Solutions for Biodiversity in Organic agriculture?

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Term Paper in Landscape Ecology, 2008 Nils Harley Boisen

Term Paper

LAØ370; Landscape Ecology, 2008 Department of Landscape Architecture and Spatial Planning, Norwegian University of Life Sciences

Agricultural Intensification and Landscape Homogeneity: omogeneity: Solutions for Biodiversity B y in Organic Agriculture?

Nils Harley Boisen Modern agricultural intensification constitutes one of biodiversity’s biodiversity’s greatest current threats. Increased agricultural intensity intensity comes at the cost of heightened landscape homogeneity and fragmentation on a range of spatial scales. Developed eveloped countries in temperate regions experienced their greatest rate of land use change by the industrial agricultural revolution prior to the 1950’s, temporally coinciding with their strongest decline of appurtenant species populations. Sustainable agricultural systems, namely organic farming, are looked upon by many as a potential solution to conventional agriculture’s continued impoverishment of landscape complexity. For the purpose of elucidating whether organic agriculture demonstrates demonstrate benefits for biodiversity, this paper evaluates a pool of comparative scientific studies which have addressed the potential biodiversity divergences ivergences between arable rable organic and conventional agriculture.

The results are discussed in light of certain essential facets of landscape ecology as well as the conclusions of three other meta-analyses meta on the same topic.

Key words: Agriculture, Biodiversity, Biodiversity Heterogeneity, Organic, Landscape ecology

Introduction Agriculture has since the start of industrialization adapted pted to methods and techniques which contribute to polarizing agricultural landscapes either into nonnon productive fallow or homogonous monocultures mono (Hendrickx et al. 2007; Reidsma et al. 2006; Stoate et al.

2001). Increased agricultural intensity has come at the cost of heightened landscape homogeneity and fragmentation on a range of spatial scales (Hendrickx et al. 2007; Reidsma et al. 2006; Stoate et al. 2001).

2001) Developed countries in temperate regions experienced their greatest rate of land use change by the industrial agricultural revolution prior to the 1950’s,, temporally coinciding coi with their strongest decline of appurtenant species populations (WWF 2006). With declines continuing, the adverse a effects of agricultural intensification ation on biodiversity has become an area of growing concern (e.g. Belfrage et al.

2005; Clough et al. 2005; Hendrickx et al. 2007; Krebs et

al. 1999; Reidsma et al. 2006; Vandermeer & Perfecto 2007). Given that the global demand for agricultural products is expected to increase by 50% within 2030 (IUCN 2006a), there re will be a challenging necessity for agriculture to reconcile with constructive ecological knowledge concerning how it can to the greatest possible degree sustainably conserve biodiversity. biodiversity Sustainable ustainable agriculture can be defined as technology and methodology ology attempting to maximize production and human health while minimizing environmental encroachment and degradation (Pretty & Hine 2001). This form of agriculture utilizes natural assets and must be locally adapted as it is intrinsically dependant on local ecological conditions. With clearly defined recommended and restricted practices, organic agriculture (OA) seeks to conserve the health of the environment and increase food security (Scialabba & Williamson 2004) through practices which are a combination of refined old school wisdom and

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Term Paper in Landscape Ecology, 2008 Nils Harley Boisen

contemporary ecological innovation. These practices are based on an ideology that are underpinned by principals of health, ecology, fairness, and care (IFOAM 2007). Because of these distinctive characteristics OA is as close as modern agricultural management systems come to sustainable agriculture.

The maintenance of non-crop vegetation among and adjacent to crops, the integration of natural habitat as buffers and/or corridors, and the lack of synthetic agrochemicals including pesticides and fertilizers are all aspects of organic management that are recognized by many world organisations, including IUCN (2006b) and FAO (2007), as having value in biodiversity conservation.

These aspects, ever disappearing under the regime of intensive industrial agriculture, are necessary for maintaining spatial heterogeneity at different scales, a pillar of species richness and diversity (Benton et al. 2003).

This paper investigates if the published literature from species diversity studies on OA vs. conventional agriculture (CA) indicates that landscape under OA management better facilitates different aspects of species diversity than conventional agriculture.

To attempt to explain the distribution of results and their explanations from studies of this character, one is disposed to view them from a landscape ecological angle.

Landscape ecology is well defined by Risser et al (1984) as the study of the development, management, and ecological consequences of spatial heterogeneity, asking with particular relevance to this topic how landscape heterogeneity interacts with fluxes of organisms, material, and energy. According to Wu and Hobbs (2007) this is typical of the early North American school of thought in landscape ecology, notably influenced by the theory of island biogeography (MacArthur & Wilson 1967) and patch dynamics (Levin & Paine 1974). Thus this study further investigates if the results of these species diversity studies on OA vs. CA farms are explained by landscape ecological aspects which appear to be common for OA.

II: mixed, i.e. biodiversity parameter significantly associated with ecological explanations attributed with either OA and CA sites by the particular study / III: insignificant, i.e. biodiversity parameter significantly associated with ecological explanations not specifically attributed to either OA or CA sites by the particular study/ IV: negative, i.e. biodiversity parameter significantly associated with ecological explanations specifically attributed to CA site by the particular study. It is important to keep the definitions of the result categories clearly in mind, e.g. one may easily forget how this study describes insignificant results with regard to biodiversity associations with farm management as significantly explained by factors not specifically attributed to OA or CA sites.

Categories of ecological explanation of result parameters were: I: vegetative structural heterogeneity, i.e. degree of structural complexity rendered from vegetation in a given plot/ II: lack of pesticide/herbicide use / III: habitat patch diversity, i.e. the abundance and composition of different non crop habitats on the given farm/ IV: degree of surrounding landscape heterogeneity, i.e. complexity of landscape surrounding the given farm. Categorization of species groups was chiefly dependant on which species the studies focused on, which generally speaking was a combination of taxa and niche; I: predatory epigeal arthropods / II: insect pollinators / III: other insect groups / IV: plants / V: birds / VI: bats (Microchiroptera spp..).

Being that a specific biodiversity parameter could have more than one ecological explanation (e.g. activity explained by both vegetative structural heterogeneity and habitat patch diversity), the total number of parameter results used in this meta-analysis originate from the 26 research publications is 123. These have been plugged into data tables in Microsoft office Excel 2007 to produce numerical- and percent wise visual depictions of the data’s distribution. In addition, three other meta-analyses on the same topic are discussed in light of this study’s findings.

Results

Methods This meta-analysis has gathered results from 26 research publications comparing different biodiversity parameters (i.e. aspects of species diversity (e.g. abundance, density, activity, richness, comosition) with ecological explanations) in arable OA vs. CA from Europe (and one study from Canada). The authors of these studies have conducted their work on the premise that differences in biodiversity parameters should be conspicuous in a scientific context based on the clear differences in farm management with logical implications for biodiversity.

Research results were divided into categories of result type, ecological explanation, and species group. Result categories were; I: significant, i.e. biodiversity parameter significantly associated with ecological explanations specifically attributed to OA sites by the particular study /

The number of research publications which address each species group is displayed in figure 1. In total (ca. 68 %, figure 2.1), and separated by ecological explanation, the greatest majority of parameter results were significant, while mixed and insignificant results tended to tie in second place, with negative results nearly nonexistent (figure2). Habitat patch diversity explained the largest amount of significant parameter results (figure 2.D) followed by degree of surrounding landscape heterogeneity (figure 2.E), followed by lack of pesticide/herbicide use (figure 2.C), and finally vegetative structural heterogeneity (figure 2.B). The few negative parameter results were explained by lack of pesticide/herbicide use (figure 2.C), habitat patch diversity (figure 2.D), and degree of surrounding landscape heterogeneity (figure 2.E). With the exception of predatory epigeal arthropods (figure 3.A) and

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Term Paper in Landscape Ecology, 2008 Nils Harley Boisen

other insect groups (figure 3.C), all ll species groups displayed a large majority of significant parameter results. results Plants and insect pollinators were respectively the two species groups with the largest amount of significant parameter results (figure 3B & 3.D).

Predatory Epigeal Arthropods Predatory epigeal arthropods displayed the most even distribution of result type, with significant significan and insignificant equally common. The two best explanations for significant results with predatory predat epigeal arthropods was vegetative structural heterogeneity and habitat patch diversity respectively, while the two best explanations for insignificant results were the degree of surrounding landscape heterogeneity and habitat patch diversity (figure 4).

The only negative result for predatory epigeal arthropods was explained by the degree of surrounding landscape heterogeneity (figure ( 4).

In Cloughs et.al (Clough et al. 2007a) study on alpha (ἄ) ( and beta (β) diversity of arthropods and plants in OA vs. CA wheat fields, β diversity of spiders was larger in conventional fields than in organic fields, and was attributed partly pa to the fact that the species composition of spiders varies more strongly with landscape composition than in organic fields.

The species composition of only one group, rove beetles (staphylinidae staphylinidae spp.), were shown to be affected by local habitat factors factor when habitat characteristics were analysed (Clough et al. 2007a). The habitat variability between betw CA fields

is lower than that in OA fields, thus in this study it seemed to Clough et al. (2007a) that the limiting factor for spiders and rove beetles was habitat to overwinter in outside arable fields due to tillage in addition to habitat patch diversity within the fields. Clough et.al (2007a) states further that it has been shown that more intensive management can result in greater heterogeneity between plots and thus increase β-diversity, diversity, albeit on a temporal and not on a spatial scale. In an earlier study addressing spider diversity in relation to this topic at a local, local landscape and regional scale, Clough et al. (2005) found that region and local management exerted no effect on species richness, and that spider activity density was higher in field edges though differed among regions. Farmland spider diversity was influenced nced by spatial differences on two scales; edge vs. centre and simple vs. complex landscapes, but not at the he two others; field management and region, which according to the authors emphasized the importance of addressing various spatial scales to adequately adequatel explain patterns of biodiversity (Clough et al. 2005).

2005) Promoting heterogeneity in land use at landscape scales was understood as one of the keys to promoting spider diversity in agro-ecosystems, ecosystems, and not specific management styles

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Term Paper in Landscape Ecology, 2008 Nils Harley Boisen which have their main distribution distrib in semi-natural habitats occurred more abundantly in OA fields, indicating a correlation between organic farming and semi-natural natural habitats. Thus Pfiffner & Luka (2003) state that seminatural habitats in combination with OA may be a practical way to conserve and enhancement species rich assemblages on agricultural land. In Puratauf et al.’s (2005) study of carabids however, the main conclusion was that surrounding landscape features were much more important than OA for enhancing local biodiversity. Here, with landscape complexity demonstrated ted to be independent of management system, carabid activity density and species richness augmented with percent grassland cover in the surrounding landscape (Purtauf et al. 2005).

(i.e. OA) (Clough et al. 2005).. Schmidt et al.’s (2005) study on epigeal spiders revealed that organic agriculture did not ot increase spider species richness, but did however enhanced spider density by 62%. Otherwise, increased non-crop crop habitats in the landscape increased local species richness of spiders from 12 to 20 species, irrespective of local management, which accordingg to Schmidt et al.

(2005) indicates that larger species pools are sustained in complex landscapes with a higher her availability of refuge and overwintering habitats.

In Pfiffner & Luka’s (2003) study, carabid species richness between integrated crop management (ICM) fields and OA fields was in most cases highest in OA fields. Spider species richness displayed mixed result between ICM and OA fields (Pfiffner Pfiffner & Luka 2003).

2003) Carabid and spider activity density was higher in the OA fields as opposed to the ICM fields. Endangered species, and species with a narrow ecological preference were more abundant in OA fields (Pfiffner & Luka 2003).

2003) Weed abundance in OA was demonstated onstated to be the greatest factor influencing carabid fauna, and weed diversity in OA the greatest factor influencing spider fauna (Pfiffner & Luka 2003). Furthermore, according to Pfiffner & Luka (2003) many species of carabids and wolf spiders (Lycosidae ( spp.)

Insect Pollinators For insect pollinators the three chief explanations for significant results were habitat hab patch diversity, degree of surrounding landscape heterogeneity, and lack of pesticide/herbicide use (figure figure 5). Returning to Clough et al.’s study from 2007a, a greater regional β diversity, as well as site ἄ and β diversity of wild bees was associated with the positive response to OA demonstrated by flowering plants (discussed further down). In part due to lack of herbicide application in OA., CA managed fields in comparison contain few broad-leaved broad flowering plant species which can attract bees (Clough et al. 2007a). In addressing bumblebee (Bombus Bombus spp.) abundance and species richness, Rundolf et al.’s (2008) study matched pairs of OA and CA farms located in homogenous and heterogeneous landscapes. Here, abundance and species richness were significantly gnificantly positively related OA and surrounding landscape heterogeneity (Rundlof et al. 2008).

However, significantly higher species richness richn and abundance was only associated with OA farms located in homogeneous landscapes (Rundlof et al. 2008).

Furthermore, this influence of landscape context on bumble bee abundance was most relevant for species with medium sized colonies (Rundlof et al. 2008).

2008) The greater

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ee abundance OA farms was partially correlated to bumblebee higher flower abundance on these sites (Rundlof et al.

2008). As colony size is believed to reflect the spatial scale at which bumblebees make use of foraging resources, the foraging ranges of species with medium sized colonies are most negatively affected by fragmentation of foraging habitat (Rundlof et al. 2008).. An earlier study by Rundlof and Smith (2006) addressed butterfly abundance and diversity also using matched pairs of OA and CA farms located in homogenous and heterogeneous landscapes.

Here as well abundance and species diversity was shown to bee significantly correlated with both OA and surrounding landscape heterogeneity (Rundlof & Smith 2006).

2006) Moreover, like in Rundolf et al’s (2008) later bumblebee study, the positive influence on abundance and species specie richness was most evident in intensively farmed homogeneous landscapes (Rundlof & Smith 2006).

2006) Ekroos et al (2007) looked at the role of organic and conventional entional field boundaries for bumblebees and butterflies and found no significant effect of OA field boundaries on butterfly diversity, but did however see higher (though not significant) abundance and species richness of bumblebees. Ekroos et al. (2007) interprets this as the effects of organic farming possibly being overpowered by the influences es of landscape structure in heterogeneous landscapes, as the abundance of nectar flowers clearly was an important factor explaining butterfly diversity and bumblebees are likely to benefit from higher weed abundance and species diversity in the fields, both being factors otherwise associated with OA from other studies (e.g. Hald 1999; Hyvönen et al. 2003).

2003) Both bumblebee and butterfly abundance were also investigated by Belfrage et al. (2005),, though here h the paired OA and CA farms consisted of pairs of large and small farms. Butterfly numbers were 40% higher on smaller compared to larger farms, and 65% higher on OA farms compared to CA farms (Belfrage et al. 2005).

2005) Bumblebee numberss were 40% higher on smaller compared to larger farms, and 13 % higher on OA farms compared to CA farms, though the latter was not

considered significant. Thus, the greatest differences in butterfly and bumblebee abundances was between small OA and large CA farms (Belfrage et al. 2005).

2005) The greatest difference in landscape between the small and large farms in this study was higher crop species diversity per hectare as well and smaller field sizes. In Feber et al.’s (2007) study addressing butterfly abundance in relation to this topic, OA farms overall attracted significantly more butterflies (Lepidoptera spp..) with greater species diversity than CA counterparts, with significantly more samples recorded over the uncropped field margin m than the crop edge. Moreover, the difference in butterfly abundance between crop edge and field margin was greater in CA than OA systems (Feber et al. 2007).

2007) Organic farms in this study possessed proportionally more grass leys and larger larg hedgerows, which is Feber et al.’s (2007) best explanation for differences in butterfly abundance and richness between the two farm types.

Other Insect Groups Stemming from only two studies (mainly one), the few parameter results (significant and mixed) for the species group of other insect groups were lack of pesticide/herbicide use, habitat patch diversity, and degree of surrounding landscape heterogeneity (figure 6).

Wickramasinghe et al. (2004) examined moth (Lepidoptera spp.)) diversity and abundance, in addition to the general abundance of other nocturnal insects, and specifically the abundance of 18 nocturnal insect families of importance to bat foraging. This was for the purpose purpo of quantifying available bat prey in relation to a previous study by Wickramasinghe et al. (2003) (discussed further down).

Compared to the CA farms in the study, nocturnal insect abundance, species richness, and moth species specie diversity were significantly higher on the OA farms (Wickramasinghe et al. 2004).

2004) Moreover, insect abundance in OA pastoral and water habitats was significantly higher than analogous logous habitats on CA farms. Furthermore in regard to the above mentioned 18 insect families, only 5 were significantly more abundant on OA farms, and here

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as well most highly correlated with OA pastoral, water, as well as woodland habitats. (Wickramasinghe et al. 2004).

Wickramasinghe et al. (2004) states that nearly all of these sites were surrounded by trees or bushes, which likely provide shelter for emerging aquatic insects and dead leaf beds that are essential habitats for other insect groups. This fact, in addition to the lack of herbicide application is indicated as the best explanation for the results. Species richness of insect herbivores (belonging to Coleopteran, Dipteral, Hymenoptera, and Lepidoptera) of non-crop plants (Canada thistle; Cirsium arvense) was addressed by Clough et al. (2007b). OA and surrounding landscape heterogeneity, but not higher densities of thistles in the landscape was shown to enhance insect herbivore species richness (Clough et al. 2007b). Host-plants located in OA fields were extra prone colonization than those in the CA fields (Clough et al. 2007b).

Plants In regards to plants the significant results were chiefly explained by habitat patch diversity, with lack of pesticide/herbicide use and degree of surrounding landscape heterogeneity more or less tied as the secondary explanation (figure 7). The four negative results for plants were equally attributed to lack of pesticide/herbicide use and habitat patch diversity (figure 8).

Once again returning to Clough et al.’s (2007a) study, site ἄ diversity of plants was greater on OA farms, and was attributed to the more extensive (heterogeneous) type of farm management in OA, as well as the lack of herbicide application and greater insect pollinator abundance. However, as mentioned earlier one can wonder if the latter is partially the result of greater flowering plant diversity. The species richness of plants which utilise insect pollination was shown in Gabriel & Tscharntke’s (2007) study addressing if insect pollinated plants benefit from OA to be much higher in OA than CA fields. This trend was even more significant in field edges rather than centres (Gabriel & Tscharntke 2007). Conversely, the abundance of non-insect pollinated plants showed an opposite trend, being more numerous in CA farms and to an increasing degree at CA field edges (Gabriel & Tscharntke 2007). Gabriel & Tscharntke (2007) attributed these results to greater pollinator densities in organic fields as well. Thus as agricultural intensification implies a disruption of plant-pollinator interactions, this may cause important shifts in plant community structure (Gabriel & Tscharntke 2007).

Returning to Belfrage et al.’s (2005) study, in addressing herbaceous plant species, 89 % more species were recorded on small compared to large farms and 75% more on OA compared to CA farms. A separate study by Gabriel et al. (2006) addressing the diversity of plant communities at different spatial scales found that that ἄ, β, and γ diversity were higher at all scales in OA compared to

CA fields, and like their previously mentioned study (Gabriel & Tscharntke 2007) this trend was more significant at the field edge compared to the centre. β diversity at the meso and macro scale explain the majority of species richness which according to Gabriel et al. (2006) indicated substantial environmental heterogeneity among fields and regions (Gabriel et al. 2006). This was particularly true for rare plant species on OA farms, while the richness of common species independent of farming system was explained by β diversity at the micro and meso scale (among plots and fields) (Gabriel et al. 2006). Aude et al. (2004) looked at the conservation value of herbaceous vegetation in hedgerows. Here, OA hedgerows contained significantly more plant species than CA equivalents, and the species composition was most significantly similar to seminatural communities (Aude et al. 2004). Aude et al. (2004) thus concluded that that OA is better for conserving herbaceous plant diversity in hedgerows given otherwise intensively cultivated surrounding landscapes. The OA and CA farms in this study were situated analogous landscapes regarding connectivity, fragmentation, etc. Thus the authors explain the semi-natural plant species composition of the OA hedgerows as likely being the effect lacking of chemical fertilizer application in adjacent OA fields, as eutrophication of field boundaries is shown to result in smaller stress-tolerant species and dominance of taller nutrient avaricious ruderals (Aude et al. 2004). With surrounding landscape parameters being similar between the OA and CA study pairs, the authors speculate that the greater species richness of the OA hedgerows was possibly due to greater ornithochory in these sites, as OA in many circumstances has been shown to be favourable for bird fauna (as discussed further down) (Aude et al. 2004). The reduced species richness in the CA hedgerows was chiefly attributed to herbicide application though there was no direct evidence for this (Aude et al. 2004). Petersen et al.’s (2006) study revealed that nutrient greedy annual ruderal plant species as well as annual species otherwise common to nutrient rich conditions were most prevalent in CA field borders, whereas perennial stress-tolerant species were more numerous around OA boarders. That stress-tolerant species could be indicators of organic farming were explained by the authors of this study as likely being the result of such species possessing a competitive edge over ruderals in biotopes with low nutrient input (low productivity) as well as disturbance (Petersen et al. 2006).

In addition, lack of herbicide application in OA is also given as a likely explanation (Petersen et al. 2006). These factors could also explain how the differences recorded between OA and CA in this study were only visible 3–4 years after conversion to organic farming since the edge habitats would need time to readapt to new conditions (Petersen et al. 2006). Romero et al.’s (2007) study which addressed weed diversity in inner fields and crop edges,

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here too it appeared that OA enhanced total weed diversity, however in this particular study the concentration of diversity in the crop edge was much lower in OA fields.

The greater total weed diversity in OA stemmed from a greater diversity of broad-leaved leaved species in the inner fields, and was attributed to CA’s thicker cover of herbicideherbicide treated and heavily fertilized arable fields (Romero et al.

2007).

Plant diversity in Gibson et al.’s (2007) study found that at the whole farm scale OA farms possessed significantly higher plant species richness than CA counterparts, withh OA fields also demonstrating higher plant abundance. The authors note this entails a greater number of species per metre in OA arable fields despite the lower fraction of arable land on OA farms (Gibson et al.

2007). The study also showed that OA farms contained in total more areas of semi-natural natural habitat (woodland, field margins and hedgerows combined) (Gibson et al. 2007).

2007) Moreover, woodland area itself on OA farms was also significantly ificantly greater, with more unbroken pockets of woodland while woodland on CA farms often consisted of more linear patches (Gibson et al. 2007).. However, within the patches of semi-natural natural habitat there was no significant signifi differences in plant species richness between OA and CA farms, though high species richness did approach significance in conventional woodlands and organic hedgerows (Gibson et al. 2007).. Gabriel et al. (2007) attributed their significant results for plant species richness and abundance to lack of herbicide / chemical fertilizer application, as well as to spillover from the increased degree of semi-natural natural habitat on the OA farms, while whi the greater amount of semi-natural natural habitat was only explained by the lack of herbicide / chemical fertilizer application.

Another study looking at plant species richness in contrasting landscapes by Boutin et al (2008) indeed found greater species richness in OA fields and hedgerows.

Though there was no apparent differences in species composition on between the two farming types, many plants however were recorded only on OA farms, among these several of conservation interest such herbaceous forest

perennials (Boutin et al. 2008).

2008) An examination of the landscape surrounding the study sites showed that OA hedgerows were situated in a matrix comprised more of non-crop habitats, in particular icular old-field old patches and to a slighter degree forest patches. According to Boutin et al (2008) the significant results of this study could be explained by the increased heterogeneity of the surrounding landscape andscape which worked synergistically with OA management.

Looking at the effects of landscape complexity on arable weed species diversity Roschewitz Ro et al. (2005) uncovered that the γ diversity of weed vegetation at the field scale, the was higher in OA than in CA fields, and that the greater γ diversity of OA fields was just weakly associated with greater landscape complexity. Moreover, γ and ἄ diversities in the seed rain and seed bank were generally higher in organic than in conventional fields (Roschewitz et al. 2005).. This significant trend of greater γ diversity for OA farms was principally valid in fields located in simple landscapes dscapes with an elevated proportion of arable land, with γ diversity in conventional fields strongly linked with landscape complexity (Roschewitz et al. 2005).

Thus in complex landscapes with large proportions of nonnon crop areas the γ diversities of OA and CA fields were nearly similar (Roschewitz et al. 2005).

2005) The authors interpreted this as a suggestion that OA fields are more or less self-sufficient sufficient ecological units, independent from species immigration migration from surrounding habitats in complex landscapes (Roschewitz schewitz et al. 2005).

2005) Furthermore, the same pattern of γ diversity was revealed for ἄ and β diversities (Roschewitz et al. 2005).

2005) However, though landscape complexity did not appear to influence the weed species diversity of OA fields straight away, the potential to do so seemed apparent in the following followi years by looking at the more diverse germinable seed bank in complex landscapes (Roschewitz et al. 2005).

2005) When separating the β and gamma diversities into broad-leafs leafs and grasses, grasses did not differ between farming systems, though were negatively correlated with the proportion of arable land (Roschewitz et al. 2005).

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Roschewitz et al. (2005) believed that broad-leaved species were less numerous on CA fields since these are more susceptible to herbicide application. There was otherwise no difference in the quantity of red-list species between OA and CA farms, though the number of species was shown to be higher in complex as opposed to simple landscapes, demonstrating the value of alternative habitats (Roschewitz et al. 2005).

Birds For birds, the significant results were chiefly explained by degree of surrounding landscape heterogeneity, followed by habitat patch diversity and vegetative structural heterogeneity, respectively (figure 8).

Once again returning to Belfrage et al.’s (2005) study, this time addressing bird fauna, total species richness was 55% higher and the number of bird territories 78% higher on OA farms. This trend increased in importance from large CA farms, to large OA farms, to small CA farms, to small OA farms (Belfrage et al. 2005).

However, in Chamberlain et al.’s (1999) study there was no significant difference in bird species diversity between OA and Ca farms during the breeding season, yet OA farms contained consistently elevated densities of specific species as well all species combined than CA counterparts outside the breeding season. Furthermore, no one bird species was more abundant on CA farms, while the species displaying greater densities on OA farms were principally recorded in the field boundaries (Chamberlain et al. 1999).

Hedges in the OA field boundaries of this study were higher and wider with more trees, as well as field sizes being smaller, and an analysis pointed to the importance of habitat structure in explaining the divergence in bird density between farm types (Chamberlain et al. 1999).

Furthermore, the stronger trend in bird densities outside the breeding season was explained by the authors as likely associated with seasonal variation in territorial behaviour, with the strongest individuals constraining the habitat use of less dominant individuals thus forcing these into suboptimal habitats (Chamberlain et al. 1999). In Freemark et al.’s (2001) study from Ontario (the only study in this meta analysis not from Europe) addressing bird fauna, species richness and total abundance was superior on OA farms.

According to the authors analyses differences in farm type practices explained 23.7%, habitat 26%, and habitat and practices together 5.7%. The gradient of bird species richness and abundance appeared to follow the gradient of habitat heterogeneity implied by more pasture, winter grain, and non-crop habitats (hedgerow, woodland) to sites with few bird species associated with larger fields with more rowcroping and spring grain, passes and tilling, and use of herbicides and chemical fertilizers (Freemark & Kirk 2001).

Bats

Significant results for bats were equally explained by habitat patch diversity and vegetative structural heterogeneity (figure 9). Though bats as a species group is represented by only one study in this meta-analysis, the results of this particular study have such wide reaching implications that it seems only pertinent to include them. In this study, bat activity was quantified using acoustic surveys which made it possible to identify 89 % of bat passes to species, and a further 9% to genus, as well as quantify foraging activity by tallying total feedings buzzes as well as feeding buzzes per pass. OA farms displayed 61% higher bat activity of all recorded species, including far more passes over “organic” water habitats when habitats were viewed separately (Wickramasinghe et al.

2003). Foraging activity was 84% higher on organic farms, with the number of feeding buzzes per pass also significantly higher over the organic farms and also correlating to hedgerow height (Wickramasinghe et al.

2003). The species richness of bats was however not significantly different between the two management forms with 14 of the sixteen documented species being observed over the organic farms and 11 over the conventional (Wickramasinghe et al. 2003). Looking more closely at the life history traits of the bat species sampled by Wickramasinghe et al. (2003), it is observed that those which are more specialized and therefore sensitive to habitat degradation were all species which had the highest bat activity over the organic farms (Wickramasinghe et al.

2003). The authors attribute this to taller hedgerows and improved water quality on the OA farms. The uncommonly substantial results of Wickramasinghe et al.’s (2003) bat study suggests that given otherwise similar landscape features, the habitat structure and quality found on these particular organic farms were more suitable not necessarily for the diversity of bats in southern England, but certainly their foraging and general activity, which of course will influence bat population demography in many complex ways. The addition of higher foraging activity over the organic farms also indicates greater habitat quality for prey insects and if we recall Wickramasinghe et al.’s (2004) study on nocturnal insects, it was stated here that the activity of bats that mainly ate Lepidoptera was significantly correlated with the abundance of this order.

Discussion The theory of island biogeography has directed much of its focus on the influences of habitat fragmentation and isolation on population dynamics (Collinge 1996; Wu & Hobbs 2007), and thus appears particularly applicable for studies located in intense modern agricultural landscapes.

It is easy to depict an arable agricultural landscape as a “sea” of homogenous inhospitable matrix, potentially containing an archipelago of more or less favourable noncrop habitats of differing sizes, and distances from each other. In this meta-analysis a greater amount of habitat

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(patches) with higher quality (partitioned in this study to vegetative structural heterogeneity and habitat patch diversity) was intrinsic with nearly every OA farm in comparison to CA counterparts either by direct measurements or as implied by authors when such pairs were located in otherwise homogenous landscapes.

Therefore, the concept of metapopulation dynamics can be a key component in explaining the significant, mixed, and insignificant results of this meta-analysis. Freckleton and Watkinson (2002) in reference to Hanski (1997) state in general that four conditions must be fulfilled for a metapopulation to persist; i: “suitable habitat occurs in discrete patches that may be occupied by local reproducing populations”/ ii: “even the largest local populations have a measurable risk of extinction (unless the largest population is the source of a source-sink system)”/ iii: “habitat patches must not be too isolated to prevent recolonization following local extinctions” iv: “local populations do not have completely synchronous dynamics”.

This dispersal is a key process for meta-population persistence through colonization of empty patches and supplementation of occupied patches (Rabasa et al. 2007).

According to Van Dyck & Baguette (2005) dispersal is a process consisting of emigration (crossing habitat boundaries), landscape traversal (including resource-poor matrix) and settlement. In such cases dispersal is initially realized as a by-product of a search for daily resources by routine movements, or by special movements intended to result in displacement and settlement at a distance from the previous or natal site. Vuilleumier et al. (2007) states that colonization of habitat patches is more likely to occur between adjacent patches, and accordingly the spatial configuration of patches in the matrix affects the capability of a given species to disperse. Thus the amount of, and distance between, habitat patches amidst the matrix (in this case agricultural fields) defines their degree of dispersal facilitation by acting as “stepping stones”.

While some few thistles in a large field of grain could be stepping stones between each other or two pockets of edge habitat for a given species of parasitic wasp (Apocrita sp.), a given bird species such as quail (Bonasa bonasia) would likely require that such isolated patches in the field be far more substantial and frequent for it to brave predation dangers the matrix between.

Conversely magpies (Pica pica) would not need the stepping stones at all. This exemplifies that what also controls the degree of dispersal for individuals in a fragmented landscape is scale dependant; i.e. the accessibility of habitat patches is related to the individual species dispersal capabilities (Turner 2005), and as stated above it appears that the OA farms of this meta-analysis also contain greater patchiness possessing offering different degree of qualities for the individual species groups. However, stepping stone habitat is only one side of the connectivity coin. The other side is linear “corridor”

habitat. Such linear habitats are presumptively believed to be important for facilitating dispersal for individual species depending on their landscape context, habitat type, scale, and nature of the connected patches (MacDonald 2003).

Fewer than half of the studies gave persuasive results in a meta-analysis by Beier & Noss (1998) on the effectiveness of dispersal facilitation by corridors. Nevertheless the authors state that evidence from the well designed studies suggest that corridors in fact are very effective (Beier & Noss 1998). A study by Pocock & Jennings (2007) for example found that shrews (Sorex spp.) and bats were strongly sensitive to boundary loss. Again, in looking at the OA farms of this meta-analysis, the greater majority of them are described as possessing a higher proportion and /quality of non-crop field edges which are likely to facilitate dispersal for a greater array of species depending on their qualities.

According to Rabasa et al. (2007), many studies demonstrate that habitat quality is possibly of greater importance than size and degree of connectivity in shaping metapopulation dynamics. It is common to regard an amount of habitat area as a potential indication of population size for a given species while overlooking the crucial role that even miniscule differences in quality can play (Pellet et al. 2007). This is particularly important regarding how one identifies potential source and sink habitat patches for a metapopulation, as the largest patches are not necessarily the sources (Tveit 2008). An example of this was partially observed in Chamberlain et al.’s (1999) study (above) when they explain the possible reasons for greater bird densities on OA farms only outside the breeding season. Potentially greater breeding success despite lower density of individuals through ideal despotic distribution would in the case of Chamberlain et al.(1999) make his OA sites an unanticipated source. In addition to being more structurally divers as indicated by this metaanalysis, stability is an apparent quality more intrinsic to OA non-crop habitats as indicated by plant studies which inclusion of species associated with more stable conditions (e.g. Aude et al. 2004; Boutin et al. 2008; Gabriel et al.

2006; Pfiffner & Luka 2003). Disturbance in this case is most likely experienced through the drift of agro-chemicals (synthetic herbicides/pesticides and fertilizers) applications which can have enormous spatial and temporal implications for species richness and abundance.

Furthermore, soil conservation practices such as no-till are increasing in popularity in OA (Badgley et al. 2006).

Vuilleumier et al. (2007) expects that metapopulation persistence can be highly influenced by disturbances that directly increase the extent of local extinctions and /or reduce the pool of empty patches that could be colonized.

The thin edge of the connectivity coin often overlooked, is the boundary between habitat and surrounding matrix. The ecological characteristics of matrix are also influential on migration rates between

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fragments (Jules & Shahani 2003). Depending on a particular species the habiat-matrix boundary can either be relatively impervious, or highly porous (Collinge 1996).

Many see the agricultural matrix initially as desolate and inhospitable for patch-species and focus is thus directed on the value of patches. However Jules & Shahani (2003) suggest that matrix can in fact be of value for certain species. In their example of pollinators, an increase in the amount of flowering plant species in the matrix can increase its permeability, while simultaneously augmenting pollinator numbers which could in turn augment pollination effectiveness in proximal patches. We can recall that this resembles the theorized catch twenty-two explanation by Gabriel et al. (2007) for greater abundance and richness of insect pollinated plants on OA farms, i.e. more pollinators = more flowering plants = more pollinators = etc. A significantly greater matrix quality is one area in particular where OA management is indeed drastically different from CA management. This is primarily the result of greater stability as described above, which as seen in this study can further result not only in a greater diversity of flowering plants, but a greater diversity of these plants within the fields, i.e. matrix. Furthermore, OA farms are in general more extensive and by being so incorporate more flowering plants (quite often nitrogen fixing plants, e.g. legumes) in their crop rotations (Stolze et al. 2000) as actual crop plants (e.g. soybean Glycine max) and/or as green manure plants (e.g. red clover Trifolium pretense). This extension of vegetative structural heterogeneity in the OA matrix in addition to greater stability has obvious positive implications for matrix quality regarding arthropods, with further implications for species which are top-down dependant (recource limited) on these, such as bats (Hendrickx et al. 2007; Wickramasinghe et al. 2004).

It is logical that the positive effects of OA on facilitation of dispersal and settlement should be particularly important for species groups with relatively limited dispersal capabilities such as certain plants and apterous arthropods, as well as species not limited by dispersal capabilities but otherwise possess narrow habitat requirements (regarding e.g. size, shape, disturbance regime, distances between separate resources such as nesting sites and foraging sites, etc) such as small to medium sized bumblebees (Bombus spp.), Skylark (Alauda arvensis), or grass snake (Natrix natrix). For example a study by Aviron et al. (2005) addressing ground/tiger beetle (Carabidae) assemblages in relation to land use intensity level, divided these up into species with low (LM) vs. high mobility (HM) with the hypothesis that LM species perceive the landscape at fine scales and HM species at coarser scales. The results revealed that LM species, often large and apterous, responded best to the proportion of woody elements in landscape context at the intermediate scale of analysis (250 m) and to the amount of

crops at fine scale (50 m) (Aviron et al. 2005). Likewise, HM species, usually smaller, displayed no response to landscape descriptors at any scale (Aviron et al. 2005).

Thus landscape context had a significant effect on carabid species assemblages based on their dispersal abilities, however lower than that of habitat type (Aviron et al.

2005). The greatest proportion of area covered by landscape descriptors of importance to LM carabids (woodland and fallow land, permanent grasslands, and increased hedgerow density) were in the least intensive agricultural landscapes (Aviron et al. 2005). Moreover, the lesser the extent of recurrent disturbances in the less intensely driven landscapes, such as ploughing, may also have been important consequence for the LM species as these are most affected by due to greater needs for stable resources (Aviron et al. 2005). In light of examples like this it is clear that one major weakness in many OA vs. CA biodiversity studies with mixed or insignificant results is clearly due to the generalization of species groups. Nearly every study that went a step further than simple taxonomic divisions and additionally looked at the dispersal abilities and/or niche width of their samples discovered that those individual species which are particularly limited in regard to dispersal or niche were those with the most positive associations to greater vegetative structural heterogeneity, habitat patch diversity, and/or surrounding landscape heterogeneity (e.g. Boutin et al. 2008; Gabriel et al. 2006; Pfiffner & Luka 2003; Roschewitz et al. 2005; Rundlof et al. 2008; Wickramasinghe et al. 2003). One more misleading error seems to be not taking into account the surrounding landscape heterogeneity at all. Every study which specifically looked at farm pairs in landscapes with different degrees of complexity uncovered that OA vs. CA pairs in landscapes with high heterogeneity showed either a synergistic effect between OA and surrounding landscape, or little difference at all. Likewise pairs in homogenous landscapes on the contrary showed results clearly favouring OA (e.g. Aviron et al. 2005; Boutin et al. 2008; Clough et al. 2007a; Clough et al. 2007b; Purtauf et al.

2005; Roschewitz et al. 2005; Rundlof & Smith 2006; Rundlof et al. 2008; Schmidt et al. 2005).

Other Meta-analyses Asking if OA benefits biodiversity, Hole et al.

(2005) appear to think that the answerer is yes based on their own meta-analysis on this topic. Their explanation of the visible benefits of OA management are based on its provision of a greater quantity and/or quality of both crop and non-crop habitat than in CA (Hole et al. 2005).

Furthermore they list three broad characteristics of management largely intrinsic though not exclusive to OA which are likely to benefit farmland biodiversity, these being; i: “prohibition/reduced use of chemical pesticides and inorganic fertilisers” ii: “sympathetic management of non-crop habitats and field margins” iii: “preservation of

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mixed farming” (Hole et al. 2005). Bengtsson et al. (2005) also found that OA is frequently associated with positive effects on species richness and abundance, though these effects are likely to differ between organism groups and landscapes. Here, birds, predatory insects, soil organisms and plants responded positively to organic farming, whereas non-predatory insects and pests did not (Bengtsson et al. 2005). Furthermore, OA’s positive effects on abundance were most important at the plot and field scales, but not for farms in matched landscapes (Bengtsson et al. 2005). The meta-analysis by Fuller et al. (2005) agrees that organic farming is associated with higher levels of biodiversity and that significant differences are recorded based on species groups. Like Bengtsson et al. (2005), Fuller et al. (2005) agrees that plant studies in particular displayed far more consistent and pronounced results in favour of OA than studies on other taxonomic groups, however Fuller et al. (2005) disagrees with Bengtsson et al.

(2005) regarding studies of predatory invertebrates and states that these only occasionally demonstrated significant responses. For other species groups Fuller et al. (2005) further states that though many significant differences are uncovered, these often possess wide confidence intervals.

This meta-analysis corresponds with these other more thorough journal published meta-analyses in seeing that OA farms differ from CA equivalents in habitat extent, composition and management (Fuller et al. 2005). Granted this, and the exclusion of agrochemicals in the matrix, it seems sensible to expect far more significant results from studies on this topic than has been the case. Like this study, Fuller et al. (2005) points to the differential impacts of temporal and spatial scales on the colonization traits of organisms as a likely explanation to the discrepancies regarding results. Plants for example are more immediately affected by herbicide and fertilizer application and can quickly recolonize from the seed bank following conversion to OA, while recolonization of other species groups appears to be more directly dependant on proximity to population sources in time and space (Fuller et al. 2005).

Otherwise these other meta-analyses seem to overlook potential differences in niche requirements and dispersal capabilities of specific species within a taxonomic group under focus.

Moreover, the result interpretations within these other meta-analyses are not discussed to any great detail in light of landscape ecological theories other than with brief statements, as if such theories are too obvious to question.

This is strange given statements from authoritive authors such as Brown & Lomolino (2000), e.g. “It (island biogeography theory) has not kept pace with relevant theory and our growing appreciation for the complexity of nature, especially with empirical findings”, or Pellet et al.

(2007) “Although widely used in conservation, metapopulation models are based on multiple simplifying assumptions that rarely have been validated empirically”,

while others like Collinge (1996) state “Ecologists are increasingly able to understand and predict the consequences of human-induced loss and isolation of native habitats due to the concepts of island biogeography and metapopulation dynamics, combined with empirical field studies in fragmented habitats”.

Conclusion The story of modern agriculture has sadly emphasized uncritical maximization of short term gain rather than long term sustainability, which has resulted in the fragmentation and homogenization of our cultural landscapes. In landscapes rendered most penurious by intense conventional agriculture, islands of organic management appear to stand out like biodiversity oases, while differences in biodiversity parameters between organic and conventional management seem more obscure in landscapes still rich in complexity. It is these obscurities in the apparently richer landscapes that organic agriculture may or may not be playing important roles for biodiversity which to date have mostly eluded scientific remark, e.g. greater matrix quality, higher diversity of insect pollinated plants, or greater nesting success for farmland birds.

Whether agricultural land surrounds protected areas or smaller fragments of natural to seminatural environment, or conversely dominates the view of the landscape, it deserves cultivation that conserves the ecological integrity of the landscape by facilitating heterogeneity and the migration of species. Organic management with its positive trends in biodiversity need not be the modern biodiversity solution in agriculture, but continues offering important lessons which to a larger extent should be dissected by the facets of landscape ecology. Future studies addressing this topic will benefit from a higher degree of classification by species phenology, guild and degree of dispersal capability, together with more highly defined and directly tested explanatory factors in farm management and landscape scale and qualities, with larger spatial and temporal samples sizes.

Acknowledgements This study was facilitated by the course in landscape ecology at the Department of Landscape Architecture and Spatial Planning with Norwegian University of Life Science. Thanks to my wife’s tolerance of me working long hours at school. Thanks to the course teachers Mari Tveit and Gary Fry for constructive insight. This paper was supported by the lånekassen.

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