Animal pollination is a fundamental process in terrestrial ecosystems and is essential for maintenance of wild plant communities and agricultural systems (Potts et al. 2010). Faunal pollination is a key ecosystem service in agricultural productivity. In contrast with the other regulation FES, the contribution of animal pollination has been widely quantified.

According to Klein et al. (2007), 35% of crops depend on pollinators globally, while Kremen et al. (2002) estimated 66% for the 1,500 crop species of the world amounting to between 15 and 30% of food production. Williams (1996) estimated for European crops that over 80% of the 264 species assessed require animal pollination. Roubik (1995) estimated that productivity of approximately 70% of tropical crops is improved by animal pollinators. Regarding wild plant species, 80% of flowering plants are directly dependent on insect pollination for fruit and seed set globally (Klein et al. 2007, Ollerton et al. 2011).

Given the morphological diversity of plants, the degree of self-compatibility and the diversity of reproductive organs in the flowers of crops, a great diversity of vectors is required for efficient animal pollination (Williams 2002, Blüthgen and Klein 2011). Insects are the most important animal pollinators by virtue of their abundance and foraging behaviour (Williams 2002). Thousands of species of bees, flies, wasps, beetles, butterflies and moths contribute to pollination of many crops, such as gourds, oilseeds, berries and tobacco, amongst many others (Roubik 1995, Williams 2002, Blanche and Cunningham 2005), as well as a countless number of wild plant species. Bees are probably the most recognised pollinators (>12,000 species; e.g. Kremen et al. 2002, Larsen et al. 2005, Morandin and Winston 2005, Potts et al. 2010, Kerr et al. 2015).

Biological control is the natural process responsible for the regulation of species’ population growth through ecological interactions –mutualism, parasitism and predation. This has been highlighted as a relevant regulation FES given the key role in restraining the spread of crop pests and diseases (Wilby and Thomas 2002, Fiedler et al. 2008, Karp and Daily 2014).

Oerke (2006) made an estimation of potential and actual losses due to pests for wheat, rice, maize, potatoes, soybeans and cotton, between 2001 and 2003, worldwide. Arthropod pests destroy 8–15% of these crops and without natural biological control and pesticides, this figure could reach 9–37%. According to the estimation done by Losey and Vaughan (2006), crop damage due to the absence of arthropod native predators might cost approximately US $4.5 billion more than the actual cost given pest control services.

Predation is one of the best-known mechanisms of biological control for agricultural pests and birds and bats have been identified as the main contributors, by their predation of species responsible for crop damage (Mols and Visser 2007, Maas et al. 2013, Karp and Daily 2014, Railsback and Johnson 2014). Increasing knowledge of the relevance of predators for pest control has increased the concern to conserve the conditions required to maintain these predators (e.g. Williams-Guillén and Perfecto 2010, Railsback and Johnson 2014).

Parasitoidism is considered another important mechanism of agricultural pest control (Drechsler and Settele 2001, Letourneau et al. 2015). The main providers identified are flies and parasitoid wasps, which lay eggs on or in the body of a host, in this case pest insects, eventually killing the hosts and diminishing the spreading of the pest.

Mutualism has been identified as another mechanism that can contribute to pest control. Perfecto and Vandermeer (2006) provided evidence that the mutualistic relationship between the Aztec ant and a coccid has a positive effect on coffee plants by reducing the numbers of the coffee borer beetle, coffee’s main pest. This exemplifies the complexity of biological control mechanisms and how an imbalance in ecological condition can negatively impact this FES.

Disease control is also recognised as a relevant FES (Ostfeld and LoGiudice 2003, Foley et al. 2005, McCauley et al. 2015). Wild and domestic animals are vectors for a wide range of infectious diseases that are potentially transmitted to humans (see Molyneux et al. 2008, Civitello et al. 2015). Healthy populations of animals (i.e. populations with the minimum number of sexually mature individuals required to secure their viability) and high diversity provide less risk of human infection, since the probabilities of vectors (e.g. flies and ticks) targeting humans as hosts decreases with higher availability of other host species (Keesing et al. 2006, Civitello et al. 2015). Disease control is a FES directly related to human health and well-being.

Animals are also relevant actors in seed dispersal. They drive plant gene flow, population dynamics and spatial structure in undisturbed habitats and contribute to regeneration of deforested habitats, by moving seeds from one site to another (Russo et al. 2006, García and Martínez 2012). Animals are considered long-distance vectors; they contribute to seed dispersal mainly by defecation and epizoochory (seeds adhere to the outside of animal bodies). These include ants, frugivorous terrestrial, arboreal and flying mammals and frugivorous and/or caching birds (Greene and Calogeropoulos 2002). Animal seed dispersal is an essential mechanism in the maintenance of temperate and tropical ecosystems (García and Martínez 2012) and are particularly important for large-seeded plants (Greene and Calogeropoulos 2002, McConkey and Drake 2006, Wenny et al. 2016). Approximately one-half of seed plant species are dispersed by animals (Wenny et al. 2016, Egerer et al. 2018).

The ecological value of faunal dispersal is well known (Russo et al. 2006, Wenny et al. 2016). However, in comparison with animal pollination and pest control, the quantitative assessment of the seed dispersal service by fauna is scarce. Seed dispersal benefits are spatially and temporally distant from the mother plant, making them difficult to measure, especially for tree species and species not used directly by humans (Wenny et al. 2016) and further quantitative assessment is required for this FES.

The economic value of animals for seed dispersal is even less well known than their ecological value (Wenny et al. 2016). Some studies have indirectly estimated the value of animal seed dispersal through the economic valuation of associated food and fibre consumed by humans (e.g. Fujita and Tuttle 1991, Paoli et al. 2001). However, studies on direct valuation are scarce. A direct economic valuation is made by Hougner et al. (2006), who value seed dispersal carried out by the Eurasian Jay in an oak forest, through the estimation of the cost of replacing birds by human force.

Some of the studies where the role of animals in seed dispersal has been assessed are in tropical ecosystems. McConkey and Drake (2006) highlighted the relevance of flying foxes to sustain Pacific island forests, since these are the only existing animals capable of dispersing large seeds over long distances in such isolated habitats. Egerer et al. (2018) showed that bird dispersal provides a benefit to wild chilli plants in Guam through increased seedling emergence of gut-passed seeds in comparison to depulped seeds and whole fruits.

Effects of invasive species on native species are well documented (e.g. D’Antonio et al. 2004, Alpert 2006) and, more recently, their effects on ecosystem services has also drawn attention (Pejchar and Mooney 2009, Pyšek and Richardson 2010, Peh et al. 2015, Walsh et al. 2016). According to Pejchar and Mooney (2009), the impact of alien species is usually well quantified for provisioning services (food, fibre and fuel). However, impacts on regulation FES are rarely calculated, but are likely to be substantial.

Amongst the reported effects of invasive species on animal pollination services are: the disruption of mutualism between native bees and plants by invasive bees, the range expansion in pollinator-limited invasive plants and consequent distraction of pollinators from native plant species (Stokes et al. 2006, Traveset and Richardson 2006). According to the review made by Morales et al. (2017), the impacts of invasive pollinators on pollination are predominantly negative for native plants, mixed for crops and positive for invasive plants. Although invasive pollinators can be beneficial for some native plants in highly disturbed habitats and some crops in intensively modified agroecosystems (e.g. Ricketts et al. 2004), they cannot replace the role of a diverse pollinator assemblage for wild plant reproduction and crop yield.

Invasive species like weeds, insects and plant pathogens (mainly fungi) can become pests and have major impacts on crops. For instance, a well-documented case is the Mediterranean fruit fly, native from West Africa, but now found worldwide, which causes damage to over 250 types of crops. The cost estimated for California reaches US $1 billion (Mooney 2005). Similarly, Pimentel et al. (2005) made a detailed review of the environmental and economic costs associated with alien species in the United States. Related to crops, pasture and forest losses, they identify 500 weed species, feral pigs, European starlings, over 900 insect species and 20 plant pathogen species, as the main agents. The cost of losses, damages and control techniques reaches an annual value of approximately US $50,000 million.

Animal seed dispersal can be a disservice when this involves the spread of invasive plants. Just like the service, the knowledge on how animals contribute to the success of invasive plants is limited (Gosper et al. 2005). However, several mechanisms have been identified: the invasive plant species rely on common native disperser species with generalist diets; the invasive plant is reunited with the disperser species of its native range — like the case of Rubus spp. and blackbirds (Turdus merula) in Australia; and a new association between plant and animal can occur — like the case of the accidental spread of seeds of wind dispersed pines, Pinus spp., by seed predating cockatoos, Calyptorhynchus spp., in Australia (Richardson et al. 2000). Additionally, the dispersal of native plants is affected by the competition of dispersal service from invasive plant species (Gosper et al. 2005).

Equally relevant is the effect of invasive species on disease control: invasive plants and animals can act as novel hosts for diseases. Pyšek and Richardson (2010) provided detailed examples of how several invasive species affect human health, acting as vectors (e.g. rodents and bats as vectors of rabies, leptospirosis and hepatitis) or acting directly (e.g. snake bites).

Overpopulation of disease organisms or disease vector organisms and/or the absence of defence organisms can increase the risk of spread for human disease. Many cases of disease outbreaks in human history have been related with invasive pathogens, due to the continual expansion and interchange of human population worldwide (Dobson and Carper 1996, Pejchar and Mooney 2009). For instance, the introduction of smallpox, measles and typhus with European arrivals to the New World increased mortality of the native human population at unprecedented rates (Dobson and Carper 1996). More recently, the increase of mosquito-borne diseases, like yellow fever and dengue, has been attributed to invasive mosquitoes in America and Asia (Pejchar and Mooney 2009). The negative effect can also be indirect, for example, the invasion of the American plant Lantana camara in East Africa. L. camara is now the habitat of the tsetse fly, vector of sleeping sickness.

Native species may also represent a risk for human health if the natural control of population growth is altered or if human contact with vectors increases. For instance, Ostfeld and LoGiudice (2003) evidenced how the risk of human exposure to Lyme disease increases due to the decrease in diversity of other hosts for ticks (Lyme disease vectors). Equally, McCauley et al. (2015) showed how changes in rodent and flea community composition due to LUCC, increase the abundance of Mastomys natalensis, transmitter of plague, in agricultural habitats in Tanzania.

Since the beginning of agriculture, humans have faced crop pests (Oerke 2006), which have had major impacts in human history. Pests, such as rusts on wheat, ergot on rye potato blight, gypsy moth and the boll weevil, have had deep social and economic consequences (Horsfall 1983). Currently 10–16% of global crop production is lost due to pests (Bebber et al. 2013).

Amongst the known causes of occurrence of crop pests is the imbalance of natural biological control, produced by a change in the abundance of natural enemy populations. For instance, a decrease in predator populations allows the increase of prey population (e.g. Drechsler and Settele 2001, Wilby and Thomas 2002, Karp and Daily 2014). Other causes are the absence of indigenous populations which facilitates the success of invasive species with similar ecological requirements (Pejchar and Mooney 2009) and the concentration of food resources, especially in perennial monocultures (Risch 1981, Altieri 2018). Although crop pests have been present since the appearance of agriculture, modern agricultural practices, like agricultural intensification (e.g. Wilby and Thomas 2002), manipulation of soil fertility and irrigation (e.g. Fuller et al. 2012) and use of chemical pesticides (Rosenzweig et al. 2001) have exacerbated these causes (Tilman 1999).

Crop raiding is the term used to describe the action of wild animals foraging or trampling crops (Hill 2016). In this context, wildlife is considered a pest. However, this is not produced by an imbalance in wildlife populations, but by the increasing overlap of human and wildlife niches, due to continuous human population growth and the anthropogenic transformation of habitat (Hill 1997, Campbell et al. 2000). The most commonly identified actors are medium and large sized mammals (e.g. monkeys, wild pigs, hippopotamus, elephants; Naughton-Treves 1998, Engeman et al. 2010, Ango et al. 2016), but some studies also refer to small mammals and birds (e.g. Naughton-Treves and Treves 2005). Amongst the identified factors influencing crop raiding are the distance from cropland to natural habitat patches, the crop type and hunting practices (Naughton-Treves 1998). Drought, leading to paucity of production in (non-irrigated) natural lands, can also act as a push factor alongside the pull factor of higher productivity in irrigated or improved agricultural areas (Mulligan 2018).

Literature on this subject is extensive and mostly consists of case studies. The approaches to quantify losses vary considerably and are not comparable from site to site (McGuinness and Taylor 2014). The impacts have been assessed in human settlements adjacent to natural protected areas, where the raiding occurs frequently (e.g. Sekhar 2002, Linkie et al. 2007, Hedges and Gunaryadi 2010). However, there are also studies that address this phenomenon outside of protected areas (e.g. Ango et al. 2016, Chaves and Bicca-Marques 2017).

The extent of damage varies widely depending on where the raiding occurs and the type of crops and wildlife species involved. For instance, the socioeconomic impact might be higher in developing countries in non-protected areas with farmers losing their livelihood and rarely being compensated for the losses, thereby creating antagonism towards wildlife (Linkie et al. 2007). In contrast, in protected areas, prevention and compensation measures are more frequently enforced (Sekhar 2002, Davies et al. 2011).

The approaches to estimate monetary losses are variable, varying in unit of measurement and spatial scale. For example, Chakravarthy and Thyagaraj (2005) estimated a loss of US $8 per kilogram of dry capsules of cardamom caused by the Bonnet macaque (Macaca radiate), while Engeman et al. (2010) estimated that Rhesus macaque (Macaca mulatta) and Patas monkey (Erythrocebus patas), both invasive species, causes a nationwide economic impact of US $1.46 million per year in Puerto Rico.

Human-driven environmental changes strongly influence the occurrence of faunal disservices. Simultaneously, these environmental changes have an adverse effect on faunal services through the negative impact on the providers, mainly caused by the loss or transformation of habitat.

Landscape heterogeneity refers to the variety of different landscape conditions within a landscape (i.e. area that is spatially heterogeneous in at least one factor of interest, Turner and Gardner 2015) as with mixed habitats or land cover types. A closely related concept is landscape complexity, which can be defined as the level of difficulty observed in understanding the interactions of the landscape components (Papadimitriou 2010). The relationship between these concepts is controversial. Heterogeneity has been described as a function of complexity (e.g. Chen and Xu 2015), at the same time, heterogeneity has been considered an attribute of complexity (e.g. Papadimitriou 2010); furthermore, the terms have been used interchangeably (e.g. Miles et al. 2012).

The inconsistency in the use of terms makes the comparison and synthesis of studies difficult (Reyes Sandoval 2017). However, for practical purposes, we consider that loss of complexity/heterogeneity or landscape homogenisation/simplification refers to the same phenomenon: loss of components and/or loss of the interaction amongst components in a landscape.

The idea that the diversity of landscape components is a key determinant for biodiversity is widely accepted (Fahrig et al. 2011, Katayama et al. 2014). Increased landscape heterogeneity is generally associated with increased biodiversity, since high habitat and resource diversity allows high diversity of species, while the opposite, i.e. landscape homogeneity, is associated with low biodiversity (Parks and Mulligan 2010, Stein et al. 2014).

A consequence of LUCC due to agriculture is landscape homogeneity, as different land cover and habitat types are converted to more uniform agricultural land. Therefore, the proportion of agricultural land is the most commonly used indicator of homogenisation in studies where the relationship between biodiversity and landscape heterogeneity is assessed (e.g. Letourneau et al. 2015, Maas et al. 2016, Jonason et al. 2017). Other indicators include distance from original habitat (e.g. Blanche and Cunningham 2005, Ricketts et al. 2008) and diversity and management indices (Gardiner et al. 2009, Williams-Guillén and Perfecto 2010, Chaplin-Kramer et al. 2011).

Several studies support a positive relationship amongst landscape heterogeneity, species diversity and abundance of FES providers (Table 3). Although neutral or mixed relationships have also been evidenced (Jonsen and Fahrig 1997, Chaplin-Kramer et al. 2011), due mostly to species’ particular ecological traits and range sizes (Katayama et al. 2014), landscape heterogeneity has proven to be a relevant factor in ecosystem functioning and population dynamics. Sustainable landscape management is suggested as the most important means of maintaining healthy populations of FES providers (Ricketts et al. 2008, Maas et al. 2013, Letourneau et al. 2015). There is also evidence that homogenised landscapes favour the occurrence of disservices by reducing the diversity and abundance of beneficial arthropods, such as pollinators and parsitoid insects and vertebrate predators (e.g. Letourneau et al. 2015, Senapathi et al. 2015, Maas et al. 2016) and thus increasing the outbreaks of herbivore and diseases pests (e.g. Altieri 1999, McCauley et al. 2015).

Along with landscape homogenisation, agricultural intensification has led to original habitat loss and concurrently to habitat fragmentation. Habitat fragmentation refers to the reduction of continuous tracts of habitat to smaller, spatially distinct remnant patches (Wilson et al. 2016). Fragmentation alters habitat connectivity and quality, affecting biodiversity and ecosystem functioning negatively (Haddad et al. 2015). Equally, reduction of the original habitat of animal populations has increased the conflict between humans and wildlife and the risk of disease transmission (Campbell et al. 2000, Xavier et al. 2012).

The degradation of ecosystems by landscape homogenisation, habitat loss and fragmentation results in decreased carrying capacity to sustain all the organisms that inhabit these ecosystems, leading to continued population losses. The loss of populations precedes species extinction and, therefore, the reduction of biodiversity (Ceballos and Ehrlich 2002).

Several studies have suggested that the loss of genetically distinct populations globally is both absolutely and proportionally several times greater than the rate of extinction of species (Hughes et al. 1997, Ceballos and Ehrlich 2002, Gaston et al. 2003). Genetic variation amongst and within populations confers resilience to environmental change whereas the loss of individuals or populations increases the vulnerability of species, destabilises ecosystem functions and affects ES provision (Luck et al. 2003).

Habitat loss and fragmentation are the main causes of population decline (Fahrig 1997, He and Hubbell 2011, Wilson et al. 2016). Hughes et al. (1997) estimate the population diversity, defined as the number of populations on the planet, for 82 species (35 vertebrates, 23 plants, 19 arthropods, four molluscs and one platyhelminth) in the range 1.1–6.6 billion populations. By using the midrange estimation (3 billion populations), assuming a linear function between population and habitat loss and that two-thirds of all populations exist in tropical regions, they estimate that 16 million populations are lost annually across these 82 species alone.

Ceballos and Ehrlich (2002) made an indirect estimation of mammal population loss globally, by comparing present and historic ranges of 173 declining species, reaching a collective loss of 50% of range area. Regarding bird populations, Gaston et al. (2003) estimated a loss of approximately 22% of breeding bird individuals so that an average of 87 billion breeding bird individuals remain from approximately 112 billion estimated before 1700, which is considered the starting date for development of the current pattern of LUCC due to agriculture.

Global declines in pollinator populations are widely recognised (Biesmeijer et al. 2006, Gallai et al. 2009, Potts et al. 2010) and habitat loss is considered the main threat, particularly for habitat and plant specialists (Ricketts et al. 2008, Potts et al. 2010, Winfree et al. 2015). Equally, decline in predator and parasitoid populations due to habitat loss has been reported (Williams-Guillén and Perfecto 2010, Letourneau et al. 2015).

LUCC causes microclimatic changes in the remaining patches of ecosystem related to temperature, wind and humidity (Meyer and Turner 1992). There is evidence that deforestation can modify local rainfall and droughts pattern, changes in moisture and humidity can also negatively affect canopy, understorey and litter organisms and can increase fire frequency in tropical and arid ecosystems (Goldammer and Seibert 1990, Rao 2009), increasing the mortality of animal populations.

Along with climatic modification, physical changes diminish animal habitat suitability by reducing the quantity and quality of nesting, sheltering, and foraging sites (Frumhoff 1995). These changes can affect ecological interactions, survivorship, reproductive fitness and distribution of populations, particularly for highly specialised organisms (Dale 1994, Afrane et al. 2006, Rao 2009). Finally, the decrease in population sizes at the interface between two land cover types, known as the edge effect, is also enhanced by habitat fragmentation, caused by deforestation (Levin et al. 2009).

Environmental changes caused by LUCC may adversely affect biological control processes. Spatial and temporal distribution and proliferation of insects, weeds and pathogens is largely determined by climate, therefore microclimatic changes in temperature, light and water supply can drive overpopulation of pests (Rosenzweig et al. 2001). Pest proliferation has detrimental consequences for ecosystems (Chapin et al. 2000, Wilby and Thomas 2002, Foley et al. 2005). For example, Imperata cylindrica, an aggressive indigenous grass, which colonises forest lands of Asia that are cleared for slash-and-burn agriculture, forms a monoculture grassland with no vascular plant diversity and few mammalian species in comparison with the native forest (Chapin et al. 2000).

Crop pests produce major losses for crop yields, therefore, farmers have resorted to the use of pesticides as a means of control. In the last six decades, there has been a dramatic increase in the use of pesticides. Along with agricultural intensification, herbicides, insecticides and fungicides have produced highly negative effects on species abundance and diversity (Geiger et al. 2010, Isenring 2010) and also threaten water quality (Vymazal and Březinová 2015) and human health directly (see Budzinski and Couderchet 2018) . There is evidence of the adverse effect of chemical pest control on farmland and wildlife populations worldwide (e.g. amphibians and reptiles, Gibbons et al. 2000, farmland birds, Boatman et al. 2004, benefitial arthropods, Desneux et al. 2007). Direct adverse effects include higher mortality due to poisoning, reduced fecundity and detrimental changes in physiology and behaviour. Indirect effects include reduction of habitat, due to destruction of non-invasive vegetation, reduction of food resources for predators by indiscriminate elimination of arthropod populations and imbalance in ecological interactions.

Regarding animal pollination, the high diversity in morphology and reproductive strategies of plants requires a similar diversity of pollinators (Blüthgen and Klein 2011). Therefore, a decrease in pollinator diversity potentially causes a decline in wild plant and crop diversity (Biesmeijer et al. 2006, Potts et al. 2010), as well as reduced crop productivity. Blanche and Cunningham (2005) observed a highly significant reduction in fruit set due to pollinator exclusion in atemoya crops. The risk is greater for wild or crop species that rely on a narrow range of pollinator species. Although the threshold of diversity, required to maintain pollination stability, depends on the biology and variety of crops, landscape structure and regional pollinator community, the evidence suggests that stability is higher with a diverse and abundant pollinator community (Klein et al. 2007).

Equally, a detrimental effect on natural pest control in crops has been identified due to a reduction in natural enemy diversity (e.g. rice crops, Drechsler and Settele 2001, Wilby and Thomas 2002, cacao plantations, Maas et al. 2013, coffee plantations, Karp and Daily 2014). Straub et al. (2008) indicated that higher diversity of predators implies higher complementarity on functional roles: feeding on different pest species, at different life stages of the pest, using diverse strategies and differential partitioning of space and/or time (e.g. eating pest insects from different parts of the plant or during different seasons). This explanation could also be applied to parasitoid species.

Human disease control can be affected by reduction in species richness. A ‘dilution effect’ (sensuKeesing et al. 2006), where increased species diversity reduces disease risk for individual species, has been described for some diseases (e.g. tick-borne diseases, Norman et al. 1999, Ostfeld and LoGiudice 2003, viral pulmonary disease, Ruedas et al. 2004, mosquito-borne diseases, Ezenwa et al. 2006, Allan et al. 2009). This indicates richness loss can lead to more disease. Keesing et al. (2006) provides a detailed explanation of the mechanisms through which higher species richness decreases disease risk, including: reducing the rate of encounter between susceptible and infectious individuals, reducing the probability of transmission given an encounter, decreasing the density of susceptible individuals, increasing the recovery rate and increasing the death rate of infected individuals.

Seed dispersal is also affected by diversity loss. García and Martínez (2012) found a clear positive relationship between richness of frugivorous birds and all components of seed dispersal (i.e. seed richness and abundance and arrival and colonisation rates). Just like pollination and biological control, this suggests the existence of functional complementarity and/or facilitation amongst dispersers.

In general, even though initial species loss can be compensated by remaining species with similar functions, significant species loss will eventually reduce provisioning of FES. Therefore, to secure FES production, it is essential to conserve species richness.

Along with species richness, population size or abundance, are determining factors for FES provision. Since population losses are higher than diversity losses (Ceballos and Ehrlich 2002, Gaston et al. 2003), these can have major implications on the magnitude and quality of FES provision.

Losses in pollinator populations produce a negative impact in wild plant communities, affecting the integrity of natural vegetation (Williams 2002, Biesmeijer et al. 2006). Additionally, population declines reduce crop production (Kremen et al. 2002, Larsen et al. 2005, Klein et al. 2007), causing important economic losses (Losey and Vaughan 2006, Gallai et al. 2009) and jeopardising food sufficiency worldwide (Aizen et al. 2009).

Equally affected is the pest control service, where abundance of natural enemies, predators and parasitoid species, largely determines the abundance of species that can become pests (Drechsler and Settele 2001, Mols and Visser 2007, Railsback and Johnson 2014). Like pollinators, losses in natural enemy populations cause losses in natural and agricultural systems (Losey and Vaughan 2006, Oerke 2006).

Regarding the disease control service, population size of hosts has a complex effect on transmission dynamics. Through model-based analysis, Norman et al. (1999), and Gilbert et al. (2001) suggested that intermediate abundances of non-viraemic hosts (i.e. where pathogens do not enter the bloodstream) allow persistence in viraemic hosts, whereas high or low abundances lead to vector fadeout. Keesing et al. (2006) provided an example of how variation of population sizes of two rodent species through time affects disease spread: when there is a high density of chipmunks, there is a reduction in tick burdens on white-footed mice (the most competent reservoir for the Lyme bacterium). Losses in populations can lead to unpredictable effects on spread of vector transmitted diseases.

Decline in frugivorous populations reduce availability and quality of seed dispersal services (McConkey and Drake 2006, Peres and Palacios 2007). McConkey and Drake (2006) demonstrated that there is a threshold in population size for service provision; this is when the functionality of dispersers is lost, even before the individuals become rare. Therefore, the losses in disperser populations should not be dramatic to have a great impact on the seed dispersal service.

Thus, a decrease in abundance of FES providers has a negative impact on FES provision. Even though the reduction is small, the consequences on FES production can be significant given the complex interactions amongst the providers and the ecosystem functioning. Population losses imply more immediate effects than the loss of richness.