field 62* Trigonotylus pulchellus (Hahn, 1834) Old field 61* Nabis punctatus Costa, 1847 Old field 53* Lygus rugulipennis Poppius, 1911 Old field 50* Platyplax salviae (Schilling, 1829) Old field 47* Stenodema calcaratum (Falleń, 1807) Dike meadow 84*** Halticus apterus (Linnaeus, 1761) Dike meadow 81*** Stictopleurus punctatonervosus (Goeze, 1778) Dike meadow 80** Stenotus binotatus (Fabricius, 1794) Dike meadow 78** Eusarcoris aeneus (Scopoli, 1763) Dike meadow 73** Megaloceroea reticornis (Geoffroy, 1785) Dike meadow 71** Dicyphus geniculatus (Fieber, 1858) Dike meadow 60* Eurydema oleraceum (Linnaeus, 1758) Dike meadow 60* Level of significance: * Orius minutus (Linnaeus, 1758) Dike meadow 60* P \ 0.05, ** P \ 0.01, *** Orthocephalus saltator (Hahn, 1835) Dike meadow 50* P \ 0.001

(16.3 %), whereas the overlapping effects of habitat quality and surrounding landscape (11.0 %) and of land-use type and surrounding landscape (11.3 %) were almost the same. The Mantel test showed significant relationship between the species composition of Heteroptera assemblages and the vegetation (r = 0.56; P \ 0.001). The URT method showed that the species number of true bug assemblages was influenced by land use type, vegetation diversity, cover of vegetation at a height of 40 cm above ground and the proportion of surrounding grasslands (Fig. 4). According to the GLMM, total species richness was significantly lower in pastures (z =-5.91; Fig. 2 The influence of the surrounding habitats on Heteroptera assemblages within different radii around the sampling plots accord- P \ 0.001) and hay-meadows (z =-3.99; P \ 0.001) ing to CCA. The surrounding habitats within a radius of 500 m than in dike habitats. However, there was no significant explained the most of the total inertia difference between dikes and old fields (z =-1.22; P = 0.224). Vegetation diversity affected the total species influence on true bug assemblages with an explained inertia richness of true bugs positively (z = 4.13; P \ 0.001), but of 26.7, 37.1 and 34.8 %, respectively. The overlapping the effect of vegetation coverage at a height of 40 cm effect of habitat quality and land-use type was higher above ground on the total species richness was not

123 J Insect Conserv

Table 3 Effects and signiﬁcance of the variables inﬂuencing the structure of true bug assemblages according to CCA v2 d.f. pseudo- P F

Surrounding grasslands (%) 0.583 1.28 3.673 0.001 (0.189) (1.16) (1.800) (0.012) Surrounding forests (%) 0.465 1.28 2.854 0.001 (0.167) (1.16) (1.585) (0.041) Surrounding arable fields (%) 0.551 1.28 3.443 0.001 (0.178) (1.16) (1.689) (0.017) Average height of vegetation 0.388 1.28 2.337 0.006 (0.141) (1.16) (1.339) (0.077) Total cover of vegetation 0.240 1.28 1.404 0.163 (0.074) (1.16) (0.705) (0.719) Cover of vegetation at a height 0.434 1.28 2.636 0.001 of 10 cm above ground (0.095) (1.16) (0.902) (0.473) Cover of vegetation at a height 0.261 1.28 1.529 0.110 Fig. 3 NMDS scatterplot showing the natural grouping of samples of 40 cm above ground (0.131) (1.16) (1.248) (0.141) based on the whole data of collected true bugs and the effects of the significant variables. Samples are marked with a letter and a number. Number of plant species 0.432 1.28 2.632 0.001 D and T mean Doć and Ta´pe´ habitat-complexes, respectively, and (0.108) (1.16) (1.029) (0.298) each grassland is marked with a number (for more details, see Diversity of vegetation 0.397 1.28 2.400 0.008 Table 1). The arrows indicate the effect of variables influencing the (0.117) (1.16) (1.112) (0.232) structure of true bug assemblages according to CCA (see Table 3). Abbreviation of variables: see ‘‘Materials and methods’’ Proportion of dead vegetation 0.406 1.28 2.457 0.005 (0.099) (1.16) (0.936) (0.395) Land use type 1.753 3.26 4.632 0.001 significant (z = 1.32; P = 1.187). The proportion of (0.693) (3.16) (2.196) (0.001) surrounding grasslands affected the total species richness The marginal and partial (in parentheses) effects are given of Heteroptera assemblages negatively (z =-6.57; P \ 0.001). The total abundance of true bugs was affected by land use type, particularly in the Doć habitat-complex (Fig. 4), and according to GLMM, it decreased significantly only in pastures (z =-4.09; P \ 0.001). In hay-meadows and old fields the abundance did not differ significantly from the abundance in dike habitats (z =-0.73; P = 0.467 and z = 1.54; P = 0.124, respectively). The vegetation coverage at a height of 10 cm above ground also influenced the total abundance of bugs significantly (z = 4.74; P \ 0.001), as well as the proportion of surrounding arable fields (z = 7.08; P \ 0.001).

Factors influencing the species richness of the different feeding groups of true bugs Fig. 4 Univariate regression tree (URT) for the species richness (a) and abundance (b) of total true bug assemblages. For each node a threshold value of the variable is given that determines branches. In The URT method showed that similarly to the total the case of the branches based on land use, the type of land use is assemblages, the species richness of the different groups of given instead of a value. The different lengths of the lines following true bugs was mainly influenced by land use type and each split are proportional to the variance explained by the split. The vegetation diversity of grasslands (Fig. 5). Land use type numbers below each leaf show the number of samples falling into that leaf. Abbreviations of variables: see ‘‘Materials and methods’’ affected the number of generalist, specialist phytophagous and predaceous species. The species richness of the various herbivores: z =-3.70; P \ 0.001 and z =-5.93; feeding groups was significantly lower in regularly mown P \ 0.001, specialist herbivores: z =-3.17; P = 0.001 or grazed grasslands than in dike habitats (generalist and z =-4.80; P \ 0.001, predators: z =-2.83; 123 J Insect Conserv

P = 0.005 and z =-4.41; P \ 0.001 for hay-meadows and pastures, respectively). In contrast to agriculturally used meadows, the species richness was not significantly lower in old field samples than in dike samples (generalist herbivores: z =-0.04; P = 0.968, specialist herbivores: z =-1.61; P = 0.108, predators: z =-0.52; P = 602). Vegetation diversity affected the species richness of generalist (z = 3.08; P = 0.002) and specialist (z = 4.57; P \ 0.001) herbivores and predators positively (z = 3.80; P \ 0.001), but it did not influence significantly the number of grass-feeding species (z = 0.15; P = 0.880). Beyond vegetation diversity, a significant influence of the number of plant species on the species number of predaceous bugs (z = 3.11; P = 0.002) was found. The effects of the other variables of habitat quality suggested by URT, was not, or just marginally significant in any case according to GLMM, i.e. the effect of vegetation coverage at 40 cm above ground on the species richness of both generalist herbivores (z = 1.62; P = 0.106), specialist herbivores (z = 1.67; P = 0.095) and predators (z = 1.84; P = 0.066), as well as the effect of the vegetation height and vegetation coverage at a height of 10 cm above ground on the number of grass-feeding species (z = 0.85; P = 0.397 and z = 0.81; P = 0.420). The URT method also showed that the surrounding habitat types had little impact on species richness of true bugs. According to GLMM, the impact of the proportion of Fig. 5 Univariate regression tree (URT) for the species richness of grass-feeding (a), predatory (b), generalist herbivorous (c) and surrounding forests on the number of generalist herbivo- specialist herbivorous (d) true bugs. For each node a threshold value rous species was not significant (z = 0.93; P = 0.350), of the variable is given that determines branches. In the case of the whereas the proportion of surrounding arable fields had a branches based on land use, the type of land use is given instead of a significant negative influence on the species richness of value. The different lengths of the lines following each split are proportional to the variance explained by the split. The numbers grass-feeding true bugs (z =-2.65; P = 0.008). below each leaf show the number of samples falling into that leaf. Abbreviations of variables: see ‘‘Materials and methods’’

Factors influencing the abundance of the different feeding groups of true bugs the diversity, the height and the coverage of the vegetation) on the abundance of true bugs. According to GLMM, we The abundance of true bugs was less affected by land use found that vegetation diversity influenced the abundance of type then was their species richness (Fig. 6). Land use type specialist herbivores negatively (z =-8.93; P \ 0.001); affected only the abundance of predaceous species. the number of plant species and the vegetation height According to GLMM, the abundance of predaceous species influenced the abundance of grass-feeding species nega- was significantly lower in hay-meadows and pastures than tively (z =-2.76; P = 0.006 and z =-2.74; P = 0.006), in dike habitats (z =-3.45; P \ 0.001 and z =-4.57; whereas the abundance of predatory species was affected P \ 0.001, respectively), but we did not find significant by the vegetation coverage at a height of 40 cm and by the differences between old field and dike habitats (z = 0.36; proportion of dead vegetation (z = 3.90; P \ 0.001 and P = 0.720). The variables of surrounding landscape, as z =-5.33; P \ 0.001). The URT method indicated the suggested by the URT, also had little impact on the influence of vegetation diversity on the abundance of grass- abundance of true bugs. The effect of the proportion of feeding species but this effect was not significant according arable fields on the abundance of specialist and generalist to GLMM (z = 0.67; P = 0.504), neither was there an herbivores was not significant according to GLMM effect of plant species richness (z =-1.53; P = 0.126) or (z = 1.39; P = 0.164 and z = 1.26; P = 0.207, respec- the cover of vegetation at 10 cm above ground (z = 1.53; tively). The URT method suggested a major influence of P = 0.136) on the abundance of generalist herbivorous various vegetation properties (i.e. number of plant species, species (Fig. 6). 123 J Insect Conserv

Hildebrandt 2003). The species number of other feeding groups was strongly influenced by land use types and was higher in old field and dike habitats than in agricultural grasslands. High arthropod species richness of Tisza dikes was also reported previously (e.g. Galle´ et al. 1995; Krausz et al. 1995). In addition to the dikes of the River Tisza being affected by the migration of insect species (Galle´ et al. 1995; Krausz et al. 1995), we emphasize the source and habitat functions of dikes for true bugs. Many areas are supposed to be potential corridors, i.e. grass strips, hedgerows, road verges could realize their greatest function as habitats for resident arthropods (e.g. Maelfait and De Keer 1990; Feber et al. 1996; Pywell et al. 2005; Helden and Leather 2004; Noordijk et al. 2009). The similarity of Heteroptera assemblages between dike habitats situated in a distance of 30 km from each other suggests strong con- nection along the dike. However, this is presumably operates at small geographical scales and insect communities differed clearly at a larger spatial scale, e.g. between dike habitats of the upper and lower reaches of the River Tisza (Galle´ et al. 1995). Due to the connectivity of river landscapes (Naiman et al. 2005), tributaries also influence the species composition of insect assemblages at the landscape scale (Krausz et al. 1995). Although we did not Fig. 6 Univariate regression tree (URT) for the abundance of grass- find significant differences in the species composition of feeding (a), predatory (b), generalist herbivorous (c) and specialist true bugs between the flooded and nonflooded sides of the herbivorous (d) true bugs. For each node a threshold value of the dike, it is presumably not a general pattern, as e.g. variable is given that determines branches. In the case of the branches Orthoptera (Krausz et al. 1995) and spider (Galle´ et al. based on land use, the type of land use is given instead of a value. The different lengths of the lines following each split are proportional to 2011) assemblages differed between the two sides. The the variance explained by the split. The numbers below each leaf structure and composition of animal assemblages can show the number of samples falling into that leaf. Abbreviations of change rapidly in riparian areas as they are affected variables: see ‘‘Materials and methods’’ strongly by flooding events (Lambeets et al. 2009). High species richness and abundance of true bugs were Discussion also observed previously during early succession stages (Southwood et al. 1979; Brown 1982; Frank and Ku¨nzle Or findings that land use type and vegetation affected 2006). We stress the higher species richness and abundance primarily the species composition and richness of Het- of predaceous true bugs in an early succession stage of an eroptera assemblages in an agricultural area are similar to abandoned field comparing to managed swards. This find- those of several authors (e.g. Zurbru¨gg and Frank 2006; ing suggests the importance of wildflower areas and Frank and Ku¨nzle 2006; Di Giulio et al. 2001). We found abandoned fields near to agricultural fields as remnants and little variation, however, in the number of the grass-feeding shelters for predaceous bugs and possibly for other natural species between the various land use types, in spite of the enemies. However, the distribution pattern of predaceous fact that some species were associated with certain habitat Heteropterans could be influenced by various other factors, types or grass species in the Great Hungarian Plain (e.g. A. e.g. connectivity (Nicholls et al. 2001), succession age longirostris in Alopecurus grasslands, Amblytylus nasutus (Frank and Ku¨nzle 2006), and pest management in agri- Kirschbaum, 1856, S. calcaratus in Poa pratensis stands cultural areas (Kinkorova´ and Kocourek 2000). Moreover, (Koppańyi 1965) as well as A. carinata, Chorosoma spp., the colonization pattern and thereby the effective role as etc. in Fescue meadows (Torma et al. 2010)). In general, natural enemies can differ between the brachypterous and grass-feeding species (Nickel and Hildebrandt 2003) and macropterous forms within a species (Roth 2003). common polyphagous species (Di Giulio et al. 2001) Leaving the grass-feeding bugs out of consideration, the showed little variation between differently managed species number of Heteropteran bugs was found to be grasslands, contrary to monophagous herbivors (Nickel and strongly influenced by land use type. However, we agree 123 J Insect Conserv with Nickel and Hildebrandt (2003) that there is a general predatory arthropods (Entling et al. 2007; Lambeets et al. need for distinction between short-term and long-term 2008, 2009; Heikkinen and MacMahon 2004; Woodcock effects of grassland management. Land use type can et al. 2007; Galle´ and Torma 2009; Galle´ et al. 2010, 2011). change the structure and species composition of vegetation Another aspect of vegetation effects on predators is that (Collins et al. 1998;Gu¨sewell et al. 1998) and thereby it vegetation affected particularly the generalist predators has an indirect or long-term effect on invertebrate assem- (Koricheva et al. 2000), as a specialised enemy should be blages by affecting host availability and habitat conditions less sensitive to plant diversity than a generalist predator (Curry 1994; Tscharntke and Greiler 1995; Nickel and because the former is less dependent on the diversity of Hildebrandt 2003). In this sense we stress the influence of alternative food offered in a more diverse habitat (Sheehan indirect effects of land use on true bugs via changing the 1986). The collected predatory Heteropterans (e.g. Nabis species composition and diversity of vegetation. Other spp., Orius spp.) are mainly generalist predators and their environmental factors, such as soil conditions (Sanderson alternative plant diet was often reported (e.g. Stoner 1972; et al. 1995), pollution (Bra¨ndle et al. 2001) are also known Lattin 1989, 1999; Gillespie and McGregor 2000), so to influence true bug assemblages across the vegetation. vegetation diversity could also be an important factor for In accordance with many studies (e.g. Tilman 1986; predaceous true bugs in this manner. Siemann 1998; Siemann et al. 1998; Murdoch et al. 1972; According to the present study, the proportion of sur- Nagel 1979; Southwood et al. 1979; Prendergast et al. rounding habitat types has a significant impact on the 1993), our results also show that vegetation diversity has a species richness and composition of Heteroptera assem- positive effect on the number of species, but not on their blages. The surrounding landscape obviously influences the abundance (e.g. Tilman 1986; Siemann 1998; Siemann area of suitable habitats and presumably affects the ability et al. 1998; Perner et al. 2003, 2005). The findings of of bug species to move and disperse (e.g. Tillman et al. studies about correspondence between plant species rich- 2009; Gange and Llewellyn 1989). However, previous ness or diversity and density of arthropods are often con- studies (Zurbru¨gg and Frank 2006; Torma et al. 2010) troversial or ambiguous (Perner et al. 2005; Rhainds and suggested that the surrounding landscape had no significant English-Loeb 2003; Siemann et al. 1998). Our findings impact on the structure of Heteroptera assemblages. show that the vegetation has no direct effect on herbivore Although we found a significant effect of the surrounding abundance in general (Koricheva et al. 2000; Perner et al. landscape on the Heteroptera assemblages, it was partly 2005) but plant diversity can affect directly the abundance due to the correspondence between the surrounding land- of the more specialised or sessile herbivores (Koricheva scape and the land use type of the studied plots, i.e. the et al. 2000). The aggregated distribution observed in the proportion of forests was higher close to dikes due to the case of several specialist species and thereby their relation riparian forest of the River Tisza, as well as the proportion to a certain land use type was obviously due to the distri- of arable fields was obviously higher in the proximity of bution and density of their host plants. For example, the old fields, as it can be seen on the scatterplot of NMDS. high density of Lamiaceae in old fields resulted in the association of Platiplax salviae and Lasiacantha capucina with these sites. This pattern suggests a direct effect of Conclusions vegetation as the resource concentration hypothesis pre- dicted (Root 1973). On the other hand, the observed cas- Although we also found that both habitat quality, landscape cading effect of vegetation diversity up to predator trophic structure and land use type played a role in shaping the level also suggests the validity of the enemies hypothesis distribution of species similarly to findings of e.g. Jean- (Root 1973) and thereby the importance of top-down reg- neret et al. (2003), Biedermann et al. (2005), we emphasise ulation on herbivor abundance. Previous studies also sug- that it is the diversity and the composition of the vegetation gested that increasing plant diversity and thus increasing which can be the main factors for conservation the diver- herbivore diversity could potentially cascade up to higher sity of grassland insects in agricultural landscapes as trophic levels, leading to a greater diversity of parasites and determinants of keystone structures (Tews et al. 2004). predators, too (Hunter and Price 1992; Siemann 1998; Our present study also demonstrates the importance of Knops et al. 1999). However, we found that the vegetation dikes to conserve species diversity of true bugs in the coverage and the proportion of dead vegetation also agricultural landscape of Tisza River Basin as they increase affected the abundance and species richness of predaceous both the area and connectivity of grassland habitats. bugs significantly, which suggests the importance of Although the remaining grasslands of the lower Tisza- structural diversity of vegetation and not plant species valley similarly to other Hungarian grassland habitats (e.g. diversity per se. The structure or architectural complexity Bata´ry et al. 2007) still contain a diverse insect fauna, the of the vegetation is a significant influential factor for many negative influence of agriculture intensification is obvious. 123 J Insect Conserv

Grassland remnants need more attention in the sense of NKFP 6/013/2005 grant and the HU-RO/0901/205/2.2.2 project. land use and conservation and traditional land use types Further we thank the anonymous referee for the useful comments on an earlier version of the manuscript. should be given priority before agricultural changes can reduce the natural value of Hungarian grasslands.

Acknowledgments We are grateful to L. Ko¨rmo¨czi and M. Zalatnai Appendix for their help in sampling the vegetation and we also thank G. Laskay for the language correction. The present study was supported by See Table 4.

Table 4 The list of the collected Heteroptera species D1 D2 D3 D4 D5 D6 T1 T2 T3 T4

Tingidae Agramma atricapillum (Spinola, 1837) 0 0001 01000 Agramma confusum Puton, 1879 0 0001 00030 Derephysia foliacea (Falleń, 1807) 0 0010 00000 Dictyla echii (Schrank, 1781) 0 0002 00000 Dictyla humuli (Fabricius, 1794) 0 0 0 1 16 2 3 9 156 31 Catoplatus horvathi (Puton, 1879) 0 0200 00000 Corythucha ciliata (Say, 1832) 0 0100 10000 Kalama tricornis (Schrank, 1801) 0 0100 00000 Lasiacantha c. capucina (Germar, 1836) 0 0 0 24 0 0 0 556 6 15 Lasiacantha gracilis (Herrich-Scha¨ffer, 1830) 0 0000 000013 Oncochila scapularis (Fieber, 1844) 0 0000 00200 Oncochila simplex (Herrich-Scha¨ffer, 1830) 0 0000 00100 Stephanitis pyri (Fabricius, 1822) 0 0100 00000 Tingis (s. str.) auriculata (Costa, 1843) 0 0061 00000 Tingis (s. str.) cardui(Linnaeus, 1758) 0 0100 10003 Tingis (Neolasiotropis) pilosa Hummel, 1825 0 0020 00000 Miridae Acetropis carinata (Herrich-Scha¨ffer, 1842) 0 1803 00000 Acetropis longirostris (Puton, 1875) 271 0013 70010 Adelphocori ticinensis(Meyer-Du¨r, 1843) 0 0001 00000 Adelphocoris lineolatus (Goeze, 1778) 0 01281211001733 Adelphocoris seticornis (Fabricius, 1775) 0 0026 4 0 11817 Agnocoris rubicundus(Falleń, 1829) 0 0001 00000 Amlytylus nasutus (Kirschbaum, 1856) 0 0211 10000 Apolygus spinolai (Meyer-Du¨r, 1841) 0 0032 00040 Campylomma verbasci(Meyer-Du¨r, 1843) 0 0001 000720 Capsodes gothicus (Linnaeus, 1758) 1 0073 20010 Capsus ater (Linnaeus, 1758) 11 4052 00000 Chlamydatus pulicarius (Falleń, 1807) 0 0014 2 038026 Chlamydatus pullus Reuter, 1870 1 1101 5 0196420 Criocoris crassicornis (Hahn, 1834) 0 0110 00006 Deraeocoris ruber (Linnaeus, 1758) 0 0040 00002 Deraeocoris ventralis Reuter, 1904 0 0041 00111 Dicyphus geniculatus Fieber, 1858 0 0 0 2 18 500513 Europiella alpina(Reuter, 1875) 0 0002 00000 Europiella artemisiae (Becker, 1864) 0 0200 00000 Globiceps horvathi Reuter, 1912 0 0010 00000 Halticus apterus (Linnaeus, 1761) 2 0 3 55 71 50 1 15 51 144

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Table 4 continued D1 D2 D3 D4 D5 D6 T1 T2 T3 T4

Halticus luteicollis (Panzer, 1805) 0 0010 00000 Heterotoma meriopterum (Scopoli, 1763) 0 0000 00010 Leptoterna dolabrata (Linnaeus, 1758) 1 001534 0002 Liocoris tripustulatus (Fabricius, 1781) 0 0000 00020 Lygus gemellatus (Herrich-Scha¨ffer, 1835) 0 0001 00010 Lygus pratensis (Linnaeus, 1758) 0 0001 0 0 1181 Lygus rugulipennis Poppius, 1911 1 0012 40575 Megaloceroea reticornis (Geoffroy, 1785) 0 0 0 13 13 53 2 0 2 10 Megalocoleus molliculus (Falleń, 1829) 0 0020 0 1143 7 Myrmecoris gracilis (F. Sahlberg, 1848) 0 1011 00010 Notostira elongata (Geoffroy, 1785) 41 7 11 26 32 2 75 7 3 1 Omphallonotus quadriguttatus (Kirschbaum, 1856) 0 0000 01000 Orthocephalus saltator (Hahn, 1835) 0 0007180 13713 Orthocephalus vittipennis (Herrich-Scha¨ffer, 1835) 0 0001 10003 Orthops basalis (Costa, 1852) 0 0000 00001 Phytocoris insiquinis Reuter, 1846 0 0100 00000 Plagiognathus arbustorum Fabricius, 1794) 0 0001 00000 Plagiognathus chrysanthemi (Wolff, 1804) 0 0000 10000 Plagiognathus fulvipennis (Kirschbaum, 1856) 0 0080 00024 Polymerus brevicornis (Reuter, 1878) 0 0010 00000 Polymerus cognatus (Fieber, 1858) 0 0200 00000 Polymerus nigritus (Falleń, 1829) 0 0030 00000 Polymerus unifasciatus (Fabricius, 1794) 0 0008 10000 Polymerus vulneratus (Panzer, 1806) 0 0000 10001 Solenoxyphus fuscovenosus (Fieber, 1864) 0 0000 05000 Stenodema calcaratum (Falleń, 1807) 6 0 0 29 72 32 8 0 17 77 Stenotus binotatus (Fabricius, 1794) 1 0 0 2 25 122 0 0 31 50 Systellonotus triguttatus (Linnaeus, 1767) 0 0200 0 3371 0 Trigonotylus caelestialium (Kirkaldy, 1902) 1 0 1 21 24 3 49 50 6 4 Trigonotylus pulchellus (Hahn, 1834) 0 0042 0123660 2 Trigonotylus ruficornis (Geoffroy, 1785) 0 0001 00000 Anthocoridae Orius (Heterorius) majusculus (Reuter, 1879) 0 0000 10000 Orius (Heterorius) minutus (Linnaeus, 1758) 0 0026 300914 Orius (s. str.) niger Wolff, 1804 0 0 0 1 77 35 3 87 138 265 Temnostethus (Ectemnus) r. reduvinus (Herrich-Scha¨ffer, 1835) 0 0010 00000 Xylocoris (s. str.) cursitans (Falleń, 1807) 0 0000 10000 Nabidae Himacerus (s. str.) apterus (Fabricius, 1798) 0 0010 00000 Nabis (s. str.) ferus (Linnaeus, 1758) 0 0111 01200 Nabis (s. str.) p. pseudoferus Remane 1949 2 1123 4 730128 Nabis (s. str.) p. punctatus Costa, 1847 0 0121 23842 Berytidae Berytinus clavipes (Fabricius, 1775) 0 0100 00010 Piesmatidae Piesma kochiae (Beckegur, 1867) 0 0000 01000 Piesma maculatum (Laporte, 1832) 0 0000 00001 Lygaeidae sensu lato

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Table 4 continued D1 D2 D3 D4 D5 D6 T1 T2 T3 T4

Boesus quadripunctatus (Mu¨ller, 1766) 0 0040 00061 Cymus claviculus (Falleń, 1807) 0 0000 00010 Cymus glandicolor (Hahn, 1831) 0 0001 03010 Dimorphopterus spinolai (Signoret, 1857) 1 0000 00000 Emblethis griseus (Wolff, 1802) 0 0000 00100 Geocoris megacephalus (Rossi, 1790) 0 0001 00000 Graptopeltus lynceus (Fabricius, 1775) 1 0001 00000 Henestaris halophilus (Burmeister, 1835) 0 0000 0230 0 0 Ischnodemus sabuleti (Falleń, 1829) 4 0010 15001 Lygaeosoma anatolicum Seidenstu¨cker, 1960 0 0000 01000 Lygaeus simulans Deckert, 1985 0 0002 11000 Megalonotus chiragra (Fabricius, 1787) 0 0001 00100 Metotoplax origani (Kolenati, 1845) 0 0100 00012 Nysius senscionis (Schilling, 1829) 0 0010 10001 Ortholomus punctipennis (Herrich-Scha¨ffer, 1839) 0 2200 00126 Oxycarenus pallens (Herrich-Scha¨ffer, 1850) 0 0 0 0 22 4 1 0 13 7 Peritrechus gracilicornis (Puton, 1877) 0 8220 31101 Platyplax salviae (Schilling, 1829) 0 0 0 27 1 0 0 10 8 19 Pterotmetus staphyliniformis (Schilling, 1829) 0 0100 0 3180 0 Raglius alboacuminatus (Goeze, 1778) 1 0000 00010 Scolopostethus grandis Horva´th, 1880 0 0001 00000 Trapezonotus arenarius (Linnaeus, 1758) 1 0000 00000 Tropistethus holosericeus (Scholtz, 1846) 0 1000 00000 Xanthochilus quadratus (Fabricius, 1798) 0 1000 00000 Coreidae Bathysolen nubilis (Falleń, 1807) 0 0000 00100 Ceraleptus gracilicornis (Herrich-Scha¨ffer, 1835) 1 0004 11001 Coreus marginatus (Linnaeus, 1758) 0 0 0 18 1 20000 Syromastes rhombeus (Linnaeus, 1767) 0 0010 00001 Alydidae Alydus calcaratus (Linnaeus, 1758) 0 0002 20359 Camptopus lateralis (Germar, 1817) 0 0000 01001 Rhopalidae Brachycarenus tigrinus (Schilling, 1817) 0 0000 00011 Chorosoma schillingi (Schummel, 1829) 0 1200 0100 0 0 Liorhyssus hyalinus (Fabricius, 1794) 0 1000 00001 Myrmus miriformis (Falleń, 1807) 52 9 32 33 11 2 62 3 9 1 Rhopalus conspersus (Fieber, 1837) 0 0000 00010 Rhopalus parumpunctatus (Schilling, 1817) 0 0021 00022 Rhopalus subrufus (Gmelin, 1788) 0 0 0 11 1 20001 Stictopleurus abutilon (Rossi, 1790) 0 0011 02010 Stictopleurus pictus Fieber, 1861 0 0001 12000 Stictopleurus punctatonervosus (Goeze, 1778) 0 0029170 01122 Plataspidae Coptosoma mucronatum Seidenstu¨cker, 1963 0 0000 10000 Coptosoma scutellatum (Geoffroy, 1785) 0 0050 30001 Thyreocoridae Thyreocoris scaraboides (Linnaeus, 1758) 0 0030 00000

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Table 4 continued D1 D2 D3 D4 D5 D6 T1 T2 T3 T4

Cydnidae Canthophorus melanopterus (Herrich-Scha¨ffer, 1835) 0 0000 10000 Tritomegas bicolor (Linnaeus, 1758) 00005 00000 Scutellaridae Eurygaster maura (Linnaeus, 1758) 1 0 0 36 1 02012 Eurygaster testudinaria (Geoffroy, 1785) 0 0011 10001 Psacastha neglecta (Herrich-Scha¨ffer, 1837) 0 0020 00000 Pentatomidae Aelia acuminata (Linnaeus, 1758) 3 12 12 63 2 79315 Aelia rostrata Boheman, 1852 0 8830 01000 Antheminia lunulata (Goeze, 1778) 0 0000 00400 Carpocoris mediterraneus Tamanini, 1958 0 0000 10000 Carpocoris purpureipennis (De Geer, 1773) 1 0030 13215 Dolycoris baccarum (Linnaeus, 1758) 1 0 0 11 2 10417 Eurydema oleracceum (Linnaeus, 1758) 0 0085 00052 Eurydema ornatum (Linnaeus, 1758) 0 0020 00000 Eusarcoris aeneus (Scopoli, 1763) 0 0 0 12 13 20036 Eusarcoris ventralis (Westwood, 1837) 0 0031 00000 Graphosoma lineolatum (Linnaeus, 1758) 0 0020 00000 Holcostethus vernalis (Wolff, 1804) 0 0030 00000 Neottiglosa leporina (Herrich-Scha¨ffer, 1830) 4 8150 09000 Neottiglosa pusilla (Gmelin, 1789) 0 0030 00000 Rubiconia intermedia (Wolff, 1811) 0 0011 00001 Sciocoris cursitans (Fabricius, 1794) 0 0000 0 1140 0 Sciocoris distinctus Fieber, 1851 3 0210 01000 Sciocoris sulcatus Fieber, 1851 0 0100 01000 Stagnomus amoenus (Brulle´, 1832) 0 0010 00000 Vilpianus galii (Wolff, 1802) 0 0100 00000 Zicrona coerulea (Linnaeus, 1758) 0 0001 00000

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