**Biodiversity and Ecosystem Functioning in Tropical Habitats — Case Studies and Future Perspectives in Atlantic Rainforest and Cerrado Landscapes**

Tathiana G. Sobrinho, Lucas N. Paolucci, Dalana C. Muscardi, Ana C. Maradini, Elisangela A. Silva, Ricardo R. C. Solar and José H. Schoereder

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/59042

## **1. Introduction**

Currently, environmental changes can be seen as an intrinsic feature of ecosystems, once finding ecosystems that do not suffer of anthropogenic pressures, either direct or indirect, is rare [1]. Such pressures come from the continuous and exponential human population growth, which propels urbanization, activities and processes directly linked to the use of fossil fuels, mining, agriculture and cattle growth. The maintenance of current human population growth implies in the supply of a huge demand for food and technology, resulting in rising pollution and loss of habitats and entire ecosystems [2-5].

Anthropogenic impacts alter the physical environmental characteristics, climate, temperature, soil and water quality, and biogeochemical cycles, interfering directly on the biota [6-11]. The immense resource consumption exerted by the human population demands an ongoing exploitation of natural resources. This causes a constant increase of greenhouse gas emissions and, consequently, the temperature around the globe, also generating an intense conversion of soil use [12,13]. The shift from natural environments into cultivated soils became noteworthy in several countries mainly after the Green Revolution, and together with the frequent use of fertilizers, have changed or destroyed natural habitats, decreasing biodiversity directly or indirectly [9,14]. In addition to these factors, species overexploitation and species invasion

© 2015 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2014 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and eproduction in any medium, provided the original work is properly cited.

have contributed to the decline of biodiversity [15]. Current data indicates that almost 30% of known species in the World became extinct or are endangered due to anthropogenic pressures.

The loss of biodiversity itself is not the only problem associated with human disturbance, as such habitat changes may have a harmful cascade effects that alter other environmental properties. Every biodiversity loss may be translated into a loss of functional diversity [16], which is related to the characteristics of organisms that allow them to perform different function in the ecosystem [16,17]. These functions, such as seed dispersal [18,19], pest and weed biological control [20], pollination [21], nutrient cycling [22], and the decomposition of organic matter, among others, are essential to ecosystem functioning maintenance. Ecosystem functioning include biogeochemical and ecosystem processes [23], responsible by matter cycling and energy flow, being directly related to resource dynamics and ecosystem stability [17]. Generally, the ecosystem functioning can be estimated as the magnitude and dynamics of ecosystem processes resultant from the interactions within and between different levels of biota [24].

Traditionally, the main parameters used for estimate ecosystem functioning are linked to plant communities, such as primary productivity and biomass stability [25-27]. Processes mediated by other organisms, including arthropods, have seldom been used for such estimates [28,29], regardless their crucial role. Arthropods are very diverse organisms, abundant everywhere, highly successful and spread across the globe. They represent more than 80% of all described biodiversity and, among them, insects are the most abundant [30]. Arthropods perform several functions in ecosystems, at different levels. These organisms inhabit from the underground soil layers to the top of tress, are engaged in several trophic levels, and interfere in the occurrence and distribution of several other organisms through intricate interactions such as predation, competition, herbivory and mutualism. Arthropods perform soil bioturbation [31], pest, weed, parasite and disease control [32-34], pollination [35,36], seed dispersal [18,37], dung and carrion removal [37], and act in decomposition and nutrient cycling [22,38,39]. Thus, it is expected that changes in the diversity of arthropods can also trigger changes in the ecosystem functioning.

Functional features of species may influence how ecosystem functioning will be altered with biodiversity loss [40]. Different species may be redundant in the functions they play in the ecosystem, and in this case species loss would be compensated by another one that performs the same function. Hence, biodiversity loss would not necessarily cause decrease on ecosystem functioning, as well as it would not increase if new species were added (redundancy hypoth‐ esis, or null hypothesis). Alternatively, some species may be singular or unique in the functions they play within the ecosystems, and their loss would eventually result in a decrease of ecosystem functioning (linearity hypothesis). Finally, the effect of species loss or gain on ecosystem functioning may be dependent on the conditions (community species composition or soil fertility, for instance) under which these biodiversity alterations occur, and the outcome would be unpredictable (idiosyncrasy hypothesis) [40].

The linearity hypothesis is a pattern frequently reported in studies carried out in temperate regions [41], while in the tropics the null hypothesis seems to be the most common. This difference in the reported patterns may be linked to higher biodiversity levels found in the tropics, suggesting a possible functional redundancy among species, which does not seem to occur in temperate regions. Notably, tropical regions harbor the highest World biodiversity [42,43], once 16 of the 25 biodiversity hotspots are located in these regions [43,44]. Conversely tropical regions also exhibit the highest rates of species loss due to human activities [45,46]. Among the main human alterations that cause biodiversity loss is land use change [47-49] and, according to estimates, it will remain as the main activity during the next 100 years [50]. Several tropical biomes have experienced high biodiversity loss due to land use change and, in Brazil, the Atlantic Forest and the Cerrado (Brazilian savanna) may be highlighted [8,51-53].

have contributed to the decline of biodiversity [15]. Current data indicates that almost 30% of known species in the World became extinct or are endangered due to anthropogenic pressures. The loss of biodiversity itself is not the only problem associated with human disturbance, as such habitat changes may have a harmful cascade effects that alter other environmental properties. Every biodiversity loss may be translated into a loss of functional diversity [16], which is related to the characteristics of organisms that allow them to perform different function in the ecosystem [16,17]. These functions, such as seed dispersal [18,19], pest and weed biological control [20], pollination [21], nutrient cycling [22], and the decomposition of organic matter, among others, are essential to ecosystem functioning maintenance. Ecosystem functioning include biogeochemical and ecosystem processes [23], responsible by matter cycling and energy flow, being directly related to resource dynamics and ecosystem stability [17]. Generally, the ecosystem functioning can be estimated as the magnitude and dynamics of ecosystem processes resultant from the interactions within and between different levels of

Traditionally, the main parameters used for estimate ecosystem functioning are linked to plant communities, such as primary productivity and biomass stability [25-27]. Processes mediated by other organisms, including arthropods, have seldom been used for such estimates [28,29], regardless their crucial role. Arthropods are very diverse organisms, abundant everywhere, highly successful and spread across the globe. They represent more than 80% of all described biodiversity and, among them, insects are the most abundant [30]. Arthropods perform several functions in ecosystems, at different levels. These organisms inhabit from the underground soil layers to the top of tress, are engaged in several trophic levels, and interfere in the occurrence and distribution of several other organisms through intricate interactions such as predation, competition, herbivory and mutualism. Arthropods perform soil bioturbation [31], pest, weed, parasite and disease control [32-34], pollination [35,36], seed dispersal [18,37], dung and carrion removal [37], and act in decomposition and nutrient cycling [22,38,39]. Thus, it is expected that changes in the diversity of arthropods can also trigger changes in the ecosystem

Functional features of species may influence how ecosystem functioning will be altered with biodiversity loss [40]. Different species may be redundant in the functions they play in the ecosystem, and in this case species loss would be compensated by another one that performs the same function. Hence, biodiversity loss would not necessarily cause decrease on ecosystem functioning, as well as it would not increase if new species were added (redundancy hypoth‐ esis, or null hypothesis). Alternatively, some species may be singular or unique in the functions they play within the ecosystems, and their loss would eventually result in a decrease of ecosystem functioning (linearity hypothesis). Finally, the effect of species loss or gain on ecosystem functioning may be dependent on the conditions (community species composition or soil fertility, for instance) under which these biodiversity alterations occur, and the outcome

The linearity hypothesis is a pattern frequently reported in studies carried out in temperate regions [41], while in the tropics the null hypothesis seems to be the most common. This difference in the reported patterns may be linked to higher biodiversity levels found in the

would be unpredictable (idiosyncrasy hypothesis) [40].

biota [24].

522 Biodiversity in Ecosystems - Linking Structure and Function

functioning.

The Atlantic Forest originally covered ca. 150 million hectares along the Brazilian coast [52], occurring in tropical and subtropical regions, and including sites with large altitude variation, humidity, temperature and rainfall regimes. Such a variation in abiotic condi‐ tions allowed the differentiation of several phytophysiognomies, high endemism and the occurrence of numerous rare species, harboring ca. of 8% of Worlds' biodiversity. Howev‐ er, recent estimates indicate that more than 90% of its vegetation cover has been lost, and the Atlantic Forest is nowadays composed by forest fragments, mostly smaller than 250 ha, immerse in a landscape of different human modified habitats [52]. The effects of habitat loss due to human activities are extensively reported in the literature concerning the Atlantic Forest [8,54-56], but the effects of biodiversity decrease on the processes related to ecosys‐ tem functioning still need more consistent information. The Brazilian Cerrado is the second larger biome of the country, occupying originally approximately 22% of its area [57], and stands out by its high biodiversity and endemism. The Cerrado is currently suffering an elevated degree of human exploitation, and nowadays remains less than 30% of its original area [43]. High fragmentation and conversion in pasture or agriculture areas, cause biodiversity loss, soil erosion, arrival and establishment of invasive species, and shifts on fire regimes, carbon cycles and climate [51-53].

In this chapter we aimed to evaluate the relationship between biodiversity and ecosystem functioning in the tropical biomes Atlantic Forest and Cerrado. We report here three case studies that investigate different ecosystem processes modulated by arthropods: litter decomposition, seed dispersal and protection against herbivores. In these studies we seek to understand the relative importance of species richness and of the presence of keystone species on the studied ecosystem functions.

We performed the first case study in a secondary forest fragment, in the Atlantic Forest biome. In this study we test the relationship between litter decomposition and the biodiversity of several functional groups of soil and litter arthropods. We performed the second case study in the Cerrado biome, testing the effect of ants biodiversity that visit extrafloral nectaries on the protection of these plants against herbivores and herbivory.

Lastly, in the third case study, carried out in the same region of the first one, we analyzed the effect of ant biodiversity on seed removal, comparing secondary forests and *Eucalyptus* crops. In this study we test the direct effect of land use change on the relationship between biodi‐ versity and functioning. From the analysis of these three case studies we concluded this chapter presenting some future perspectives of studies on this subject, to solve some knowledge gaps related to the biodiversity and ecosystem functioning relationship in tropical ecosystems.

### **2. Case study 1**

Decomposition is the process that transforms nutrients retained in organic matter into their inorganic form, making them available in the soil to the primary producers [58,59], and is therefore a key supporting process for the functioning of ecosystems. This process is ruled by three main factors: the physicochemical environment, the quality of the decomposing material and the soil and litter fauna [58,60-63]. These factors present different interaction routes [64] and the relative importance of each component changes in different time and spatial scales [65].

The physicochemical environment is related to the climate, or microclimate, mainly humidity and temperature [66,67]. Abiotic conditions may indirectly affect decomposition, altering litter characteristics, or directly, controlling the activity of decomposers [66,68]. Litter quality is usually associated to foliar material degradability [69], as the concentration of some nutrients has been frequently associated to its palatability to organisms [70]. Usually a higher initial nitrogen concentration reflects in a higher organic matter quality to decomposers. Finally, organisms living in litter and soil are crucial for decomposition processes and nutrient release [60,71-73]. These organisms revolve, mix, break and digest the detritus, metabolizing the litter constituents [58]. Among the components of the soil community, fungi and bacteria are the main decomposing agents. Nevertheless, the micro and mesofauna of soil and litter arthropods have an important role in the decomposition process, through fragmentation of organic matter, through the mixing and vertical movement of organic matter [74]. The existence of an abundant and diversified arthropod fauna is expected, then, to favor an enhanced nutrient cycling [75] and a subsequent faster plant growth [76].

The abundant arthropod fauna composing soil and litter communities can be categorized into different guilds or functional groups, according to their activities, which may affect the microbial community by several ways [77]. Fungivores and bacteriovores consume exclusively the microorganisms, decreasing their abundance. Moreover, they can decrease their prey species richness, through an intense predation, or else an increase of this species richness, through the top-down control of the more competitive species, mediating their coexistence. Detritivores, on the other hand, consume part of organic matter together with the film of microorganisms, releasing the broken and partially digested organic matter in their faeces. As a result, besides their negative effects on microorganisms due to predation, detritivores may increase litter fragmentation, resulting in more decomposing surface and higher decomposer abundance an species richness. These organisms interact in complex food webs and therefore diversity and abundance changes of a given functional group or guild may alter abundance, diversity and functioning of another group [78,79]. It is important hence the investigation of the functional groups role on the decomposition process, as different guilds may interfere more than others.

The process of litter decomposition, as well as the intricate relationships among the diverse components of the edaphic fauna associated to the litter, offers an excellent study system of the relationship between biodiversity and ecosystem processes, mainly in tropical environ‐ ments with their huge diversity. In this study case we verify how soil and litter arthropod biodiversity affects litter decomposition in a tropical habitat. Our hypothesis is that increasing arthropod abundance and species richness cause higher decomposition rates, and that some functional groups may have stronger roles in this process.

### **2.1. Methods**

**2. Case study 1**

544 Biodiversity in Ecosystems - Linking Structure and Function

and a subsequent faster plant growth [76].

than others.

Decomposition is the process that transforms nutrients retained in organic matter into their inorganic form, making them available in the soil to the primary producers [58,59], and is therefore a key supporting process for the functioning of ecosystems. This process is ruled by three main factors: the physicochemical environment, the quality of the decomposing material and the soil and litter fauna [58,60-63]. These factors present different interaction routes [64] and the relative importance of each component changes in different time and spatial scales [65]. The physicochemical environment is related to the climate, or microclimate, mainly humidity and temperature [66,67]. Abiotic conditions may indirectly affect decomposition, altering litter characteristics, or directly, controlling the activity of decomposers [66,68]. Litter quality is usually associated to foliar material degradability [69], as the concentration of some nutrients has been frequently associated to its palatability to organisms [70]. Usually a higher initial nitrogen concentration reflects in a higher organic matter quality to decomposers. Finally, organisms living in litter and soil are crucial for decomposition processes and nutrient release [60,71-73]. These organisms revolve, mix, break and digest the detritus, metabolizing the litter constituents [58]. Among the components of the soil community, fungi and bacteria are the main decomposing agents. Nevertheless, the micro and mesofauna of soil and litter arthropods have an important role in the decomposition process, through fragmentation of organic matter, through the mixing and vertical movement of organic matter [74]. The existence of an abundant and diversified arthropod fauna is expected, then, to favor an enhanced nutrient cycling [75]

The abundant arthropod fauna composing soil and litter communities can be categorized into different guilds or functional groups, according to their activities, which may affect the microbial community by several ways [77]. Fungivores and bacteriovores consume exclusively the microorganisms, decreasing their abundance. Moreover, they can decrease their prey species richness, through an intense predation, or else an increase of this species richness, through the top-down control of the more competitive species, mediating their coexistence. Detritivores, on the other hand, consume part of organic matter together with the film of microorganisms, releasing the broken and partially digested organic matter in their faeces. As a result, besides their negative effects on microorganisms due to predation, detritivores may increase litter fragmentation, resulting in more decomposing surface and higher decomposer abundance an species richness. These organisms interact in complex food webs and therefore diversity and abundance changes of a given functional group or guild may alter abundance, diversity and functioning of another group [78,79]. It is important hence the investigation of the functional groups role on the decomposition process, as different guilds may interfere more

The process of litter decomposition, as well as the intricate relationships among the diverse components of the edaphic fauna associated to the litter, offers an excellent study system of the relationship between biodiversity and ecosystem processes, mainly in tropical environ‐ ments with their huge diversity. In this study case we verify how soil and litter arthropod biodiversity affects litter decomposition in a tropical habitat. Our hypothesis is that increasing

We carried out this study from July 2008 to February 2009, in a ca. 300 ha secondary forest in Viçosa, Minas Gerais, Southeast Brazil (20°45'S e 42°55'W). The main vegetation is composed by Semidecidual Seasonal Atlantic Forest, located within the domain of the Atlantic Forest [80]. In the study area we set two 75m parallel transects, apart 5 m from each other. Along each transect we delimited 15 1m2 squares, 5m distant from each other, in a total of 30 sampling points. We collected approximately 200g of freshly fallen leaves from predominant tree species in each sampling point. These leaves were mixed and oven-dried at 60°C for 72 hours. Dried leaves were weighted and separated in groups of 5g, which were placed into litter bags, measuring 15 x 15 cm, with a mesh of 2 mm [81,63]. In each sampling point we set 15 litter bags and, after 30 days we started to remove them. Litter bags were removed fortnightly along 225 days. At the end of the experiment we took a 20 cm deep soil sample in each sampling point, which were taken to soil analyses. The soil analyses were performed at the Soil Lab analyses of the Federal University of Viçosa, and consisted of organic matter content and macroporosity, variables that could interfere in the decomposition process.

After removal, we placed litter bags in Berlese funnels for 48 hours, to extract the arthropods. After their identification, arthropods were sorted according to their feeding habits: detriti‐ vores, fungivores and predators [82-86]. The arthropods that we cannot sort in the above categories, because we could not identify feeding habits, were classified as "other arthropods", and were considered only in the analyses that included all arthropods.

After the arthropod extraction, litter was oven-dried at 60°C for 72 hours and weighted to compare with the initial weight (5g). We considered litter weight loss as the difference between initial and final weight (after 225 days), and we used this as an estimate of decomposition in each sampling point.

### **2.2. Statistical analyses**

To test the hypothesis that more arthropod richness and abundance leads to a higher litter decomposition rate we used a model selection approach [87,88]. The response variable was litter weight loss, and explanatory variables were: total abundance and richness of arthropods, abundance and richness of fungivores, detritivores and predators, as well as macroporosity and soil organic matter. Before structuring the model, we carried out a correlation test among the explanatory variables, using the package "psych", and whenever two variables presented a correlation higher than 0.7 we removed the variable considered biologically less relevant [89]. Variables that presented correlation higher than 0.7 were: total arthropod species richness and predator species richness (0.73) and total arthropod abundance and detritivore abundance (0.94). We opted, then, to remove predator species richness and total arthropod abundance, as the former represents a possible action of organisms distant from the focal process of decom‐ position and the latter because it is an estimate more general that detritivore abundance.

The procedure of model selection involved the "MuMIn" package [90], that allows the construction of all possible models starting from the global model containing all variables. For each model, the procedure calculates model weight, based on the Akaike Information Criteria– AICc(ω). After doing so, it ranks all models and the best models are those containing lower AICc and higher weight values. We standardized and centralized all explanatory variables [91], using the package "arm" and the models were built with these transformed variables prior to model selection. All models within ΔAICc < 2 bounds were considered to obtain a good evidence of support [87]. In the case of more than one model, we averaged the models to obtain only one final model with averaged model coefficients, including their respective confidence internal. Parameters for which the confidence interval crossed zero were consid‐ ered non-significant [88]. All analyses were performed under the platform R [92].

### **2.3. Results and discussion**

### *2.3.1. Litter arthropod fauna*

We sampled 2,284 individual and 198 arthropod species, from seven classes: (i) Arachnida, (ii) Malacostraca, (iii) Symphyla, (iv) Chilopoda, (v) Diplopoda, (vi) Entognatha and (vii) Insecta. The class with more orders was Insecta (10 orders), followed by Arachnida (four), Entognatha (three), Malacostraca, Symphyla, Chilopoda and Diplopoda (one order each). The most abundant arthropods in our sampling were Acari and Collembola, which are usually described as more abundant in soil and litter [93]. Besides, high abundance of these two groups had already been reported by [94] and [95], who studied forest fragments in the same region. Oribatid mites were the most respresentative group in all sampling, and these mites have an important role in decomposition process, as most are detritivore [93]. Collembola also presented high abundance and species richness in the samples. These organisms are fungivores and their trophic activity includes both the direct consumption of microorganisms and organic matter fragmentation [96]. Besides, they constitute an important source of food to predatory organisms, being very important in food webs to soil and litter [97].

### *2.3.2. Arthropod biodiversity and ecosystem functioning*

Opposed to what we expected, there was no effect of arthropod species richness and abun‐ dance on decomposition rates, both considering total arthropods and when they were sorted by their feeding habits. Although our final model presented soil macroporosity and detritivore species richness as explanatory variables, their 95% confidence interval includes zero, and were considered non-significant (Table 1).

Our results contrast with others, which reported a positive effect of species richness on ecosystem processes [98-103]. The lack of relationship in our study may have occurred due to a high functional redundancy among arthropod species [40,104]. Accordingly, we infer that the studied community is composed by species with similar functions, thus species loss does not cause changes on ecosystem functioning. However, this possible redundancy assumed in this case study does not necessarily exclude another hypothesis to explain the biodiversity-


The procedure of model selection involved the "MuMIn" package [90], that allows the construction of all possible models starting from the global model containing all variables. For each model, the procedure calculates model weight, based on the Akaike Information Criteria– AICc(ω). After doing so, it ranks all models and the best models are those containing lower AICc and higher weight values. We standardized and centralized all explanatory variables [91], using the package "arm" and the models were built with these transformed variables prior to model selection. All models within ΔAICc < 2 bounds were considered to obtain a good evidence of support [87]. In the case of more than one model, we averaged the models to obtain only one final model with averaged model coefficients, including their respective confidence internal. Parameters for which the confidence interval crossed zero were consid‐

ered non-significant [88]. All analyses were performed under the platform R [92].

organisms, being very important in food webs to soil and litter [97].

*2.3.2. Arthropod biodiversity and ecosystem functioning*

considered non-significant (Table 1).

We sampled 2,284 individual and 198 arthropod species, from seven classes: (i) Arachnida, (ii) Malacostraca, (iii) Symphyla, (iv) Chilopoda, (v) Diplopoda, (vi) Entognatha and (vii) Insecta. The class with more orders was Insecta (10 orders), followed by Arachnida (four), Entognatha (three), Malacostraca, Symphyla, Chilopoda and Diplopoda (one order each). The most abundant arthropods in our sampling were Acari and Collembola, which are usually described as more abundant in soil and litter [93]. Besides, high abundance of these two groups had already been reported by [94] and [95], who studied forest fragments in the same region. Oribatid mites were the most respresentative group in all sampling, and these mites have an important role in decomposition process, as most are detritivore [93]. Collembola also presented high abundance and species richness in the samples. These organisms are fungivores and their trophic activity includes both the direct consumption of microorganisms and organic matter fragmentation [96]. Besides, they constitute an important source of food to predatory

Opposed to what we expected, there was no effect of arthropod species richness and abun‐ dance on decomposition rates, both considering total arthropods and when they were sorted by their feeding habits. Although our final model presented soil macroporosity and detritivore species richness as explanatory variables, their 95% confidence interval includes zero, and were

Our results contrast with others, which reported a positive effect of species richness on ecosystem processes [98-103]. The lack of relationship in our study may have occurred due to a high functional redundancy among arthropod species [40,104]. Accordingly, we infer that the studied community is composed by species with similar functions, thus species loss does not cause changes on ecosystem functioning. However, this possible redundancy assumed in this case study does not necessarily exclude another hypothesis to explain the biodiversity-

**2.3. Results and discussion**

566 Biodiversity in Ecosystems - Linking Structure and Function

*2.3.1. Litter arthropod fauna*

**Table 1.** Summary of model averaging results, detailing the explanatory variables present in the final average model. Parameters estimates were obtained from standardized variables.

functioning observed: the linearity hypothesis. Two curves may be generated by these two hypotheses: a linear relationship (Type I curve) in the case of singular species and an asymp‐ totic curve (Type II curve) in the case of redundant species [105]. Therefore, both hypotheses may be explained by the same curve, depending on the scale data was sampled. Linearity, then, would be a component of redundancy curve, but that would only be expected in cases with low diversity. From a given species richness a saturation of the functions would occur, with species playing similar roles. Data obtained in the present study would fit into this latter diversity scale. To test such assumption one can manipulatively reduce arthropod abundance and richness, studying a broad range of species richness, and effectively testing the redun‐ dancy hypothesis in tropical environments.

Another possible explanation to the absence of relationship between arthropod diversity and litter decomposition is the similar litter constitution across all sampling points. It is known that the chemical and physical composition of litter has an effect on decomposition rates [61,62,70,106,107]. The manipulation of litter diversity and composition in litter bags may lead to the establishment of different arthropod communities, according to leaf degradability. The manipulation of species richness and composition of plants under decomposition may lead to different responses of arthropod species richness, which might mirror variation in decompo‐ sition rates.

Furthermore, in this study we evaluated only the role of arthropod diversity of the soil-litter system. It is known, however, that fungi and bacteria (the microflora) are the decomposers and responsible for organic matter mineralization [58,60,71]. Conversely arthropods act indirectly on decomposition and, even though they facilitate the action of true decomposers, detecting their action on decomposition may be less straightforward.

The absence of relationship between litter decomposition and arthropod species richness and abundance must be evaluated with caution, because assuming functional redundancy among species may be uncertain. Such outcome may lead us to the wrong conclusion that species loss does not affect ecosystem functioning at all, and this early, which may be wrong as discussed above.

Our conclusion is that litter decomposition process in the tropics and other hyper diverse habitats may be more complex than it is for the well known temperate habitats. Studying only one component may not give a precise response, due to the immense assembly of components in complex habitats and their equally complex interactions. Manipulative experiments, microbial activity estimates, as well as manipulation of litter diversity and composition should give us a more precise knowledge of the biodiversity role on ecosystem key processes.

## **3. Case study 2**

Besides the ecosystem processes, as decomposition and nutrient cycling, species interactions are also important in the maintenance of ecosystem stability. The importance of predation and competition in community structuring is well studied [108-110], although mutualism may also have a central role in species distribution in ecosystems [111,112].Ants may establish a wide variety of mutualistic associations with plants [113,114]. Plants may offer resources like shelter, food, or both, that may be used by ants in several ways. The mutualistic interactions between ants and plants vary from diffuse, such as secondary seed dispersal [115] and use of extrafloral nectaries (EFNs) by generalist species [116], to more specialized interactions, such as domatia colonization by *Azteca* sp. [117-119,]. On the other hand, ants may also be beneficial to plants, increasing seed dispersal or reducing herbivory, for example [120].

EFNs, are nectar producing structures associated to plant vegetative organs, as leaves or petioles [121]. Extrafloral nectar is a liquid resource, composed by glucose, sucrose and fructose, and containing sometimes amino acids and proteins [122]. EFN-bearing plants are more visited by ants than plants without them [123], and ants that use extrafloral nectar as a resource may establish a generalist association of protection in exchange for food [114,119]. Therefore, the benefits arising from this interaction may explain its success [124]. The interac‐ tion between ants visiting EFNs and the plants has been the subject of several studies, although there are some divergences among the obtained results. Several studies relate advantages for EFN-bearing plants, such as decreasing the herbivory and the abundance of herbivores, or even positive effects on plant fitness [120,121,123,125-127]. However, some studies did not spot beneficial effects of visiting ants [128,131,132], indicating that in some cases ants may not be efficient in reducing herbivory [128]. The outcome of the interaction between ants and plants may depend on feeding habits of the herbivores, due to the interaction between ants and sapfeeding insects. Several authors [123,130-134] suggest that generalist ants feeding on their honeydew protect these insects from predators and competitors over chewing insects [130]. Moreover, plant protection may be related to ant species composition, as different ant species present varied behavior and defensive characteristics [129].

The interaction among ants, EFN-bearing ants and herbivores is very common in the Brazilian Cerrado. This biome is composed by herbs, shrubs and small trees that vary in density, composing different phytophysiognomies [135]. These physiognomies are usually divided in three groups, characterized by fields, savannas and forests [136]. *Qualea grandiflora* (Vochy‐ siacveae) is among the several EFN-bearing plant species of Cerrado, which are medium to large-sized trees that reach 30 meters [137]. It is very studied in the Cerrado as it has a large distribution and abundance, and also because it attracts several ant species to their EFNs, placed at the basis of the petioles, near to leaf insertion [138].

This case study evaluates whether the ants foraging on *Qualea grandiflora* protect these plants against herbivory. We tested the hypotheses that increased ant species richness and abundance (i) decreases herbivore species richness and abundance, and (ii) changes the proportion of herbivore guilds in Cerrado.

### **3.1. Methods**

microbial activity estimates, as well as manipulation of litter diversity and composition should give us a more precise knowledge of the biodiversity role on ecosystem key processes.

Besides the ecosystem processes, as decomposition and nutrient cycling, species interactions are also important in the maintenance of ecosystem stability. The importance of predation and competition in community structuring is well studied [108-110], although mutualism may also have a central role in species distribution in ecosystems [111,112].Ants may establish a wide variety of mutualistic associations with plants [113,114]. Plants may offer resources like shelter, food, or both, that may be used by ants in several ways. The mutualistic interactions between ants and plants vary from diffuse, such as secondary seed dispersal [115] and use of extrafloral nectaries (EFNs) by generalist species [116], to more specialized interactions, such as domatia colonization by *Azteca* sp. [117-119,]. On the other hand, ants may also be beneficial to plants,

EFNs, are nectar producing structures associated to plant vegetative organs, as leaves or petioles [121]. Extrafloral nectar is a liquid resource, composed by glucose, sucrose and fructose, and containing sometimes amino acids and proteins [122]. EFN-bearing plants are more visited by ants than plants without them [123], and ants that use extrafloral nectar as a resource may establish a generalist association of protection in exchange for food [114,119]. Therefore, the benefits arising from this interaction may explain its success [124]. The interac‐ tion between ants visiting EFNs and the plants has been the subject of several studies, although there are some divergences among the obtained results. Several studies relate advantages for EFN-bearing plants, such as decreasing the herbivory and the abundance of herbivores, or even positive effects on plant fitness [120,121,123,125-127]. However, some studies did not spot beneficial effects of visiting ants [128,131,132], indicating that in some cases ants may not be efficient in reducing herbivory [128]. The outcome of the interaction between ants and plants may depend on feeding habits of the herbivores, due to the interaction between ants and sapfeeding insects. Several authors [123,130-134] suggest that generalist ants feeding on their honeydew protect these insects from predators and competitors over chewing insects [130]. Moreover, plant protection may be related to ant species composition, as different ant species

The interaction among ants, EFN-bearing ants and herbivores is very common in the Brazilian Cerrado. This biome is composed by herbs, shrubs and small trees that vary in density, composing different phytophysiognomies [135]. These physiognomies are usually divided in three groups, characterized by fields, savannas and forests [136]. *Qualea grandiflora* (Vochy‐ siacveae) is among the several EFN-bearing plant species of Cerrado, which are medium to large-sized trees that reach 30 meters [137]. It is very studied in the Cerrado as it has a large distribution and abundance, and also because it attracts several ant species to their EFNs,

increasing seed dispersal or reducing herbivory, for example [120].

present varied behavior and defensive characteristics [129].

placed at the basis of the petioles, near to leaf insertion [138].

**3. Case study 2**

588 Biodiversity in Ecosystems - Linking Structure and Function

Sampling was carried out in Panga Ecologic Station (PEE), situated in Minas Gerais, Brazil (19o 09'20"-19o 11'10" S, 48o 23'20"-48o 24'35" W). The area is a 409 ha of Cerrado, with several phytophysiognomies [139, 140]. Climate is Aw, tropical with a rainy summer and dry winter [141]. The average temperature during winter is 18°C and during summer is 23°C, and monthly rainfall is 60 mm during winter and 250 mm during summer.

Insects were sampled during January 2013, during the rainy season. We chose 90 individuals of *Qualea grandiflora*, 30 in each phytophysiognomy: *Cerradão* (Forest), *Cerrado Stricto Sensu* (Dense Savanna) and *Campo Cerrado* (Field Savanna). As it is known that ant species richness and abundance may vary with tree density in Cerrado [142], we expect that sampling in three different plant densities would produce a higher range of variation on ant community parameters. We sampled herbivores by beating, using an entomological umbrella of 1 m2 [143, 144]. We did 10 beatings in each tree, and all insects were collected. All herbivores were counted, identified up to the family level, and sorted into two groups (guilds): leaf chewing insects and sap-sucking insects. Ants were sampled by pitfall traps, placed in the trunk of trees at 1.5 meters above ground level. In each tree we installed four pitfalls, to maximize ant sampling. Pitfall traps remained open for 48 hours, and ants were identified to the lower level as possible (genus or species). When identification up to species level was not possible, we asserted the individuals into morphospecies. Herbivores eventually collected in the pitfall traps were added to the beating sampling.

### **3.2. Statistical analyses**

To test whether an increase of ant species richness and abundance decreases herbivore species richness and abundance, we carried out an ANCOVA (analysis of covariance), with Poisson distribution, considering phytophysiognomies as covariates. To test the hypothesis that increasing ant species richness and abundance decreases the proportion of leaf chewing insects and increases sap-sucking insects, we carried out an ANCOVA to one of the herbivore guilds, using a binomial distribution, corrected for overdispersion. Only one analysis is needed in this case, as the two response variables (proportion of each guild) are complementary. All analyses were carried out in R platform and models were simplified by removing non-significant variables and obtaining the minimal adequate model [145].

### **3.3. Results and discussion**

We sampled 2,597 ants, from 150 species, 25 genera and seven subfamilies. The most abundant subfamily was Formicinae (1,293 individuals), followed by Myrmicinae (737), Dolichoderinae (426), Pseudomyrmecinae (110), Ectatomminae (22), Ponerinae (eight) and Heteroponerinae (one). The subfamily with the highest species richness was Myrmicinae (55 species), followed by Formicinae (40), Pseudomyrmecinae and Dolichoderinae (20 species each), Ectatomminae (three), Ponerinae (two) and Heteroponerinae (one). We sampled 233 herbivore insects, and Coleoptera was the most abundant order (141 individuals), followed by Hemiptera (75), Lepidoptera (11) and Orthoptera (six). Herbivores were sorted in 97 species, and the order with highest species richness was Coleoptera (62 species), followed by Hemiptera (24), Orthoptera (six) and Lepidoptera (five).

Ant species richness did not affect both herbivores abundance and species richness. As the studied plant species is a mirmecophile, it was not expected to maintain obligatory associations with ant species, being visited by generalist ant species foraging both during the day and the night. Several species in a given community may have a redundant role in some ecological functions, such as predation [146]. Moreover, as mirmecophyles are usually visited by several ant species, species richness may not contribute effectively for herbivore decrease. Therefore, ant species richness visiting EFN-bearing plants may not contribute to herbivore decrease because (i) they may be highly redundant, and/or (ii) they encompass non-aggressive ant species.

Nevertheless, we observed a decrease of herbivore abundance with the increase of abundance of ants in the trees of *Cerradão* and *Campo Cerrado* (χ<sup>2</sup> =0.7; p=0.02) (Figure 1). The higher number of ants present in a given site gives a higher probability of encounters between them and herbivores, decreasing the number of herbivores [147]. Some ant species may present aggres‐ sive behavior, or efficient recruitment ability, and if the most frequent ants have these attributes, there would be a higher chance of herbivore attack and decreasing.

Conversely, we also observed that herbivore abundance increased with the abundance of ants in *Cerrado Stricto Sensu* (χ<sup>2</sup> =1.7; p=0.05) (Figure 1). Another interesting result regards a higher sap-sucker insect abundance in *Cerrado Stricto Sensu*, in comparison with the other two phytophysiognomies (Figure 2). This latter result may explain the positive relationship between ant and herbivore abundance, because most of the herbivores belong to sap-sucking insects, and the positive association between them and ants is well known. As there are more sap-sucking insects in the *Cerrado Stricto Sensu*, ants may be consuming more sugars from the honeydew than from the EFNs [148], possibly leading to a dominance of more aggressive, abundant and frequent ants [149]. Therefore, the association between ants and EFN-bearing plants may shift from mutualistic to antagonistic when sap-sucking insects are present. When there is high resource competition on the plants, ants have a tendency to consume more honeydew than nectar [150], protecting sap-sucking insects and harming the plants.

Although the above scenario may arise from the three trophic level interaction mentioned above, it has been suggested that ants may also repel chewing insects by consuming honeydew, decreasing their abundance and activity on plants [133]. The different responses of herbivores to abundance of ants found in our study suggest that the effect of ants on herbivores is dependent of herbivore feeding habits. Several studies have reported non-obligatory interac‐ tions between ants and sap-sucking insects [123,130-134]. In this interaction, ants feed on the honeydew and protect the insects against predators and competitors [130]. Such an interaction may produce an explanation for the above results, because most ants participating in this interaction are generalists, consuming both honeydew and nectar from EFNs. Therefore, both sap-sucking insects and plants may be protected by the ants, as both provide resources to them. In this scenario, leaf chewing herbivores would be repelled or predated by the foraging ants.

(one). The subfamily with the highest species richness was Myrmicinae (55 species), followed by Formicinae (40), Pseudomyrmecinae and Dolichoderinae (20 species each), Ectatomminae (three), Ponerinae (two) and Heteroponerinae (one). We sampled 233 herbivore insects, and Coleoptera was the most abundant order (141 individuals), followed by Hemiptera (75), Lepidoptera (11) and Orthoptera (six). Herbivores were sorted in 97 species, and the order with highest species richness was Coleoptera (62 species), followed by Hemiptera (24),

Ant species richness did not affect both herbivores abundance and species richness. As the studied plant species is a mirmecophile, it was not expected to maintain obligatory associations with ant species, being visited by generalist ant species foraging both during the day and the night. Several species in a given community may have a redundant role in some ecological functions, such as predation [146]. Moreover, as mirmecophyles are usually visited by several ant species, species richness may not contribute effectively for herbivore decrease. Therefore, ant species richness visiting EFN-bearing plants may not contribute to herbivore decrease because (i) they may be highly redundant, and/or (ii) they encompass non-aggressive ant

Nevertheless, we observed a decrease of herbivore abundance with the increase of abundance

of ants present in a given site gives a higher probability of encounters between them and herbivores, decreasing the number of herbivores [147]. Some ant species may present aggres‐ sive behavior, or efficient recruitment ability, and if the most frequent ants have these

Conversely, we also observed that herbivore abundance increased with the abundance of ants

sap-sucker insect abundance in *Cerrado Stricto Sensu*, in comparison with the other two phytophysiognomies (Figure 2). This latter result may explain the positive relationship between ant and herbivore abundance, because most of the herbivores belong to sap-sucking insects, and the positive association between them and ants is well known. As there are more sap-sucking insects in the *Cerrado Stricto Sensu*, ants may be consuming more sugars from the honeydew than from the EFNs [148], possibly leading to a dominance of more aggressive, abundant and frequent ants [149]. Therefore, the association between ants and EFN-bearing plants may shift from mutualistic to antagonistic when sap-sucking insects are present. When there is high resource competition on the plants, ants have a tendency to consume more

honeydew than nectar [150], protecting sap-sucking insects and harming the plants.

Although the above scenario may arise from the three trophic level interaction mentioned above, it has been suggested that ants may also repel chewing insects by consuming honeydew, decreasing their abundance and activity on plants [133]. The different responses of herbivores to abundance of ants found in our study suggest that the effect of ants on herbivores is dependent of herbivore feeding habits. Several studies have reported non-obligatory interac‐ tions between ants and sap-sucking insects [123,130-134]. In this interaction, ants feed on the honeydew and protect the insects against predators and competitors [130]. Such an interaction may produce an explanation for the above results, because most ants participating in this

=1.7; p=0.05) (Figure 1). Another interesting result regards a higher

attributes, there would be a higher chance of herbivore attack and decreasing.

=0.7; p=0.02) (Figure 1). The higher number

Orthoptera (six) and Lepidoptera (five).

1060 Biodiversity in Ecosystems - Linking Structure and Function

of ants in the trees of *Cerradão* and *Campo Cerrado* (χ<sup>2</sup>

species.

in *Cerrado Stricto Sensu* (χ<sup>2</sup>

**Figure 1.** Relationship between abundance of ants and abundance of herbivore insects in the three studied phytophy‐ siognomies. The continuous curve (circles) represent the decrease of herbivores in *Cerradão* and *Campo Cerrado* (χ<sup>2</sup> =0.7; p=0.02), and the dashed line (triangles) the increase of herbivores in *Cerrado Strictu Senso* (χ<sup>2</sup> =1.7; p=0.05).

However, ant species distribution and their consequent effect on interactions may be modu‐ lated by habitat type and conditions [151]. Habitats with larger resource availability may facilitate the coexistence of ant species [131]. As Cerrado is composed by vegetation types with different tree abundances [136], the relationships among organisms may also vary accordingly. EFN-bearing plants were found to be more frequent in the Forest formations of Cerrado (*Cerradão*) than in other phytophysiognomies, which indicates more extrafloral nectar in this vegetation type [152]. As tree density and species richness influences ant species richness due to higher resource availability to generalist and specialist species [142], mutualistic interactions may be dependent of resource amount and distribution. More heterogeneous habitats may generate diverse resource availability, promoting the found differences among the phytophy‐ siognomies studied here. As there are more resources in the *Cerradão*, ant foraging may be more opportunistic, resulting in a less effective protection against herbivory by chewing insects. Additionally, as there are more resources provided by EFNs, associations between sapsucking insects and ants may be less effective, decreasing their abundance.

**Figure 2.** Relationship between SAP-sucking insects in the three phytophysiognomies. C – *Cerradão*, CC – *Campo Cerra‐ do*, CSS – *Cerrado Stricto Sensu*. C and CC did not differed statistically.

**Figure 3.** Relationship between the proportion of chewing insects and ant species richness. While ant species richness increased chewing herbivores in *Cerradão* (circles; continuous line), in decreases this proportion in *Cerrado Stricto Sensu* and *Campo Cerrado* (triangles; dashed line). The relationship between sap-sucking insects and ants follows a pattern contrary to the above, as these proportions are complementary.

## **4. Case study 3**

**Figure 2.** Relationship between SAP-sucking insects in the three phytophysiognomies. C – *Cerradão*, CC – *Campo Cerra‐*

**Figure 3.** Relationship between the proportion of chewing insects and ant species richness. While ant species richness increased chewing herbivores in *Cerradão* (circles; continuous line), in decreases this proportion in *Cerrado Stricto Sensu* and *Campo Cerrado* (triangles; dashed line). The relationship between sap-sucking insects and ants follows a pattern

*do*, CSS – *Cerrado Stricto Sensu*. C and CC did not differed statistically.

1262 Biodiversity in Ecosystems - Linking Structure and Function

contrary to the above, as these proportions are complementary.

The conversion of pristine environments into human-modified landscapes is rising around the World. Such habitat conversions may culminate in altered environmental conditions, reduc‐ tion in the availability of resources and decrease in habitat heterogeneity [153]. Consequently, many authors have been warning to the existence of a biodiversity crisis [154,155]. In general the conversion of natural systems introduces newer and simplified ecosystems composed by one or few economically valuable crop species. Whereas habitat loss per se is enough to generate local extinctions [156,157], what is observed is that these habitats are usually substi‐ tuted by agricultural systems as well. Therefore, most human-modified landscapes are altered by the joint action of these processes, habitat loss and conversion.

Habitat heterogeneity can be defined as the variety and the relative amount of different microhabitats available to organisms, and has been considered over the years a major variable determinant on local species richness and abundance [158-160]. Structurally more complex habitats provide more spatial niches and different types of resource exploitation, thus increasing species diversity [153,161-162], although this relationship may not be always straight [163,164]. Habitat heterogeneity reduction, for instance, can lead to lower resource availability, changes in environmental conditions and eventually species and ecosystem functions losses [165,166].

*Eucalyptus* crops are one of the economic activities that may lead to the above mentioned biodiversity loss in the Brazilian biomes. This culture was introduced in the country by the beginning of 19th century, and up to 2012 it is estimated to cover 5.105.246 hectares [167]. Once *Eucalyptus* is classically grown as a monoculture in Brazil, habitats are extremely simplified and homogeneous, potentially triggering the mentioned effects on biodiversity.

Several functions may be altered in *Eucalyptus* plantations, potentially due to homogenization, such as litter decomposition [97,168], nutrient cycling [169] and seed dispersal [170]. The latter is usually associated with habitat recovery, as there are several advantages for plants such as avoiding rodent predation [171,172], dispersion for nutrient-rich sites [173], protection against fire [174] and smaller competition with the parental plant [175]. Thereby, such mutualism may play a central role on local plant dynamics [176].

Seed dispersal by animals is considered a diffuse interaction, once can be performed by several generalist frugivores [177,178]. Ants are mainly reported as secondary dispersers, as they take fallen diaspores to their nests. Within the nests conditions are more suitable for the seed, because it is protected from herbivores [179-181], it is a nutrient-rich microhabitat [182,183] and free of competition with the parental plant [175,180,184].

A common observed trend is the reduction in ant species richness with habitat conversion, such as pastures [185], crops [186] and in *Eucalyptus* cultures [187,188]. Species composition is also usually altered by habitat conversion [185,189]. These changes may strongly alter seed dispersal dynamics by ants in fragmented and modified landscapes [185,190], although no mechanism was proposed to explain this correlation. In this study we tested the effect of natural habitat conversion into *Eucalyptus* crops and the consequences for seed removal by ants. We hypothesized that higher ant species richness increases seed removal, and this relationship is more pronounced in natural forest ecosystems than in *Eucalyptus* crops. Furthermore, we hypothesized that ant species composition changes between natural and *Eucalyptus* crops, and this shift is also responsible for supposed differential seed removal rates. Finally, as we observed differential seed removal between studied habitats we tested the presence of keystone species would influence removal rates.

### **4.1. Methods**

### *4.1.1. Study site*

We carried out the study in Viçosa, Minas Gerais (20°45'S, 42°50'W), Brazil, during summer 2010/2011, the rainy season. The pristine vegetation in this region is within the Atlantic Forest Domain, and is classified as Seasonal Semi-deciduous Forest. From the 1930's decade an intense fragmentation process has begun, and the native vegetation was mainly substituted by coffee crops and pastures. Now days, the landscape is highly fragmented, and is composed by several secondary forest fragments, intermingled with pasture, coffee and *Eucalyptus* crop, among others. We arbitrarily chose five forest fragments and five neighboring *Eucalyptus* for our study sites.

### *4.1.2. Experimental design*

We used *Mabea fistulifera* (Euphorbiaceae) seeds, an abundant native myrmecochoricous species with elaiosome. Seeds of this species have a diplochoric dispersion, primarily ballistic and secondarily mainly by ants. Seeds, collected directly from branches of several native trees, were obtained from Forest Seeds Laboratory of the Federal University of Viçosa two months before the experimental set up. Seeds had their elaiosomes preserved and were maintained in cold chamber at 20°C straight after natural dehiscence and kept until their use.

In each of the 10 sampling sites (5 native forests and 5 *Eucalyptus* crops), we set 10 sampling units, which were distributed 10 meters apart from each other. Each sampling unit consisted of one ant sampling point and one seed removal spot, distanced 2 meters from each other. Ant sampling points consisted of unbaited pitfall traps (diameter 8 cm, 12 cm height), buried at soil level, and half filled with a killing solution of water, detergent and salt. Seed removal spots consisted of the provision of 10 seeds of *M. fistulifera*, which were covered by a cage with a mesh of 10 mm, to avoid seed removal by vertebrates [192]. Both pitfall traps and seeds remained in the field for 48 hours. After that, we counted the number of remaining seeds and the number and identity of ant species.

We identified ants to genera using the keys by [192], and when possible to species by com‐ parisons with the reference collection of the Community Ecology Lab/UFV, where voucher specimens were deposited. Species identification was confirmed by a specialist.

### **4.2. Statistical analyses**

natural habitat conversion into *Eucalyptus* crops and the consequences for seed removal by ants. We hypothesized that higher ant species richness increases seed removal, and this relationship is more pronounced in natural forest ecosystems than in *Eucalyptus* crops. Furthermore, we hypothesized that ant species composition changes between natural and *Eucalyptus* crops, and this shift is also responsible for supposed differential seed removal rates. Finally, as we observed differential seed removal between studied habitats we tested the

We carried out the study in Viçosa, Minas Gerais (20°45'S, 42°50'W), Brazil, during summer 2010/2011, the rainy season. The pristine vegetation in this region is within the Atlantic Forest Domain, and is classified as Seasonal Semi-deciduous Forest. From the 1930's decade an intense fragmentation process has begun, and the native vegetation was mainly substituted by coffee crops and pastures. Now days, the landscape is highly fragmented, and is composed by several secondary forest fragments, intermingled with pasture, coffee and *Eucalyptus* crop, among others. We arbitrarily chose five forest fragments and five neighboring *Eucalyptus* for

We used *Mabea fistulifera* (Euphorbiaceae) seeds, an abundant native myrmecochoricous species with elaiosome. Seeds of this species have a diplochoric dispersion, primarily ballistic and secondarily mainly by ants. Seeds, collected directly from branches of several native trees, were obtained from Forest Seeds Laboratory of the Federal University of Viçosa two months before the experimental set up. Seeds had their elaiosomes preserved and were maintained in

In each of the 10 sampling sites (5 native forests and 5 *Eucalyptus* crops), we set 10 sampling units, which were distributed 10 meters apart from each other. Each sampling unit consisted of one ant sampling point and one seed removal spot, distanced 2 meters from each other. Ant sampling points consisted of unbaited pitfall traps (diameter 8 cm, 12 cm height), buried at soil level, and half filled with a killing solution of water, detergent and salt. Seed removal spots consisted of the provision of 10 seeds of *M. fistulifera*, which were covered by a cage with a mesh of 10 mm, to avoid seed removal by vertebrates [192]. Both pitfall traps and seeds remained in the field for 48 hours. After that, we counted the number of remaining seeds and

We identified ants to genera using the keys by [192], and when possible to species by com‐ parisons with the reference collection of the Community Ecology Lab/UFV, where voucher

specimens were deposited. Species identification was confirmed by a specialist.

cold chamber at 20°C straight after natural dehiscence and kept until their use.

presence of keystone species would influence removal rates.

1464 Biodiversity in Ecosystems - Linking Structure and Function

**4.1. Methods**

*4.1.1. Study site*

our study sites.

*4.1.2. Experimental design*

the number and identity of ant species.

To test the relationship between seed removal and ant species richness we used an ANCOVA, in which the response variable was the proportion of removed seeds in each site, and the explanatory variables were ant species richness and environment type (native forests or *Eucalyptus* crops). As the response variable was a ratio, we used a binomial error distribution, corrected for overdispersion when necessary. We performed this analysis in the software R [92] and we did residual analysis to check for model fit and distribution suitability.

To test whether ant species composition changes between studied habitats, we performed a NMDS (Non-metric Multidimensional Scaling), using Bray-Curtis dissimilarity index. We computed species abundances as the number of traps they occurred in each sampling site. The significance of differences was checked through PERMANOVA [193]. This analysis was performed in the package vegan within the software R [92]. We tested if the most frequent ant species act as keystone species [194] in seed removal by an ANOVA, in which we compared seed removal in the presence and in the absence of each species.

We removed one of the *Eucalyptus* areas a priory from all the analyses due to a heavy rain that removed all seeds, and another *Eucalyptus* area from the ANCOVA after the residual analysis as it was considered an outlier, therefore reducing our total sampling units to eight (five native and three *Eucalyptus* crops).

### **4.3. Results**

We sampled 43 ant species, from 25 genera and seven subfamilies. From these, 23 species occurred exclusively in the native forests, five were exclusive from *Eucalyptus* and 15 occurred in both. The most frequent species were *Pheidole radoskowskii* Mayr, 1884 and *Ectatomma muticum* Mayr, 1870, which occurred in 40.24% and 30.49% of pitfall traps, respectively.

As expected, ant species richness was higher in native forest than in *Eucalyptus* crops (χ<sup>2</sup> =6.93; p=0.008). Moreover, seed removal rate increased with the number of ant species (F1,6=11.01; p=0.021; Fig. 4), however it was higher in the *Eucalyptus* than in the native forest (F1,6=8.75; p=0.032). Conversely, species composition did not differ between the two habitats (Fig. 5, PERMANOVA F1,7=1.12, p=0.32). Neither the presence of *P. radoskowskii* (F1,80=0.87; p=0. 35), more frequent in *Eucalyptus*, nor of *E. muticum* (χ<sup>2</sup> =0.94; p=0.33), more frequent in native forests, influenced seed removal.

## **5. Discussion**

Differences in species richness, abundance and composition may affect ecosystem functioning [195]. In this case study we investigated the role of these three biodiversity components on seed removal by ants in native and Eucalyptus forests. Concerning species richness our results confirm the general pattern of reduction in modified habitats. The main causes reported for such pattern include habitat loss, homogenization, and harshness conditions for native species [196]. In comparison with native forest, *Eucalyptus* crops may be homogeneous habitats, which might have contributed to its lower species richness. Ant species richness is strictly related with environmental features such as higher plant species diversity, litter amount and habitat complexity [142,197-200]. From these, plant species diversity and habitat complexity decrease in *Eucalyptus* crops, which may have caused the loss of ant species that did not survive in the modified habitat. We observed the expected positive relationship between seed removal and ant species richness, both in the native forests and *Eucalyptus* crops. Nevertheless, the maximal seed removal at *Eucalyptus* was around 30% while at native forest was about 65%, which may be related to the smaller capacity of *Eucalyptus* crops of harboring species when compared to native forests. This pattern could also be attributed to keystone species (sensu [201]) at native forest, thus promoting the observed higher removal rates. However, the sole effect of potential keystone species did not explain the rates we observed, as seed removal did not change in their absence. Therefore, we have no evidence to consider the existence of some specialist seed remover species inhabiting either of the environments, reinforcing the role of ant species richness in the studied process.

On the other hand we did not find differential species composition between the two habitats types, thus we cannot assign the higher seed removal at native forest due to some keystone species. Moreover, seed removal rates at native forest did not differ when we analyzed the effects of the presence of the most abundant ant species (*E. muticum*). Therefore, we conclude that species richness is the only biodiversity component influencing the ecosystem process in the studied system. The positive relationship between ant species richness and seed removal rate may have important concerns on conservation. The maintenance of natural species richness levels can contribute to a suitable ecosystem functioning due to the role of the seed dispersal for seedling establishment and the community assembly.

**Figure 4.** Seed removal rates increased with ant species richness (F1,6=11.01; p=0.021), and were higher in *Eucalyptus* crop.

Biodiversity and Ecosystem Functioning in Tropical Habitats — Case Studies and Future Perspectives… 17 http://dx.doi.org/10.5772/59042 67

**Figure 5.** NMDS map of species composition according to treatment (Native or *Eucalyptus* crop). We analyzed signifi‐ cance by using Permanova test, which was non-significant.

### **6. Conclusions and perspectives**

might have contributed to its lower species richness. Ant species richness is strictly related with environmental features such as higher plant species diversity, litter amount and habitat complexity [142,197-200]. From these, plant species diversity and habitat complexity decrease in *Eucalyptus* crops, which may have caused the loss of ant species that did not survive in the modified habitat. We observed the expected positive relationship between seed removal and ant species richness, both in the native forests and *Eucalyptus* crops. Nevertheless, the maximal seed removal at *Eucalyptus* was around 30% while at native forest was about 65%, which may be related to the smaller capacity of *Eucalyptus* crops of harboring species when compared to native forests. This pattern could also be attributed to keystone species (sensu [201]) at native forest, thus promoting the observed higher removal rates. However, the sole effect of potential keystone species did not explain the rates we observed, as seed removal did not change in their absence. Therefore, we have no evidence to consider the existence of some specialist seed remover species inhabiting either of the environments, reinforcing the role of ant species

On the other hand we did not find differential species composition between the two habitats types, thus we cannot assign the higher seed removal at native forest due to some keystone species. Moreover, seed removal rates at native forest did not differ when we analyzed the effects of the presence of the most abundant ant species (*E. muticum*). Therefore, we conclude that species richness is the only biodiversity component influencing the ecosystem process in the studied system. The positive relationship between ant species richness and seed removal rate may have important concerns on conservation. The maintenance of natural species richness levels can contribute to a suitable ecosystem functioning due to the role of the seed

**Figure 4.** Seed removal rates increased with ant species richness (F1,6=11.01; p=0.021), and were higher in *Eucalyptus*

dispersal for seedling establishment and the community assembly.

richness in the studied process.

1666 Biodiversity in Ecosystems - Linking Structure and Function

crop.

Although a positive relationship between biodiversity and ecosystem functioning is common‐ ly reported [28, 202], we did not find such evidence from all the studies presented here. The main results presented allow us to conclude that the general effect of arthropods on ecosystem functioning is dependent on the studied process and the proximity with their agents. As more direct is the action of arthropods on the ecosystem processes, more detectable are their effects on functioning.

Arthropods contribute indirectly in the process of litter decomposition by modifying the substrate to the decomposers (the microbiota), besides acting through predation in a top-down effect on these microorganisms. Therefore, their indirect effect on litter decomposition may have produced the lack of relationship between their biodiversity and ecosystem functioning. Similarly, in the second case study we observed an absence of ant species richness and plant defense against herbivory. Nevertheless, in this study we noticed that the abundance of ants partially resulted in a decrease of the abundance of herbivore insects. The results obtained in this case study may reflect the plenty of defense mechanisms against herbivory, and ants are only a further mechanism within several others by which plants achieve a better protection. In the third case study we could evaluate a process that directly involved the importance of biodiversity on ecosystem processes, as there is a direct interaction between ants and the seeds they remove, without intermediate agents between them. Therefore, we could notice that the effect of ant biodiversity on the ecosystem process was stronger when compared to the two previous case studies. The above rationale points that a greater proximity between the agent and process turns the relationship stronger and detectable.

Based on the studies presented here, we suggest the following steps to improve the studies of the relationship between biodiversity and ecosystem functioning. Firstly, the control of variables through manipulative approaches should be increased, as confounding variables might decrease the chance of unveiling significant relationships. Secondly, as described above, it should be investigated relationships in which the processes and their agents are more directly connected. As reported elsewhere [28], the effects on productivity decrease with the increase of the number of trophic levels between manipulated (biodiversity) and estimated (ecosystem process) elements. Finally, studies of less complex systems may produce stronger results, once in complex systems several agents may influence concomitantly a given process, decreasing the chance of detecting a relationship between biodiversity and ecosystem functioning. Our second case study is an example, as several agents may influence plant herbivory, besides the presence of a higher ant species richness and abundance. Similar conclusions have been found in a meta-analysis study involving several results obtained from different regions of the World [28]. Hence, the comparison among our results and those obtained by other authors indicate that, despite the high complexity and biodiversity found in tropical regions, the trends reported here are comparable to those found worldwide.

This chapter integrated different case studies relating biodiversity and ecosystem functioning, with varying degrees of proximity between the agents and the processes. Because human activities would certainly continue to produce loss of species, we suggest that future studies relating biodiversity and ecosystem processes consider the linkage among the agents involved in the processes, to improve the understanding of this relationship, as well as the prognosis involving changes in biodiversity.

### **Acknowledgements**

Authors are indebted to Lucas G. Dornelas for providing his data of case study 3. Authors received grants from Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq – Brazil), Fundação Capes – Brazil, and Fundação de Amparo à Pesquisa do Estado de Minas Gerais (FAPEMIG).

## **Author details**

Tathiana G. Sobrinho, Lucas N. Paolucci, Dalana C. Muscardi, Ana C. Maradini, Elisangela A. Silva, Ricardo R. C. Solar and José H. Schoereder\*

\*Address all correspondence to: jschoere@ufv.br

Departamento de Biologia Geral, Universidade Federal de Viçosa. Viçosa, MG, Brazil

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previous case studies. The above rationale points that a greater proximity between the agent

Based on the studies presented here, we suggest the following steps to improve the studies of the relationship between biodiversity and ecosystem functioning. Firstly, the control of variables through manipulative approaches should be increased, as confounding variables might decrease the chance of unveiling significant relationships. Secondly, as described above, it should be investigated relationships in which the processes and their agents are more directly connected. As reported elsewhere [28], the effects on productivity decrease with the increase of the number of trophic levels between manipulated (biodiversity) and estimated (ecosystem process) elements. Finally, studies of less complex systems may produce stronger results, once in complex systems several agents may influence concomitantly a given process, decreasing the chance of detecting a relationship between biodiversity and ecosystem functioning. Our second case study is an example, as several agents may influence plant herbivory, besides the presence of a higher ant species richness and abundance. Similar conclusions have been found in a meta-analysis study involving several results obtained from different regions of the World [28]. Hence, the comparison among our results and those obtained by other authors indicate that, despite the high complexity and biodiversity found in tropical regions, the trends

This chapter integrated different case studies relating biodiversity and ecosystem functioning, with varying degrees of proximity between the agents and the processes. Because human activities would certainly continue to produce loss of species, we suggest that future studies relating biodiversity and ecosystem processes consider the linkage among the agents involved in the processes, to improve the understanding of this relationship, as well as the prognosis

Authors are indebted to Lucas G. Dornelas for providing his data of case study 3. Authors received grants from Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq – Brazil), Fundação Capes – Brazil, and Fundação de Amparo à Pesquisa do Estado de Minas

Tathiana G. Sobrinho, Lucas N. Paolucci, Dalana C. Muscardi, Ana C. Maradini,

Departamento de Biologia Geral, Universidade Federal de Viçosa. Viçosa, MG, Brazil

Elisangela A. Silva, Ricardo R. C. Solar and José H. Schoereder\*

\*Address all correspondence to: jschoere@ufv.br

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1868 Biodiversity in Ecosystems - Linking Structure and Function

reported here are comparable to those found worldwide.

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**Acknowledgements**

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**Author details**


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## **Climate Change, Range Shifts and Multitrophic Interactions**

Jeffrey A. Harvey and Miriama Malcicka

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/59269

## **1. Introduction**

Climate change represents one of the most serious threats to biodiversity and ecosystem functioning. The current rate of temperature change, driven primarily by the human combus‐ tion of fossil fuels, far exceeds rates that have occurred in at least 10,000 years (lower Pleisto‐ cene) and perhaps much longer (IPCC, 2014). That last major climate change event precipitated a mass extinction that led to the sudden demise of many large quadrupeds, including such characteristic species as the woolly mammoth, woolly rhinoceros, mastodon, giant elk, sabertoothed tiger and dire wolf [1]. One of the major differences between landscapes at the time of previous climate change events and the current one is that the biosphere is now dominated by a single species, *Homo sapiens sapiens*, which has profoundly altered and simplified many terrestrial and aquatic ecosystems. Thus, in addition to climate change, natural ecosystems have been altered by other human-induced changes including deforestation, eutrophication, over-harvesting, the introduction of non-native species and various types of pollution. Consequently, species and populations are being challenged by multiple stressors, making it more difficult for them to adapt to rapid shifts in climate regimes. One can strongly argue that we no longer live in the Holocene but in the Anthropocene [2,3].

In a warming world, many species and populations are responding by changing various aspects of their life cycles, such as seasonal growth and phenology patterns, as well as by shifting their ranges pole-wards and/or to higher elevations [4,5,6,7]. The ability of species to shift their distributions is often limited by various eco-physiological constraints. These include the loss of habitat corridors through urban and agricultural expansion, which enable species to disperse over landscapes to other suitable habitat patches. Furthermore, some species lack traits, such as wings, which enable them to easily track changes in biotic conditions. As a result of these factors, we can expect ecological communities to change over time and for this to lead

© 2015 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2014 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and eproduction in any medium, provided the original work is properly cited.

to unpredictable new assemblages which may or may not stabilize as temperatures continue to rise [8].

## **2. Climate change and multitrophic interactions**

It has long been known that species do not exist in isolation in nature. Instead, the survival and persistence of species in food webs and communities is dependent upon an array of interactions with other organisms occurring over highly variable spatial and temporal scales. Indeed, ecologist Daniel Janzen [8] once remarked that 'the ultimate extinction is the extinction of ecological interactions'. More recently, Pimm and Raven [9] argued that for every species of plant that becomes extinct in the tropics, many tens of species dependent on that plant for food or shelter also disappear. Given that different species in food webs may respond differ‐ ently to climate change, warming has the potential of unravelling and/or destabilizing plantinsect communities, and that these can trickle through to affect other trophic interactions, even involving vertebrates [10]. In this chapter we discuss the effects and potential consequences of warming on trophic interactions involving plants, insect herbivores and specialist natural enemies, focusing on parasitic wasps (or parasitoids). Parasitoids are insects that develop on or in the bodies of other insects, whereas the adults are free living [11]. Hosts attacked by parasitoids are often not much larger than the adult parasitoid, meaning they are under intense selection to allocate and utilize limited host resources for different and often competing fitness functions such as reproduction and survival [12]. It is well established that many parasitoids are often highly specialized in attacking only one or a few species of hosts in nature [11]. Warming-induced changes in the environment may therefore affect various aspects of the biology and ecology of food plants and insect herbivores dependent on them, and this may trickle up the food chain, in particular affecting natural enemies that are more specialized on certain host types. Parasitoids thus make model organisms for examining a range of biotic and abiotic constraints in the environment.

### **2.1. Outline of the chapter**

The chapter will be broken down into separate sections examining the effects of warming on the biology and ecology of the three trophic levels separately, and then move on to integrate these interactions and to provide testable predictions for these processes. Given that insects are ecto-therms, it is by now well established that metabolic rate and the developmental program of insects is closely co-ordinated with changes in temperature. However, it is less well established how changes in temperature, as well as attendant changes in precipitation etc. will affect tightly linked two and three-trophic level interactions. On this basis, the chapter will be broken down thusly:

**1.** Range shifts in plants and effects to (i) primary [nutrients] (ii) secondary [defensive compounds] metabolites, as well as in plant volatiles under herbivore damage (HIPVs). How will changes in plant quality affect multitrophic interactions?

Climate Change, Range Shifts and Multitrophic Interactions 3 http://dx.doi.org/10.5772/59269 87

to unpredictable new assemblages which may or may not stabilize as temperatures continue

It has long been known that species do not exist in isolation in nature. Instead, the survival and persistence of species in food webs and communities is dependent upon an array of interactions with other organisms occurring over highly variable spatial and temporal scales. Indeed, ecologist Daniel Janzen [8] once remarked that 'the ultimate extinction is the extinction of ecological interactions'. More recently, Pimm and Raven [9] argued that for every species of plant that becomes extinct in the tropics, many tens of species dependent on that plant for food or shelter also disappear. Given that different species in food webs may respond differ‐ ently to climate change, warming has the potential of unravelling and/or destabilizing plantinsect communities, and that these can trickle through to affect other trophic interactions, even involving vertebrates [10]. In this chapter we discuss the effects and potential consequences of warming on trophic interactions involving plants, insect herbivores and specialist natural enemies, focusing on parasitic wasps (or parasitoids). Parasitoids are insects that develop on or in the bodies of other insects, whereas the adults are free living [11]. Hosts attacked by parasitoids are often not much larger than the adult parasitoid, meaning they are under intense selection to allocate and utilize limited host resources for different and often competing fitness functions such as reproduction and survival [12]. It is well established that many parasitoids are often highly specialized in attacking only one or a few species of hosts in nature [11]. Warming-induced changes in the environment may therefore affect various aspects of the biology and ecology of food plants and insect herbivores dependent on them, and this may trickle up the food chain, in particular affecting natural enemies that are more specialized on certain host types. Parasitoids thus make model organisms for examining a range of biotic and

The chapter will be broken down into separate sections examining the effects of warming on the biology and ecology of the three trophic levels separately, and then move on to integrate these interactions and to provide testable predictions for these processes. Given that insects are ecto-therms, it is by now well established that metabolic rate and the developmental program of insects is closely co-ordinated with changes in temperature. However, it is less well established how changes in temperature, as well as attendant changes in precipitation etc. will affect tightly linked two and three-trophic level interactions. On this basis, the chapter

**1.** Range shifts in plants and effects to (i) primary [nutrients] (ii) secondary [defensive compounds] metabolites, as well as in plant volatiles under herbivore damage (HIPVs).

How will changes in plant quality affect multitrophic interactions?

**2. Climate change and multitrophic interactions**

862 Biodiversity in Ecosystems - Linking Structure and Function

abiotic constraints in the environment.

**2.1. Outline of the chapter**

will be broken down thusly:

to rise [8].

Variation in responses at the species level makes it virtually impossible to predict the effects of climate change on trophic interactions, although available evidence suggests that there will be many more losers than winners and as a result we can expect ecosystem interactions to be simplified and ecosystems therefore to become less stable and resilient. **Figure 1.** Phenological interactions involving three plant species, a specialist herbivore and its specialist endoparasi‐ toid under normal conditions and under climate warming scenarios in Europe. In (A), *Pieris brassicae* (middle) the large cabbage white butterfly is trivoltine, and different generations lay their eggs on different species of large, short-lived annual brassicaceous (mustard) plants (bottom) that grow in aggregated populations. Different stages of the herbivore – eggs, young larvae, fully grown larvae, pupa and adult are shown. The mustards in turn grow only for 2-3 months during the year and at different times. *Brassica rapa* (wild turnip, left, green line) supports the first generation of *P. brassicae*, *Sinapsis arvensis* (charlock mustard, middle, red line) the second, and *Brassica nigra* (black mustard, right, blue line) the third. In turn, the gregarious endoparasitoid wasp, *Cotesia glomerata* (top) also has three generations where the adult wasps emerge in time to find and parasitize young caterpillars of *P. brassicae*, and emerge from fully grown cater‐ pillars pupating on the food plant. In (B) and (C), warmer temperatures lead to dissociation of the multi-trophic inter‐ action by differentially affecting the seasonal growth and activity patterns of the plants and insects. In (B), the plants are more strongly affected and their growing phenology is shifted to an earlier point in the season. However, the in‐ sects respond less to warming and their temporal life cycles become desynchronized with the growth cycles of their plants. There is also a period later in the season when few or no suitable plants are available. In (C), the plants only marginally shift their growing patterns to an earlier time point in the season, whereas the insects have an additional (=4th) generation later in the year. In both climate warming scenarios, there are gaps for the insects when food plants are scarce or absent. This could have profoundly negative consequences on the persistence of this trophic chain.

(c)

6


An important theme of this chapter is to stress the importance of examining multi-species interactions under climate change scenarios. This entails a close examination of mechanisms as well as the consequences of warming in a community-related context. Thus far, both climate change and invasion ecology have been studied independently or have been largely restricted to the study of bi-trophic interactions (e.g. plant-herbivore) with very few studies going to three (or more) trophic interactions [13]. There are a number of excellent reviews which focus on each of these areas [14,15,16,17] but only a few have begun to address community leveleffects, and again these generally do not integrate climate change with range shifts in plants and insects [18,19,20]. Fewer still have explored climate-change related range shifts on plantherbivore-parasitoid interactions at both reductionist and larger scale perspectives. The main aim of this chapter is therefore to explore and discuss how climate change and related abiotic changes (e.g. precipitation) will affect multi-trophic interactions from both the perspective of mechanisms and processes, scaling up from individuals to communities and ecosystems. We will discuss the potentially cascading effects of climate warming and other parameters related to climate change on ecosystems, suggesting that many will be simplified (Fig 1.), reducing their resilience against other natural and anthropogenic challenges in the environment.

### **3. Range shifts in plants**

Two aspects of warming will affect the structure of plant communities. The first involves native plants which may be adapted to cooler conditions and thus become increasingly stressed as conditions warm. This is especially true for plant species growing at the southern edge of their ranges. The second involves plants which shift their ranges northwards in order to track warming and to exploit more optimal thermal conditions. These plants will increasingly expand into habitats occupied by native plants, leading to potential increases in competition and, at least to some extent, rearrangement in plant community structure. As the climate warms, we may expect physiological and metabolic responses in both native and range expanding plants that will in turn potentially affect the behaviour and performance of higher trophic level consumers associated with them. More specifically, the metabolic changes in plants may be borne out on traits such as primary and secondary metabolism which play a key role in insect nutrition and plant defence responses. Plant volatiles, which are often induced in response to herbivore damage and which potentially have many functions, might also be affected by warming or in response to new selection pressures in range expanding plants. At the same time, range expanding plants may, at least partially, escape from some of their co-evolved enemies, such as herbivores and pathogens, also affecting the costs and benefits of metabolic investment into plant defences. Below, we examine these areas in more detail.

## **4. Primary and secondary metabolism**

**2.** Range shifts and outbreaks in herbivores in response to climate change and effects on plant responses and interactions with their natural enemies; range shifts in parasitoids: tracking their hosts or switching to novel native hosts? Examples of competitive interac‐

**3.** The effects of temperature on insect growth, development and other traits. Given that insects are ectothermic, how will higher temperatures affect their development, and what

**4.** Changes in the seasonal phenology of multitrophic interactions; does warming differen‐ tially affect the growth/life cycles of plants, herbivores and their natural enemies, and

**5.** How will warming potentially affect the dynamics of multi-trophic interactions, scaling up to food webs, communities and ecosystems, and will this in turn affect resilience and

An important theme of this chapter is to stress the importance of examining multi-species interactions under climate change scenarios. This entails a close examination of mechanisms as well as the consequences of warming in a community-related context. Thus far, both climate change and invasion ecology have been studied independently or have been largely restricted to the study of bi-trophic interactions (e.g. plant-herbivore) with very few studies going to three (or more) trophic interactions [13]. There are a number of excellent reviews which focus on each of these areas [14,15,16,17] but only a few have begun to address community leveleffects, and again these generally do not integrate climate change with range shifts in plants and insects [18,19,20]. Fewer still have explored climate-change related range shifts on plantherbivore-parasitoid interactions at both reductionist and larger scale perspectives. The main aim of this chapter is therefore to explore and discuss how climate change and related abiotic changes (e.g. precipitation) will affect multi-trophic interactions from both the perspective of mechanisms and processes, scaling up from individuals to communities and ecosystems. We will discuss the potentially cascading effects of climate warming and other parameters related to climate change on ecosystems, suggesting that many will be simplified (Fig 1.), reducing their resilience against other natural and anthropogenic challenges in the environment.

Two aspects of warming will affect the structure of plant communities. The first involves native plants which may be adapted to cooler conditions and thus become increasingly stressed as conditions warm. This is especially true for plant species growing at the southern edge of their ranges. The second involves plants which shift their ranges northwards in order to track warming and to exploit more optimal thermal conditions. These plants will increasingly expand into habitats occupied by native plants, leading to potential increases in competition and, at least to some extent, rearrangement in plant community structure. As the climate warms, we may expect physiological and metabolic responses in both native and range

stability? What are the possible effects on top-down and bottom-up regulation?

what are the consequences for the persistence of these interactions?

tions and potential displacement in interspecific competition will be explored.

will the consequences be for biotic interactions?

884 Biodiversity in Ecosystems - Linking Structure and Function

**3. Range shifts in plants**

Plant tissues contain various concentrations of nutrients, with carbon (C), nitrogen (N), and phosphorus (P) being considered as the most important, as well as amino acids [21]. Although atmospheric CO2 concentrations are deterministic at various scales, and are rising slowly at about 2-3 ppm annually, temperatures are not. Thus one can question the efficacy of experi‐ ments that expose plants (and insects) to extremely high concentrations of CO2 (e.g. 450 ppm and higher) when these levels are not expected to be reached for several decades. However, there is no way to successfully circumvent this obstacle, and many studies exposing plants to higher CO2 levels have been conducted. Many have shown that under increased CO2 regimes attendant with climate change that plants will possibly take up more C at the potential expense of foliar levels of N and P [22,23,24] and that changes in concentrations of other nutrients, such as amino acids, will also occur [25]. (However, these processes are indeed complex – for instance they vary in plants with different metabolic C pathways as well as in different successional stages but a highly detailed discussion of this is not possible here). N is considered a limiting nutrient for insect development [26,27,28,29] and it has been suggested that a reduction in foliar concentrations of N will lead to compensatory feeding in herbivores to ensure optimal levels of this nutrient are acquired [26,30,14,31]. For range expanding plants in a warming world, it is difficult to extrapolate cause-and-effect relationships related to temperature and precipitation, unless atmospheric changes in CO2 are also taken into account.

Most plants also produce secondary metabolites (or allelochemicals) whose primary function is considered to be defence against antagonists such as pathogens and herbivores [32,33,34,35]. There is a vast array of different types of allelochemicals produced by plants that are based on the phylogeny of a species group, often at the family level [36]. For instance, one can find glucosinolates, alkaloids, iridoid glycosides, furanocoumarins and many other allelochemicals produced by plants in different families. These toxins are often inducible, meaning that they are found in basal levels in intact plant tissues but markedly increase after tissue damage [37, 35]. These compounds have long been known to affect the behaviour and performance of insects that are intimately associated with the plants that produce them. High levels of phytotoxins in plant tissues can impair the development and hence fitness of herbivores, through increased rates of mortality, extended development time and reduced adult size [38, 39,40,41,42,43,44]. These effects often differ between specialist and generalist herbivores. Generalists usually do not possess highly refined mechanisms of dealing with specific secondary plant compounds, but instead rely on general responses, (e.g. P-450 mono-oxyger‐ nases) that are effective against a range of different metabolites [45]. However, they are very often less effective than the mechanisms employed by specialists which have strongly coevolved histories with certain types of plant phytotoxins. For this reason, specialists are frequently assumed to be in a co-evolutionary 'arms race' with plant defences [46] in which increased herbivory leads to increased plant defences which in turn leads to adaptations on the part of the herbivore and so on in a 'Red Queen' type stalemate scenario. In fact, elevated levels of toxins actually stimulate feeding and oviposition behaviour in the many specialist herbivores [35]. Despite this, very well defended plants are often problematical even for specialists [47].

The development of natural enemies such as parasitoid wasps is also known to be affected by host attributes as mediated through the quality of the host diet [41,42]. High levels of plant toxins ingested by the host can also detrimentally affect the growth and survival of immature parasitoids and even hyperparasitoids one trophic level higher [48,49]. One major phyloge‐ netic constraint on parasitoid larvae is that the alimentary tract is not externally connected until they terminate the relationship with the host [50]. Therefore, all host tissues ingested during development are stored in the body pending the voiding of the meconium (or even later as an adult). This includes deleterious materials such as plant toxins that were stored in the body of the herbivore host. If concentrations of these toxins become too high, then it is possible that the developing parasitoid larvae cannot safely store them and will die preco‐ ciously as a result. It is interesting that the diversity of many parasitoid taxa is higher in temperate than in tropical biomes, a trend that contradicts that shown in most other groups of insects. Bolton [51] hypothesized that tropical vegetation is generally much more toxic than vegetation in the temperate zone, and that parasitoids have not been able to adapt to hosts that ingest and store these toxins in their body tissues (the 'nasty-host-hypothesis').

Several climate-change related effects on plant allelochemistry are possible, and these may trickle their way through trophic webs. Elevated CO2 regimes will also affect plant allelo‐ chemistry in potentially different directions depending upon whether a plant's defence chemistry is C-or N-based. Plants with C-based allelochemistry may become more toxic (e.g. better defended) whereas the opposite may occur with plants with N-based allelochemistry [52,53,54,55,56,57]. A second affect will be that warming, along with changes in precipitation, as well as reduced or extended drought periods attendant with climate change will affect metabolic allocation by plants to secondary defense metabolites. Higher temperatures can sometimes lead to a reduction in plant allocation to direct defence [58]. Veteli [59] found that elevated CO2 and temperature increased plant growth but had opposite effects on the growth rate of an insect herbivore. A third effect will be that range shifting plants will escape from some of their co-evolved herbivore enemies, leading to changes in metabolic allocation towards plant defence [60]. If native specialist and generalist herbivores select for high chemical defences in native plants, then range-expanding plants may invest less in these defences if they are attacked by fewer herbivores (and thus suffer less herbivore damage) in their new range [the 'enemy-release hypothesis' [ERH],[61,62,63] and more towards other functions, such as growth that enable plants to outcompete natives (the 'evolution of increased competitive ability hypothesis' [EICA], [62,64]. A third hypothesis, the 'shifting defence hypothesis' [SDH], posits that co-evolved specialists are adapted to high concentrations of plant allelochemicals, driving selection for a reduction in them to make them less attractive as oviposition sites [65]. In new habitats, these plants escape their specialists but attract more generalists, selecting for an increase in chemical defences. Lastly, many invasive plants bring with them novel secondary metabolites to which the native herbivore fauna is not adapted and which therefore allows the plant to invest metabolic energy to other vital functions (the 'novel weapons hypothesis' [NWH], [66,67]. Thus only a single study has considered climate change-mediated changes in plant communities on natural enemies of the herbivores on native and invasive plants [68]. This revealed that there may be shifts in the intensity of natural enemy-mediated top down trophic cascades versus bottom-up plant-mediated effects on herbivores on natives and invasives. However, it is far too early to draw conclusions as many more studies are needed to tease out potential patterns in communities where shifts in plant composition are occurring as a result of warming.

39,40,41,42,43,44]. These effects often differ between specialist and generalist herbivores. Generalists usually do not possess highly refined mechanisms of dealing with specific secondary plant compounds, but instead rely on general responses, (e.g. P-450 mono-oxyger‐ nases) that are effective against a range of different metabolites [45]. However, they are very often less effective than the mechanisms employed by specialists which have strongly coevolved histories with certain types of plant phytotoxins. For this reason, specialists are frequently assumed to be in a co-evolutionary 'arms race' with plant defences [46] in which increased herbivory leads to increased plant defences which in turn leads to adaptations on the part of the herbivore and so on in a 'Red Queen' type stalemate scenario. In fact, elevated levels of toxins actually stimulate feeding and oviposition behaviour in the many specialist herbivores [35]. Despite this, very well defended plants are often problematical even for

The development of natural enemies such as parasitoid wasps is also known to be affected by host attributes as mediated through the quality of the host diet [41,42]. High levels of plant toxins ingested by the host can also detrimentally affect the growth and survival of immature parasitoids and even hyperparasitoids one trophic level higher [48,49]. One major phyloge‐ netic constraint on parasitoid larvae is that the alimentary tract is not externally connected until they terminate the relationship with the host [50]. Therefore, all host tissues ingested during development are stored in the body pending the voiding of the meconium (or even later as an adult). This includes deleterious materials such as plant toxins that were stored in the body of the herbivore host. If concentrations of these toxins become too high, then it is possible that the developing parasitoid larvae cannot safely store them and will die preco‐ ciously as a result. It is interesting that the diversity of many parasitoid taxa is higher in temperate than in tropical biomes, a trend that contradicts that shown in most other groups of insects. Bolton [51] hypothesized that tropical vegetation is generally much more toxic than vegetation in the temperate zone, and that parasitoids have not been able to adapt to hosts that

ingest and store these toxins in their body tissues (the 'nasty-host-hypothesis').

Several climate-change related effects on plant allelochemistry are possible, and these may trickle their way through trophic webs. Elevated CO2 regimes will also affect plant allelo‐ chemistry in potentially different directions depending upon whether a plant's defence chemistry is C-or N-based. Plants with C-based allelochemistry may become more toxic (e.g. better defended) whereas the opposite may occur with plants with N-based allelochemistry [52,53,54,55,56,57]. A second affect will be that warming, along with changes in precipitation, as well as reduced or extended drought periods attendant with climate change will affect metabolic allocation by plants to secondary defense metabolites. Higher temperatures can sometimes lead to a reduction in plant allocation to direct defence [58]. Veteli [59] found that elevated CO2 and temperature increased plant growth but had opposite effects on the growth rate of an insect herbivore. A third effect will be that range shifting plants will escape from some of their co-evolved herbivore enemies, leading to changes in metabolic allocation towards plant defence [60]. If native specialist and generalist herbivores select for high chemical defences in native plants, then range-expanding plants may invest less in these defences if they are attacked by fewer herbivores (and thus suffer less herbivore damage) in

specialists [47].

906 Biodiversity in Ecosystems - Linking Structure and Function

There is ample evidence in support of these different hypotheses for inter-continental inva‐ sions [63,69]. However, intra-continental invasions based on climate-change related range shifts [70,71] are likely to have more subtle effects on plant-consumer interactions because many insects can track their food plants as they move pole-wards [72]. As a result, consistent patterns remain elusive. Engelkes [68] found that range expanding plants in the Netherlands were more toxic to naïve herbivores than related natives, and that the range-expanders also produced higher levels of general defence compounds (phenolics). Fortuna [73] compared allelochemistry and herbivore performance in native and range-expanding populations of warty cabbage (*Bunias orientalis*) and reported that the insects were larger and survived better on the range-expanding than on the native populations. Moreover, the authors found both quantitative and qualitative differences in plant allelochemistry between the native and range expanding plants. It appears that *B. orientalis* is comparatively rare and local in its native range, which is perhaps evidence of biotic resistance amongst coevolved members of the native plant community as well as more top-down control exerted by native co-evolved herbivores and/or pathogens [61,64]. Other studies have found little or no differences in insect performance on native and range-expanding populations in plants [74].

Plants also release chemically-based odours via the production of volatiles that are often induced by stresses such as herbivory [75,76,77]. The precise function of these volatiles is unclear and remains the subject of considerable debate [78,79,80,81]. It is known that these volatiles are used by insects to locate plants on which to oviposit [75,82] or on which natural enemies find their prey or hosts [83,84,85,86]. If insects can drive selection for types and concentrations of volatiles, then changes in top-down pressures as plants are released from their enemies (e.g. herbivores) or potential allies (e.g. parasitoids) may also alter volatile profiles over time. However, thus far the effect of plant volatiles in range expanding plants on native natural enemies has received virtually no attention. Fortuna [87] found that adult females of the large cabbage white butterfly, *Pieris brassicae*, preferred to oviposit on shoots of a native brassicaceous plant, *Sinapis arvensis*, over those of a range expander, *Bunias orientalis*. Herbivore performance was higher on the native plant. However, its major natural enemy, the parasitoid *Cotesia glomerata*, did not distinguish between volatiles of the two plants. This suggests that herbivores and parasitoids may respond to different kinds of volatiles – herbi‐ vores clearly on those released by the plant alone and specialist natural enemies to a combi‐ nation of plant and host-or prey-related odours.

A recent study by Yuan [88] laid out a possible framework for the effects of climate change on plant volatiles, and projected this to the community level. They argued that patterns are likely to be variable and association-specific, because of the immense complexity in the blends that make up a plant's volatile profile. Moreover, other anthropogenic stresses must be factored in, leading to highly unpredictable scenarios across different scales of space and time. They urged that more studies with different systems are necessary to tease out mechanisms that may cascade up to affect higher trophic level consumers.

## **5. Effects of climate change on insect growth and development**

Insects, like all invertebrates, are ectotherms and thus they are highly susceptible to changes in temperature as well as other abiotic processes linked with climate change [14,15]. Many studies have examined how insects across different trophic levels respond to variations in temperature [14]. Rarely have these responses, however, been placed within the context of climate change, perhaps because warming has only been broadly acknowledged in the past 20 years or so by the scientific community. As with plants, insects can respond in two ways to local warming regimes. First, they must adapt behaviourally, morphologically and physio‐ logically to such processes as an increased incidence of heat waves and other attendant stresses such as droughts or higher precipitation regimes or an increase in frost-free periods [15]. Second they can shift their ranges and move pole-wards or to higher elevations (below).

Many studies have reported that the survival, development rate and adult body mass of insect herbivores and their natural enemies are affected by rearing temperatures [89]. Much less attention has been paid to transient periods of high or low temperatures (e.g. combining them under the umbrella of a single experiment, thus creating a more realistic picture of events transpiring in nature) or rainfall. It is by now known that climate warming is likely to generate more extreme weather events at local scales, rather than simply resulting in gradual changes that are measured at large spatial scales across the biosphere. When confronted with these conditions, insects have to adapt or to move to new habitats where conditions are more suitable. In time, the latter will result in range shifts, a phenomenon by now well described in many studies (see next section, below). Those insects that 'remain behind'will exhibit physio‐ logical responses to warming and phenomena associated with its such as an increase in heat waves and changes in precipitation.

## **6. Range shifts in herbivores and their natural enemies**

a native brassicaceous plant, *Sinapis arvensis*, over those of a range expander, *Bunias orientalis*. Herbivore performance was higher on the native plant. However, its major natural enemy, the parasitoid *Cotesia glomerata*, did not distinguish between volatiles of the two plants. This suggests that herbivores and parasitoids may respond to different kinds of volatiles – herbi‐ vores clearly on those released by the plant alone and specialist natural enemies to a combi‐

A recent study by Yuan [88] laid out a possible framework for the effects of climate change on plant volatiles, and projected this to the community level. They argued that patterns are likely to be variable and association-specific, because of the immense complexity in the blends that make up a plant's volatile profile. Moreover, other anthropogenic stresses must be factored in, leading to highly unpredictable scenarios across different scales of space and time. They urged that more studies with different systems are necessary to tease out mechanisms that may

Insects, like all invertebrates, are ectotherms and thus they are highly susceptible to changes in temperature as well as other abiotic processes linked with climate change [14,15]. Many studies have examined how insects across different trophic levels respond to variations in temperature [14]. Rarely have these responses, however, been placed within the context of climate change, perhaps because warming has only been broadly acknowledged in the past 20 years or so by the scientific community. As with plants, insects can respond in two ways to local warming regimes. First, they must adapt behaviourally, morphologically and physio‐ logically to such processes as an increased incidence of heat waves and other attendant stresses such as droughts or higher precipitation regimes or an increase in frost-free periods [15]. Second they can shift their ranges and move pole-wards or to higher elevations (below).

Many studies have reported that the survival, development rate and adult body mass of insect herbivores and their natural enemies are affected by rearing temperatures [89]. Much less attention has been paid to transient periods of high or low temperatures (e.g. combining them under the umbrella of a single experiment, thus creating a more realistic picture of events transpiring in nature) or rainfall. It is by now known that climate warming is likely to generate more extreme weather events at local scales, rather than simply resulting in gradual changes that are measured at large spatial scales across the biosphere. When confronted with these conditions, insects have to adapt or to move to new habitats where conditions are more suitable. In time, the latter will result in range shifts, a phenomenon by now well described in many studies (see next section, below). Those insects that 'remain behind'will exhibit physio‐ logical responses to warming and phenomena associated with its such as an increase in heat

**5. Effects of climate change on insect growth and development**

nation of plant and host-or prey-related odours.

928 Biodiversity in Ecosystems - Linking Structure and Function

cascade up to affect higher trophic level consumers.

waves and changes in precipitation.

As with other organisms, various insect taxa are adapted to climate windows and have welldefined ranges which coincide with both biotic and abiotic conditions [90]. Climate warming is already known to be driving demographic and geographic responses in insects [6,8,14,91,92]. Range shifts in herbivores and their natural enemies depend on a number of ecological factors that go beyond warming and which are often complex. For specialist herbivores that feed on only one or a few related plant species, a major impediment to movement is the availability of nutritionally suitable plants in their new habitats [6]. Generalist herbivores, on the other hand, may benefit if they are able to feed on a range of unrelated plants in both their native and invasive ranges. For many herbivores this is not a problem if their food plants also track the warming climate or if related plants with similar allelochemistries are also found in the new ranges of the insects. The oak processionary caterpillar (*Thaumetopoea processionea*) has expanded its range dramatically to the north within Europe over the past 30 years, coinciding with the recent warming episode [93]. Suitable oak trees on which the larvae can feed and develop, are found over much of Europe, helping to facilitate its spread. Its close relative, the pine processionary caterpillar (*Thaumetopoea pityocampa*), has also spread northwards as a result of recent warming and is projected to arrive in the Benelux region in the near future [94, 95]. Both species are considered as major health hazards owing to the production of numerous urticating hairs in mid-and late larval instars that contain soluble proteins and which are highly irritating to the skin and mucous membranes of humans [96]. The oak processionary caterpillar is actually more abundant now in many parts of its invasive than in its native range, perhaps because its specialist natural enemies have not effectively tracked its northwards expansion [93]. This is a worrying pattern that, if repeated in many trophic interactions, could facilitate pest outbreaks with large attendant economic costs.

Many other insects are known to be expanding into new habitats as a result of climate warming [7,97,98]. As they do so, they interact with native plants and their associated arthropod communities. The broader ecological outcome of these interactions is open to considerable debate. There is the possibility of community reorganization or reassembly as some species compete with and potentially displace others (see more detailed discussion of this below). Studies examining a suite of ecophysiological processes that underpin the ways in which these interactions work are urgently required. For instance, different species within food webs may each respond differently to changes in abiotic conditions such as temperature and moisture. The diamondback moth (*Plutella xylostella*) is native to Africa and the Mediterranean region, but has been introduced over many parts of the world where it is a serious pest of cabbages and related crops [99]. The moth only began to successful‐ ly overwinter in central Europe in the past 20 years, allowing to have two generations per year and to build up numerically faster by mid-summer [99]. It is attacked by several larval endoparasitoids, each of which exhibits differential responses to temperature. Recent warming in central Europe appears to favor thermophilic parasitoids like *Costesia vestalis* and *Dolichogenidea sicaria* over cool-favoring species like *Diadegma semiclausum* [99, J. Harvey, personal observations]. In addition to range expansions, many native insects will benefit from warming as a result of longer growing seasons and more favorable conditions for populations to grow. There is already evidence that some species are experiencing outbreaks as a result of warming as well as range expansions [100].

## **7. Effects of warming on the seasonal phenology of multitrophic interactions**

Jeffs and Lewis [17] examined the potential effects of climate warming on host-parasitoid interactions and developed three primary ways in which parasitoids might respond to warming: (1) by shifting distributions polewards or to higher elevations; (2) altering their phenology; (3) adjusting to persist in their current ranges through phenotypic plasticity or evolutionary adaptations. However, this ignores the potentially negative effects of a failure to respond to warming, or else the consequences of local changes on the survival and persistence of parasitoids. For example, warming is occurring so rapidly in many places that many species or populations may not be able to adapt in sufficient time. The authors fail to discuss the physiological costs of warming and how this might affect the acquisition and metabolic allocation of resources by the larvae (from the host) and the adult (from both host and/or nonhost sources). If development of immature stages is negatively affected, this might have profound effects on adult fitness and thus lead to declines and possible extinction. Further‐ more, phonological shifts depend on the ability of the parasitoid not only to track the host but the host's foodplant(s). Changes in important abiotic parameters may unravel trophic inter‐ actions if the species in these links respond differently to warming in terms of their life cycles. This has already been demonstrated in oak-winter moth interactions and the effects of this are negatively affecting the reproductive success of both migratory and resident insectivorous birds. There is a possibility that parasitoids of winter moths are also being negatively affected by warming. Some plants that are vital for the development and survival of specialist herbi‐ vores may also shift their seasonal growth patterns. For instance, interactions involving the large cabbage white butterfly, *Pieris brassicae*, its natural food plants and a specialist gregarious endoparasitoid, *Cotesia glomerata*, are complex in the context of life-history interactions involving the various parties. The herbivore and parasitoid each have 2-3 generations per year, each of which must seek out new food plant species in which to exploit. This is because most of its suitable food plants – brassicaceous species – are short-lived annuals or biennials, whereby different species grow at different times of the growing season [101]. Some species, such as *Brassica nigra*, grow early in the spring, whereas others, including *Sinapis arvensis* and *B. nigra* grow in late spring and summer respectively. The consequences of warming on the phenology of this trophic interaction may critically hinge on how the plants and insects each respond to increasing temperatures, and how this in turn affects the availability and suitability of the resources which they exploit as food. For the herbivore, of course, this means the availability of nutritious shoots on which the caterpillars feed, and for the parasitoid young caterpillars in which the female wasps oviposit clutches of up to 50 eggs. Warming will certainly increase the number of generations the insects have, and there already indications, based on populations in the Mediterranean region that up to 4 generations are indeed possible in more central and northern parts of Europe. However, if plant growth is temporally advanced (something that occurred in 2014) then later generations of the insects may emerge into habitats with little plant food available. There are, however, many possible scenarios, whereby the insects may experience neutral or negative effects of warming on their survival and fitness. Furthermore, this example is hardly likely to be an isolated one; indeed, many trophic interactions involving specialized consumers are under the same constraints.

## **8. Climate warming in the context of larger ecological scales**

populations to grow. There is already evidence that some species are experiencing outbreaks

Jeffs and Lewis [17] examined the potential effects of climate warming on host-parasitoid interactions and developed three primary ways in which parasitoids might respond to warming: (1) by shifting distributions polewards or to higher elevations; (2) altering their phenology; (3) adjusting to persist in their current ranges through phenotypic plasticity or evolutionary adaptations. However, this ignores the potentially negative effects of a failure to respond to warming, or else the consequences of local changes on the survival and persistence of parasitoids. For example, warming is occurring so rapidly in many places that many species or populations may not be able to adapt in sufficient time. The authors fail to discuss the physiological costs of warming and how this might affect the acquisition and metabolic allocation of resources by the larvae (from the host) and the adult (from both host and/or nonhost sources). If development of immature stages is negatively affected, this might have profound effects on adult fitness and thus lead to declines and possible extinction. Further‐ more, phonological shifts depend on the ability of the parasitoid not only to track the host but the host's foodplant(s). Changes in important abiotic parameters may unravel trophic inter‐ actions if the species in these links respond differently to warming in terms of their life cycles. This has already been demonstrated in oak-winter moth interactions and the effects of this are negatively affecting the reproductive success of both migratory and resident insectivorous birds. There is a possibility that parasitoids of winter moths are also being negatively affected by warming. Some plants that are vital for the development and survival of specialist herbi‐ vores may also shift their seasonal growth patterns. For instance, interactions involving the large cabbage white butterfly, *Pieris brassicae*, its natural food plants and a specialist gregarious endoparasitoid, *Cotesia glomerata*, are complex in the context of life-history interactions involving the various parties. The herbivore and parasitoid each have 2-3 generations per year, each of which must seek out new food plant species in which to exploit. This is because most of its suitable food plants – brassicaceous species – are short-lived annuals or biennials, whereby different species grow at different times of the growing season [101]. Some species, such as *Brassica nigra*, grow early in the spring, whereas others, including *Sinapis arvensis* and *B. nigra* grow in late spring and summer respectively. The consequences of warming on the phenology of this trophic interaction may critically hinge on how the plants and insects each respond to increasing temperatures, and how this in turn affects the availability and suitability of the resources which they exploit as food. For the herbivore, of course, this means the availability of nutritious shoots on which the caterpillars feed, and for the parasitoid young caterpillars in which the female wasps oviposit clutches of up to 50 eggs. Warming will certainly increase the number of generations the insects have, and there already indications, based on populations in the Mediterranean region that up to 4 generations are indeed possible in more central and northern parts of Europe. However, if plant growth is temporally advanced

**7. Effects of warming on the seasonal phenology of multitrophic**

as a result of warming as well as range expansions [100].

1094 Biodiversity in Ecosystems - Linking Structure and Function

**interactions**

Thus far, the relationship between temperature and insect behavior and development have been largely confined to pairwise interactions involving a plant and a herbivorous insect, or even the insect alone when reared on an artificial medium. Slightly more complexity has been achieved by incorporating a predator or a parasitoid into these studies, but the vast majority of them have been focused on optimal rearing conditions for biological control rather than on anthropogenic stresses such as climate change in natural and managed ecosystems. Indeed, in other fields where anthropogenic stresses are involved, such as invasive species or habitats loss, little attention has often been paid to insects in a multitrophic framework [10,13,102]. Climate change certainly represents a serious challenge to insects across vastly different scales of space and time because it will have cascading effects on a wide range of ecological charac‐ teristics and processes in habitats. An important challenge is to scale up the results of smallscale studies to see how these play out in communities, ecosystems and biomes. In this context we need to understand how biodiversity over large scales regulates ecosystem-level processes and how warming, by weakening processes and interactions at smaller scales, will affect this regulation.

The traditional approach in examining the relationship between biodiversity plays in ecosystem functioning has been based species interactions and the consequences of such interactions for community structure and function. These interactions can be classified as direct, involving pairwise interactions between species (e.g. predation/parasitism) or indirect, involving mediation by a third party [103]. Therefore, studying how biodiversity influences ecosystem functioning in multitrophic systems (involving mediation by a third party) is important for several reasons: (1) multiple trophic levels represent the core of ecosystems [104]; (2) as multitrophic diversity increases, average ecosystem properties could increase, decrease, stay the same or follow more complex non-linear patterns [105]. Consequently, as the number of species change within a community, the occurrence and significance of (in)direct interactions will also change. These, in turn, may be modified by abiotic factors, which may generate cascading reactions that generate large ecological changes with important ecological consequences [103,106].

Three of the well-documented global changes mediated by human activities are: increasing concentrations of carbon dioxide in the atmosphere; alterations in the biogeochemistry of the global nitrogen cycle; and ongoing land use/land cover change [106]. Human activity is now considered as the prime driver of global environmental change [106,107,108,109]. However, our ability to generate linkages at spatial (landscape) scales relevant to the human enterprise is limited at present [110]. Most importantly, the consequences of biodiversity loss on ecosys‐ tem services (e.g. primary and secondary production, nutrient cycling, pest control, pollina‐ tion, etc.) is poorly understood, as is our knowledge of the effects of warming on ecosystem processes [105]. Therefore, understanding an array of mechanisms that drive the biodiversity– ecosystem functioning relationship is thus difficult to evaluate in multitrophic systems [105], because of the unexpected consequences of warming on species interactions and demographics e.g. when the biology of one species is influenced by the biology of another species [111]; under simultaneously changing landscape characteristics such as habitat availability and landscape structure that affect biodiversity [112,113]; through the loss and/or fragmentation of habitats that drive changes in species abundance with both winners and losers [114].

Multitrophic interactions involve microbes, pathogens, plants, animals and other functional groups that are found in different positions of the food chain and provide vital functions to communities and ecosystems [115]. Here, we focus on organisms inextricably linked to plantinsect interactions. Microorganisms from diverse environments have played an important role in ecosystem sustainability. Since the spatial and temporal stresses of the microbial system may be quite different from those of plants and animals [116], many studies of the ecological responses to global changes have suggested that belowground processes, often mediated by soil microorganisms, are central to the response of ecological systems to global change [117]. Below-ground microbial mediated processes can both immobilize and release nutrients that limit primary production and can influence the long-term response of ecosystems to global change [118]. Investigations into microbial parameters involved in soil quality are increasing [118,119,120] and it has been shown that human activity can directly or indirectly affect the functioning and diversity of the soil community [116,121] and these effects are transferred aboveground where they effect the structure and function of plant-insect communities [19,122,123]. Bardgett [124] described some potential outcomes for soil microbes and carbon exchange that include: (i) increases in soil carbon loss by respiration and in drainage waters as dissolved organic carbon due to stimulation of microbial abundance and activity, and enhanced mineralization of recent and old soil organic carbon [125,126]; (ii) stimulation of microbial biomass and immobilization of soil N, thereby limiting N availability to plants, creating a negative feedback that constrains future increases in plant growth and carbon transfer to soil [127]; (iii) increased plant-microbial competition for N, leading to reduced soil N availability and microbial activity and suppression of microbial decomposition and ultimately increased ecosystem carbon accumulation [117]; (iv) increased growth of mycor‐ rhizal fungi [128,129], which receives carbon in the form of photosynthate directly from the host plant and retains this carbon, controlling its release to the soil microbial community [130]; and changes in root exudation that are known to play a potentially important role through the promotion of methanogenesis and hence carbon loss from soil as methane [131], although the mechanisms involved in this process are poorly understood.

The response of the soil free-living bacterial group (β subclass of the *Proteobacteria*) on simulated multifactorial global change was investigated in grassland vegetation dominated by annual grasses and forbs growing in sandstone-derived soil [118]. The results demonstrate that shifts in community composition were associated with increases in nitrification, but changes in abundance were not, confirming that microbial communities can be consistently altered by global changes and that these changes can have implications for communities and ecosystems.

is limited at present [110]. Most importantly, the consequences of biodiversity loss on ecosys‐ tem services (e.g. primary and secondary production, nutrient cycling, pest control, pollina‐ tion, etc.) is poorly understood, as is our knowledge of the effects of warming on ecosystem processes [105]. Therefore, understanding an array of mechanisms that drive the biodiversity– ecosystem functioning relationship is thus difficult to evaluate in multitrophic systems [105], because of the unexpected consequences of warming on species interactions and demographics e.g. when the biology of one species is influenced by the biology of another species [111]; under simultaneously changing landscape characteristics such as habitat availability and landscape structure that affect biodiversity [112,113]; through the loss and/or fragmentation of habitats

Multitrophic interactions involve microbes, pathogens, plants, animals and other functional groups that are found in different positions of the food chain and provide vital functions to communities and ecosystems [115]. Here, we focus on organisms inextricably linked to plantinsect interactions. Microorganisms from diverse environments have played an important role in ecosystem sustainability. Since the spatial and temporal stresses of the microbial system may be quite different from those of plants and animals [116], many studies of the ecological responses to global changes have suggested that belowground processes, often mediated by soil microorganisms, are central to the response of ecological systems to global change [117]. Below-ground microbial mediated processes can both immobilize and release nutrients that limit primary production and can influence the long-term response of ecosystems to global change [118]. Investigations into microbial parameters involved in soil quality are increasing [118,119,120] and it has been shown that human activity can directly or indirectly affect the functioning and diversity of the soil community [116,121] and these effects are transferred aboveground where they effect the structure and function of plant-insect communities [19,122,123]. Bardgett [124] described some potential outcomes for soil microbes and carbon exchange that include: (i) increases in soil carbon loss by respiration and in drainage waters as dissolved organic carbon due to stimulation of microbial abundance and activity, and enhanced mineralization of recent and old soil organic carbon [125,126]; (ii) stimulation of microbial biomass and immobilization of soil N, thereby limiting N availability to plants, creating a negative feedback that constrains future increases in plant growth and carbon transfer to soil [127]; (iii) increased plant-microbial competition for N, leading to reduced soil N availability and microbial activity and suppression of microbial decomposition and ultimately increased ecosystem carbon accumulation [117]; (iv) increased growth of mycor‐ rhizal fungi [128,129], which receives carbon in the form of photosynthate directly from the host plant and retains this carbon, controlling its release to the soil microbial community [130]; and changes in root exudation that are known to play a potentially important role through the promotion of methanogenesis and hence carbon loss from soil as methane [131], although the

that drive changes in species abundance with both winners and losers [114].

1296 Biodiversity in Ecosystems - Linking Structure and Function

mechanisms involved in this process are poorly understood.

The response of the soil free-living bacterial group (β subclass of the *Proteobacteria*) on simulated multifactorial global change was investigated in grassland vegetation dominated by annual grasses and forbs growing in sandstone-derived soil [118]. The results demonstrate that shifts in community composition were associated with increases in nitrification, but Alternatively, atmospheric CO2 on microbial decomposition in peat was found to be greater than when these factors operated alone, creating a stronger positive feedback on carbon loss from soil as DOC (dissolved organic carbon) and respiration [126].

Soil microbes and their activities are inexorably linked to above and below ground commun‐ ities and therefore it is necessary to understand the effects of climate change on carbon budgets [124]. However, it is important to be aware also of adverse effect of soil microbes on multitro‐ phic interactions, mostly in the case of pathogens. Soil borne pathogens can significantly alter the spatial distribution and reduce the yield and quality of plants [132,133]. Climate change generally has a beneficial effect on pathogen suppression and stress tolerance [115]. Bacteria, fungi and nematodes perform specific functions that help suppress the infection and coloni‐ zation of plant roots by pathogens in the rhizospehere [115].

Plants play a major role in controlling and mediating directly or indirectly all soil multitrophic interactions [115] and contribute to ecosystems on by providing food, shelter or "ecological islands" for microbes, fungi, insects, and other organisms. However, dead plants also continue to supply the environment with nutrients. Thus, qualitative and quantitative changes in plant physiology, chemistry, tissue composition and signaling pathways under rising CO2 and temperature regimes may influence the entire cascade of above and below-ground multitro‐ phic interactions [134,135,136,137]. Moreover, the geographical distribution of plants is also affected by changing temperature and precipitation that alter the structure of plant commun‐ ities at larger scales and which in turn generate new (positive or negative) trophic interactions [138]. The principle effect of changes in the spatial distribution of plants is also influenced by the presence or absence of mutualists (pollinator) and antagonist (pathogen and herbivore) in a changing environment [139] as well as intra-interspecific competitors (taller plants replace lower congeneric species) [140]. On the other hand, plant phenology is temperature-sensitive with strong connections with photoperiod [141]. Accordingly, there is critical springtime photoperiod, which is an important pre-requisite before floral development can begin [141,142]. Net photosynthesis (primary production) typically peaks within the range of normal temperatures [143,144], however increasing temperatures can shift this peak and extending the growing season and consequently accelerate plant growth [145]. This situation can lead to key differences with respect to interactions with insect mutualists or antagonists [16] and subsequently changes in the population dynamics can alter evolutionary trajectories [146], creating "evolutionary noise" with unpredictable consequences for the strength and persis‐ tence of trophic interactions. While evidence of adaptive responses via variation among genotypes in response to increasing temperature has been described in a marine parasite *Maritrema novaezealandensis* [147], the extinction of local populations of the angiosperm *Fagus sylvatica* still occurred at the southern range margin despite strong signals of genetic adaptation in this species to climate warming [148,149]. This variation in responses at the species level makes it virtually impossible to predict the effects of climate change on trophic interactions, although available evidence suggests that there will be many more losers than winners [6] and as a result we can expect multitrophic (ecosystem) interactions to be simplified (Fig. 2) and ecosystems therefore to become less stable and resilient [150].

**Figure 2.** Variation in responses at the species level makes it virtually impossible to predict the effects of climate change on trophic interactions, although available evidence suggests that there will be many more losers than winners and as a result we can expect ecosystem interactions to be simplified and ecosystems therefore to become less stable and resilient.

### **9. Summary and future directions**

Climate change clearly represents a major challenge to biodiversity at all levels of organization. This includes physiological and behaviorial responses of individual species through trophic interactions and beyond to the scale of communities and ecosystems. Given that they are ectotherms, insects will respond to warming in a number of ways. First, they may seek out micro-climates with more optimal temperatures for biological activity; second, they may shift their diel patterns of activity to coincide with more optimal temperature regimes; third they will adjust their broader distributions either moving to higher elevations or polewards. However, these responses also critically depend upon interactions with other species, includ‐ ing both resources (e.g, plant for herbivores and herbivores for predators) as well as indirect interactions with species two levels or more apart on the food chain.

Thus far, few studies have explored climate change on trophic interactions integrating three or more levels of a food chain. Instead, the main focus thus far has been on descriptive studies or else on two-trophic level interactions, with little attention paid to members higher up the food chain (e.g. parasitoids and hyperparasitoids), especially in a multitrophic framework. Given the paucity of studies in this area, it is virtually impossible to predict larger scale consequences of warming for communities and ecosystems and the services that emerge from them. It is therefore important to scale up the biological effects of warming on individual species of producers (plants) and consumers (herbivores and natural enemies) to trophic chains and from there to communities, in order to better understand the potential consequences on the stability and resilience of ecosystems and the biomes in which they are embedded. This is a daunting challenge, but one that must be tackled if we are to develop predictive power of climate change effects on natural and managed ecosystems. Another important area is to combine warming in an above-and below ground approach integrating trophic interactions occurring amongst insects and other invertebrates in both compartments, an area that is now being increasingly considered in community-based research [151].

## **Author details**

as a result we can expect multitrophic (ecosystem) interactions to be simplified (Fig. 2) and

**Figure 2.** Variation in responses at the species level makes it virtually impossible to predict the effects of climate change on trophic interactions, although available evidence suggests that there will be many more losers than winners and as a result we can expect ecosystem interactions to be simplified and ecosystems therefore to become less stable

Climate change clearly represents a major challenge to biodiversity at all levels of organization. This includes physiological and behaviorial responses of individual species through trophic interactions and beyond to the scale of communities and ecosystems. Given that they are ectotherms, insects will respond to warming in a number of ways. First, they may seek out micro-climates with more optimal temperatures for biological activity; second, they may shift their diel patterns of activity to coincide with more optimal temperature regimes; third they will adjust their broader distributions either moving to higher elevations or polewards. However, these responses also critically depend upon interactions with other species, includ‐ ing both resources (e.g, plant for herbivores and herbivores for predators) as well as indirect

Thus far, few studies have explored climate change on trophic interactions integrating three or more levels of a food chain. Instead, the main focus thus far has been on descriptive studies or else on two-trophic level interactions, with little attention paid to members higher up the food chain (e.g. parasitoids and hyperparasitoids), especially in a multitrophic framework. Given the paucity of studies in this area, it is virtually impossible to predict larger scale consequences of warming for communities and ecosystems and the services that emerge from them. It is therefore important to scale up the biological effects of warming on individual

interactions with species two levels or more apart on the food chain.

ecosystems therefore to become less stable and resilient [150].

1498 Biodiversity in Ecosystems - Linking Structure and Function

and resilient.

**9. Summary and future directions**

Jeffrey A. Harvey1,2\* and Miriama Malcicka2

\*Address all correspondence to: j.harvey@nioo.knaw.nl

1 Netherlands Institute of Ecology, Department of Terrestrial Ecology, Wageningen, The Netherlands

2 VU University Amsterdam, Department of Ecological Sciences, Animal Ecology, Amsterdam, The Netherlands

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**Provisional chapter Chapter 5**
