1 Coping with Global Environmental Change

HEXAGON Arieh Singer SERIES ON HUMAN AND ENVIRONMENTAL SECURIT Y AND PEACE / VOL 5 The Soils of Israel Hans Günter Brauch, Úrsula Oswald Spring, Czeslaw Mesjasz, John Grin, Patricia Kameri-Mbote, Béchir Chourou, Pál Dunay, Jörn Birkmann (Eds.) Coping with Global Environmental Change, Disasters and Security Threats, Challenges, Vulnerabilities and Risks This policy-focused Global Environmental and Human Security Handbook for the Anthropocene (GEHSHA) addresses new security threats, challenges, vulnerabilities and risks posed by global environmental change and disasters. In 5 forewords, 5 preface essays and 95 peer reviewed chapters, 164 authors from 48 countries analyse in 10 parts concepts of military and political hard security and economic, social, environmental soft security with a regional focus on the Near East, North and Sub-Sahara Africa and Asia and on hazards in urban centres. The major focus is on coping with global environmental change: climate change, desertification, water, food and health and with hazards and strategies on social vulnerability and resilience building and scientific, international, regional and national political strategies, policies and measures including early warning of conflicts and hazards. The book proposes a political geo-ecology and discusses a ‘Fourth Green Revolution’ for the Anthropocene era of earth history. Hans Günter Brauch, Adj. Prof. (PD) at the Free University of Berlin, chairman of AFES-PRESS, senior fellow at UNU-EHS in Bonn and editor of this series; he publishes on security and environment issues. Úrsula Oswald Spring, Professor at UNAM-CRIM, Mexico; first UNU-EHS chair on social vulnerability; she writes on sustainability, development, gender, disaster, poverty and collaborates with peasants.

Czeslaw Mesjasz, Assoc. Professor, Management, Cracow University of Economics; he publishes on systems and game theory, conflict resolution, negotiation, economics, finance and security. Patricia Kameri-Mbote, Professor, Strathmore University; Programme Director, International Environmental Law Research Centre, Nairobi; she writes on law, development, property, environment and gender. Béchir Chourou, Director, University of Tunis-Carthage, he taught International Relations at Tunis University; he publishes on Euro-Mediterranean relations, food policy and human security in the Arab world. Pal Dunay, Faculty Member, Geneva Centre for Security Policy, Director of International Training Course in Security Policy; he publishes on European security, the post-Soviet space and conventional arms control. Jörn Birkmann, Adj. Prof. (PD) at Bonn University, Head, Vulnerability Assessment, Risk Management and Adaptive Planning Section, United Nations University, Institute for Environment and Human Security.

Hans Günter Brauch Úrsula Oswald Spring Czeslaw Mesjasz John Grin

Patricia Kameri-Mbote Béchir Chourou Pál Dunay Jörn Birkmann (Eds.)

VOL 5 / HEXAGON SERIES ON HUMAN AND ENVIRONMENTAL SECURIT Y AND PEACE

1 Coping with Global Environmental Change, Disasters and Security

John Grin, Professor, Department of Political Science, University of Amsterdam; he publishes on societal transformations in water management, agriculture and health care, and advices practitioners.

Brauch Oswald Spring Mesjasz Grin Kameri-Mbote Chourou Dunay Birkmann (Eds.)

ISSN 1865-5793 ISBN 978-3-540-xxxxx-x

Coping with Global Environmental Change, Disasters and Security Threats, Challenges, Vulnerabilities and Risks

All chapters were anonymously peer reviewed. This is the third and final volume of the Global Environmental and Human Security Handbook for the Anthropocene (GEHSHA).

HESP / VOL 5

13

46

Vulnerability of Tropical Montane Rain Forest Ecosystems due to Climate Change Hans Juergen Boehmer

46.1

Introduction: Definitions and Key Concepts1, 2

Tropical montane rain forests provide important ecosystem services, such as supply, purification and retention of fresh water, regional water and air quality regulation, carbon sequestration, genetic and pharmaceutical resources, natural hazard and erosion regulation, recreation and ecotourism, etc.3 This type of ecosystem is highly dependent on stable conditions of several climate variables and, therefore, highly sensitive to any changes in those variables. For that reason, these forests provide excellent monitoring sites for detecting threats by climate change and illustrate its potential consequences for natural ecosystems and ecosystem services in an impressive way (Loope/Giambelluca 1998; Foster 2001). Before discussing this topic, however, some principal terms used in this chapter have to be defined.

1

2

3

I wish to thank Hans Günter Brauch, Helene H. Wagner, Dieter Mueller-Dombois, Corina Niemand, Nadja Rueger and Catherine Reynolds for helpful remarks. The comments of three anonymous reviewers significantly improved an earlier version of this manuscript. Keywords: biological invasions, catastrophe, cavitation, climatic anomalies, climatic fluctuations, cloud forests, cohort senescence, dieback, disaster, disturbance, disturbance regime, ecosystem services, extreme climatic events, foundation species, fragmentation, invasibility, long-term dynamics, natural disturbance, perturbation, population dynamics, resilience, tree demography, vegetation dynamics See Bruijnzeel (2004); Millennium Ecosystem Assessment (2005); Gullison, Frumhoff, Canadell, Field, Nepstad, Hayhoe, Avissar, Curran, Friedlingstein, Jones and Nobre (2007); Martinez, Pérez-Maqueo, Vázquez, Castillo-Campos, García-Franco, Mehltreter, Equihua and Landgrave (2009).

46.1.1

What is a Tropical Montane Rain Forest?

Like many terms in ecology, the term ‘montane rain forest’ has different connotations. Also, other terms have been used to describe forests that exist under the conditions outlined below (46.1.2, 46.1.3), like ‘cloud forest’, ‘elfin forest’, ‘mossy forest’, and ‘dwarf forest’ (Stadtmueller 1987; Richards 1996). This has caused some confusion; for instance, Beard (1955) juxtaposed ‘montane rain forest’ with ‘cloud forest’ (MuellerDombois/Ellenberg 1974, 2002), while Lamprecht (1977) used the terms synonymously. Bruijnzeel and Hamilton (2000) distinguish between lower montane forest, lower montane cloud forest, upper montane cloud forest, and subalpine cloud forest. ‘Cloud forest’ has been common for several decades,4 and is still widely used with a particular focus on tropical montane cloud forests.5 Like Whitmore (1998), the latest analysis of the world’s forests6 distinguishes between tropical upper montane forest and tropical lower montane forest. This chapter uses the term ‘tropical montane rain forest’ (van Steenis 1935; Beard 1944; Richards 1996). Its broader connotation includes all types of rain forests in tropical mountain ranges, from naturally open upper montane rain forests at higher elevations (characterized by stunted growth, twisted trunks, and epi-

4

5

6

See Beebe and Crane (1947); Ellenberg (1964); Troll (1968); Lawton and Dryer (1980); Sugden (1983); Stadtmueller (1987); Bruijnzeel and Hamilton (2000). See montane cloud forest (MCF): Hamilton, Juvik and Scatena (1995); Cayuela/Golicher/González-Espinosa/ Ramírez-Marcial/Rey Benayas 2006a; Cayuela/Rey Benayas/Echeverría 2006b); tropical montane cloud forest (TMCF): Bruijnzeel (2001); Foster (2001); Bubb, May, Miles and Sayer (2004). See Schmitt, Burgess, Coad, Belokurov, Besançon, Boisrobert, Campbell, Fish, Gliddon, Humphries, Kapos, Loucks, Lysenko, Miles, Mills, Minnemeyer, Pistorius, Ravilious, Steininger and Winkel (2009).

Hans Juergen Boehmer

790 phytic mosses) to dense lower montane rain forests with tall trees and a closed canopy (Whitmore 1998; Foster 2001). The latter are highly dependent on the frequent presence of clouds, but not on a permanent cloud cover (Stadtmueller 1987). Despite the attempted terminological clarity, it has to be kept in mind that there is a considerable diversity of montane rain forests within the tropics (Vareschi 1980; Kitayama 1995; Bruijnzeel 2001). Summing this up pragmatically, tropical montane rain forests are the predominating forest type in zones of maximum cloud condensation in mountain ranges of the tropics, and they frequently or permanently receive additional humidity through horizontal precipitation (direct canopy interception of cloud water, socalled cloud stripping).7 46.1.2

Structure and Biodiversity of Tropical Montane Rain Forests

Compared to tropical lowland rain forests, tropical montane rain forests have lower nutrient cycling rates and lower overall productivity (Grubb 1977; Whitmore 1989; Bruijnzeel 1998). Thus, trees have a lower stature with an average canopy height of about 15 to 35 meters, and increased stem density (Richards 1996; Whitmore 1998; Bruijnzeel 2001). Generally speaking, above-ground biomass decreases as elevation increases (Bruijnzeel/Hamilton 2000; Kitayama/Aiba 2002; Moser/Hertel/Leuschner 2007). At higher elevations, average canopy height can even be lower than 5 metres and trees have relatively small, thick and hard leaves (Richards 1996; Foster 2001). The following factors contribute to lower productivity in upper montane forests: reduced photosynthesis due to lower radiation and air temperatures; periodic water shortage on shallow soils; saturated soils and impeded root respiration; high soil acidity and, thus, low decomposition rates, nutrient mineralization and overall fertility.8 The complex topography of high mountains provides a large variety of important site factors and disturbance regimes (Boehmer/Richter 1997), thus creat7

Ataroff and Rada (2000), for instance, studied water fluxes in a Venezuelan montane rain forest (2300 m above sea level [a.s.l.] 3124 mm annual rainfall); there, 91 per cent of the incoming water was from rainfall, 9 per cent from cloudwater; in Holder’s study (2004) of a Guatemalan montane forest (2550 m), fog perecipitation contributed more than 7 per cent of the total input to the water budget, but was 19 per cent during the dry season.

ing a highly diverse environment with numerous mechanisms that maintain montane forest biodiversity (Zang/Tao/Li 2005). As a result, tropical montane rain forests have high levels of species diversity and endemism9 (Lewis 1971; Gentry 1995; Kessler 2001); long evolutionary time under relatively constant climatic conditions even made them biodiversity hot spots10. Gentry (1986, 1995) emphasizes natural habitat fragmentation and small-scale variation in climatic conditions as decisive triggers for high levels of endemism in tropical uplands. Regarding species composition and diversity, there are continuous elevational changes (Lovett 1998) as well as discrete compositional ecotones at certain altitudes (Martin/Sherman/ Fahey 2007; Schmitt/Denich/Frijs/Demissew/Boehmer in press). Ecotonal11 interactions, e.g. between lower and upper parts of watersheds, further stimulate species richness (Oesker/Dalitz/Günter/Homeier/Matezki 2008). Due to the high availability of water (which results from high precipitation and high relative humidity), epiphytes12 are usually the most frequent life forms aside from trees. Trees are covered by dense layers of vascular and non-vascular epiphytes (mosses, liverworts, lichens, ferns, orchids, etc.; Robins/Sugden 1979; Nadkarni 1984; Nadkarni/Matelson/Haber 1995). Moss coverage relative to that of lichens increases with elevation (Frahm/Gradstein 1991). High and largely non-seasonal annual rainfalls appear to be a driving force behind the diversity of epiphytes and terrestrial ferns (Gentry/Dodson 1987, Kessler 2001). Ferns and tree ferns (Cyatheaceae) are highly abundant and play an important role as understory species at all altitudes (figures 46.1, 46.2). In most of the lower montane forests, tree ferns are more common, while the upper montane forests are characterized by high abundances of filmy ferns (Hymenophyllaceae), which profit from the relatively

8

See Stadtmueller (1987), Bruijnzeel and Veneklaas (1998); Sollins (1998); Tanner, Vitousek and Cuevas (1998); Clark, Lawton and Butler (2000); Vitousek (2004). 9 Endemic species are restricted to a certain area or ecosystem (endemism). 10 See Terborgh (1977); Rohde (1992); Bush, Silman and Urrego (2004); Barthlott, Hostert, Kier, Küper, Kreft, Mutke, Rafiqpoor and Sommer (2008); Beck and Kottke (2008). 11 Ecotone refers to a transition zone between two structurally different (plant) communities. 12 Epiphytes refers to plants that grow attached to the stem and branches of trees.

Vulnerability of Tropical Montane Rain Forest Ecosystems due to Climate Change

791

Figure 46.1: Montane rain forest on the island of Hawaii, Kilauea area, ca. 1200m above sea level. Source: Photograph by Jutta Pscherer (2002).

cool, wet microclimate inside the stands (Richter 2008). In some montane rain forest types, lianas are characteristic features as well (Senbeta/Schmitt/Den-

ich/Demissew/Vlek/Preisinger/Woldemariam/Teketay 2005). The overall richness of vascular plant species usually decreases substantially from lower to upper


792 elevations (Kitayama 1992; Vásquez/Givinish 1998; Kessler 2001). Tree species composition varies with altitude (Lieberman/Lieberman/Peralta/Hartshorn 1996), but tree species richness does not necessarily decline with elevation (Lovett 1998, 1999). Tropical montane rain forests harbour highly diverse faunas with numerous co-existing animal species that are limited to certain habitat types within this zone.13 Habitat specialization of birds, for instance, generates high diversity (Willis/Schuchmann 1993; Watson/Peterson 1999; Jankowski/Ciecka/Meyer/ Rabenold 2009). 46.1.3

Distribution of Tropical Montane Rain Forests

The frequent or permanent presence of a dense cloud layer at certain elevation intervals along mountain ranges is a key condition for the existence of tropical montane rain forests;14 they are strongly linked to regular cycles of cloud formation (Still/Foster/Schneider 1999). In tropical high mountains there is a general correlation between cloudiness and orographic rainfall (Sarmiento 1986). The average rainfall in those areas ranges from 1200 mm to over 7500 mm per year (Evenson 1983; Bush/Hanselman/Hooghiemstra 2007), with a mean annual temperature of 18–20°C at the lower elevation limit and about 10°C at the upper limit (Bruijnzeel 2001). However, short periods of drought (infrequent mild droughts) can be considered a normal element of the natural environment (Giambelluca/Nullet/Ridgley/Eyre/Moncur/Price 1991; Werner 2003) that affects or even determines population dynamics of certain foundation plant species.15 Exceptionally wet conditions have been described for a wide range of tropical mountains, causing reduced temperature oscillations, solar irradiation, and

13 See Orians (1969); Terborgh (1971); Hernández-Baños, Peterson, Navarro-Sigüenza, Escalante-Pliego 1995; Watson and Peterson (1999); Blake and Loiselle (2000); Brehm, Homeier and Fiedler (2003); Herzog, Kessler and Bach (2005). 14 See Grubb and Whitmore (1966); Huber (1976); Nair, Asefi, Welch, Ray, Lawton, Manoharan, Mulligan, Sever, Irwin, Pounds (2008). 15 A foundation species is a dominant species in an ecosystem that controls the populations of other species (and overall community dynamics) and modulates ecosystem processes (Ellison/Bank/Clinton/Colburn/Elliott/Ford/ Foster/Kloeppel/Knoepp/Lovett/Mohan/Orwig/Rodenhouse/Sobczak/Stinson/Stone/Swan/Thompson/Von Holle/Webster 2005).

evapotranspiration of trees (Hedberg 1964; Sarmiento 1986; Stadtmueller 1987; Hamilton/Juvik/Scatena 1995). Here, seasonality is primarily based on fluctuations in horizontal and vertical precipitation (Richter 2008). In many places there are other extreme site factors like insufficient soil drainage (Hamilton/Juvik/ Scatena 1995, Mueller-Dombois 2006), nutrient leaching (Vitousek 2004), or strong winds (e.g. along the trade wind belt; Lawton 1982; Sugden 1986). The band of maximum cloud cover in the humid tropics is highly dependent on the topography of the principal mountain ranges (e.g. Bendix/Rollenbeck/ Goettlicher/Cermak 2006). Generally it is located between 1200 and 2500 m above seal level (asl; Stadtmueller 1987). Nevertheless, in many regions the belt of montane rain forest begins at lower elevations. Lower extremes tend to be ca. 800 m asl, but can be lower than 500 m, e.g. Seychelles (Flenley 1974) or the tropical Pacific islands (Mueller-Dombois/Fosberg 1998), and, on the high end, Hueck (1978) mentions Andean cloud forests even at 3900 m asl. A variety of factors affect the intensity, amount, and duration of upslope precipitation, including barrier width, slope steepness, and updraft speed. In addition, other climatic and geographic factors like the direction of the prevailing winds, micro relief, and disturbance regimes determine the distribution of montane rain forests and indicate their limits (Richter 2008). The position of the belt of maximum precipitation in the mountains also depends on the degree of aridity in the adjacent forelands. That means that severe changes in foreland ecosystems (deforestation) can affect nearby montane rain forests (Lawton/Nair/ Pielke/Welch 2001, Nair/Lawton/Welch/Pielke 2003). The distribution of tree ferns is a prominent indicator for the geographical boundaries of montane rain forests in the tropics (Kroener 1967; Troll 1970; Stadtmueller 1987; figure 46.1, 46.2). Troll (1956, 1968) developed a pioneering profile of altitudinal vegetation belts that includes a montane forest belt in the inner tropics, ranging from tropical temperate up to cold altitudes. Numerous case studies have helped to create a clearer picture of tropical montane rain forest distribution (Bruijnzeel 2001). Today, the horizontal extension of tropical montane rain forests is calculated at 924.000 km2, approximately 15 per cent of the existing tropical rain forest. Among those, tropical upper montane rainforests cover 476.000 km2, and tropical lower montane rainforests 448.000 km2.16 In the American humid tropics, the principal areas of montane rain forest distribution include Mexico (Hastenrath 1968; Cayuela/Golicher/Rey-Benayas


793

Figure 46.2: Treefern-dominated understory in a montane rain forest on the island of Hawaii, Puu Makaala area, ca. 1270m above sea leavel. Source: Photograph by Hans Juergen Boehmer (2003).

2006), Central America (Knapp 1965), particularly Guatemala (Holder 2005), Costa Rica (Zadogra 1981, 16 See Schmitt, Burgess, Coad, Belokurov, Besançon, Boisrobert, Campbell, Fish, Gliddon, Humphries, Kapos, Loucks, Lysenko, Miles, Mills, Minnemeyer, Pistorius, Till Ravilious, Steininger, Winkel 2009a, 22009b).

Nadkarni/Wheelwright 2000), Panama (Myers 1969), Jamaica (Tanner 1977), Puerto Rico (Weaver 1972a, 1972 b), Ecuador (Gradstein/Homeier/Gansert 2008; Beck/Bendix/Kottke/Makeschin/Mosandl 2008), Venezuela (Veillon 1955, 1974), Colombia (Sugden 1982a, 1982b, 1982c), and Peru (Young/León 1995). The high mountains in Southeast Asia, Malaysia, the


794 Philippines and Indonesia, in particular, are centres of montane rain forest distribution (Hamilton/Juvik/ Scatena 1995). New Guinea and the principal islands of the Pacific region (Hawaii, Samoa) have considerable amounts of tropical montane rain forests as well (Merlin/Juvik 1995; Mueller-Dombois/Fosberg 1998). In Africa, Mt. Cameroon and the high mountains of East Africa are worthy to mention (Hedberg 1951; Friis 1992, Hemp 2006). Even in relatively dry regions such as the Simien Mountains in northern Ethiopia, montane rain forests form a discrete belt on northeastern escarpments at 2200 to 2500 m a.s.l. (Richter 2008), which again illustrates the major significance of horizontal precipitation in this type of ecosystem. 46.1.4

Climate Change and Specific Vulnerability

Climate change is one of several global processes (including rapid urban development, population growth, etc.) that require special attention because it is affecting the distribution of resources and increasing the vulnerability of ecosystems. In IPCC (2007: 30) terminology, climate change refers to “a change in the state of the climate that can be identified (e.g. using statistical tests) by changes in the mean and/or the variability of its properties, and that persists for an extended period, typically decades or longer.” In short, “it refers to any change in climate over time, whether due to natural variability or as a result of human activity.” According to Cardona (chap. 3 above), vulnerability of ecosystems is “intrinsically tied to different socio-cultural and environmental processes, but (…) is also related to the fragility, the susceptibility, or the lack of resilience of the exposed elements, both from society and environment” (for overviews on the concepts of ‘vulnerability’, ‘risk’, etc. see Brauch 2005a; Birkmann 2006; chap. 2 by Brauch). In the context of climate change, vulnerability is “a function of the sensitivity of a system to changes in climate (the degree to which a system will respond to a given change in climate, including beneficial and harmful effects), adaptive capacity (the degree to which adjustments in practices, processes, or structures can moderate or offset the potential for damage or take advantage of opportunities created by a given change in climate), and the degree of exposure of the system to climatic hazards” (IPCC 2001). Interestingly, the concepts of disaster, catastrophe, disturbance, perturbation, and resilience are applied differently in the scientific and political discussion on human and environmental insecurity from the way

they are commonly understood in ecology (Harper 1977, White/Pickett 1985; Grimm/Wissel 1997; Brand/Jax 2007). This lack of interdisciplinary terminological clarity cannot be discussed here in detail; to give an example, Turner II, Kasperson, Matson, McCarthy, Corell, Christensen, Eckley, Kasperson, Luers, Martello, Marybeth, Polsky, Pulsipher and Schiller (2003: 8074) define hazards as threats to a system, comprised of perturbations and stress (and stressors), and the consequences they produce. A perturbation is a major spike in pressure (e.g., a tidal wave or hurricane) beyond the normal range of variability in which the system operates. Perturbations commonly originate beyond the system or location in question. Stress is a continuous or slowly increasing pressure (e.g., soil degradation), commonly within the range of normal variability. Stress often originates and stressors (the source of stress) often reside within the system. Risk is the probability and magnitude of consequences after a hazard (perturbation or stress).

These concepts of perturbation and stress do not fit with the broadly accepted definition of the terms in the natural sciences. There has been a long discussion in ecology about the definitions of terms like perturbation, disturbance, stress, and catastrophe (Harper 1977; White/Pickett 1985; Boehmer/Richter 1997; White/Jentsch 2001; Kammer/Moehl 2002). Widely accepted is the definition of disturbance as “any relatively discrete event in time that disrupts ecosystem, community, or population structure and changes resources, substrate availability, or the physical environment” (White/Pickett 1985: 5ff.). An ecological event is an “abruptly occurring process which significantly affects variables of an ecological unit for a period longer than its own duration. It must have a clear beginning and end, and must be of relatively short duration in relation to the time scale of the organisms considered” (Grimm/Stillman/Jax/Goss-Custard 2007: 60). The term ‘perturbation’ describes a departure from an explicitly defined normal state, behaviour or trajectory (White/Pickett 1985). Unfortunately, in the definition by Turner II, Kasperson, Matson, McCarthy, Corell, Christensen, Eckley, Kasperson, Luers, Martello, Marybeth, Polsky, Pulsipher and Schiller (2003: 8074) the term ‘perturbation’ is used as a synonym for the ecological term ‘disturbance’, more precisely for what is defined in ecology as a catastrophic disturbance (large infrequent disturbance [LID] after Turner/Baker/Peterson/Peet 1998). In ecology, catastrophes or ‘catastrophic events’ are regarded as special cases of disturbance. Events of this nature are unpredictable and cannot be included in the prevailing regime of

Vulnerability of Tropical Montane Rain Forest Ecosystems due to Climate Change regularly occurring disturbances (Boehmer/Richter 1997; Turner/Baker/Peterson/Peet 1998); accordingly, foundation species cannot become adapted to such events (Harper 1977; Boehmer/Richter 1997). On the other hand, a ‘disaster’ recurs so frequently that it becomes relevant for the fitness of successive generations, because it can be expected to occur within the species´ life cycle: “The selective consequence of disasters is therefore likely to be to increase short-term fitness and the consequence of catastrophes is to decrease it” (Harper 1977: 627); so, disasters and catastrophes can – from the viewpoint of ecology – be interpreted as the ends of a continuum. In contrast, a disaster as discussed in the context of human and environmental insecurity implies “loss and damage, and consequential impacts that the affected community is unable to absorb or to cushion the effects and recover using its own resources and reserves” (chap. 3 by Cardona). Furthermore, the distinction between ecological events (such as disturbance) and continuous processes (such as perturbation) as two fundamental types of ecological processes is central to any ecological research on ecosystem dynamics (Grimm/Stillman/ Jax/Goss-Custard 2007). To avoid confusion, the terms ‘disturbance’ and ‘perturbation’ will be used here according to White and Pickett (1985), following the introduced and accepted terminology of ecology. ‘Stress’ is commonly defined as a limitation of dry matter production, while disturbance is connected to the destruction of biomass (Grime 1979; Kammer/ Moehl 2002). Another problem with Turner II’s definition is the boundary between disturbances originating from ‘beyond the system’ (commonly named ‘exogenous disturbances’, not ‘perturbation’) and those originating from ‘within the system’ (commonly named ‘endogenous disturbances’, not ‘stress’). This boundary appears to be conceptually precise but is difficult to grasp in terms of structural content (Boehmer/Richter 1997). The montane rain forests of the tropics are considered – as stated above – fragile exposed elements with a lack of resilience. Bioclimatic models suggest that major impacts to tropical forests will result from climate change (Hannah/Midgley/Andelman/Araújo/ Hughes/Martinez-Meyer/Pearson/Williams 2007); complex synergetic effects on essential ecological factors should be expected (see 46.2). Vulnerability is considered for the purposes of this chapter as having the following aspects: 1) destruction of the established ecosystem ‘tropical montane rain forest’ through substantial changes in the determining cli-

795

mate variables; 2) impact of natural disturbances; 3) enduring disruption of ecosystem processes (irreversible modification of natural long-term dynamics); 4) loss of indigenous biodiversity; and 5) loss of ecosystem services. Nevertheless, when emphasizing the importance of climatic factors for plant species distribution, it must not be forgotten that most species depend on a much more complex system of factors that describe their ecological niche (Heger/Boehmer 2005; Svenning/Kerr/Rahbek 2009). Non-climatic stresses affecting species survival, exacerbating problems caused by climate change, or limiting species’ ability to respond to environmental changes should also be taken into account. Detailed studies of all ecological factors relevant for distribution and spread still do not exist for the vast majority of plant species. Reliable predictions of future distributions are, therefore, hardly possible (Grainger 2008), although attempted frequently (Thullier/Lavorel/Araújo 2005; Thullier/Lavorel/ Araújo/Sykes/Prentice 2005). Even so, any assessment of ecosystem vulnerability should consider the full spectrum of possible impacts (Clark 2007; Feeley/Silman 2008), even those not yet quantifiable in detail (e.g. adverse effects of climate change, including climate variability and extremes).

46.2

Climate Change and Tropical Montane Rain Forest Dynamics

The availability of water determines the distribution of plants in terrestrial ecosystems (Larcher 1995). Climatic factors such as temperature, precipitation, and humidity play a fundamental role in the physiology and ecology of plant species (Harper 1977). Changes in those factors can, therefore, cause environmental stress and pose a threat to the regeneration and reproduction of numerous species. To survive such changes, species must be either resilient to climatic extremes, adapt to the new conditions, or migrate to other suitable areas (Davis/Shaw/Etterson 2005; Parmesan 2006). In plant communities, species react individually to climatic change (Hannah/Lovejoy 2007). Alterations in climate affect forest communities by changing the abundance and competitive capacity of different species in the forests (Hannah/ Midgley/Andelman/Araújo/Hughes/Martinez-Meyer/ Pearson/Williams 2007). Altered rainfall patterns, surface temperatures, and light availability lead to a change in carbon fixation by plants (primary production), higher mortality, and changing regeneration suc-


796 cess of many plant species (Holmgren/Scheffer/Ezcurra/Gutiérrez/Mohren 2001; Stenseth/Mysterud/ Ottersen/Hurrell/Chan/Lima 2002). Climate change threatens tropical montane rain forests particularly by altering precipitation, humidity, temperature, ultraviolet radiation, storm frequency and duration (Hamilton/Juvik/Scatena 1995, Foster 2001). In tropical high mountains, vascular plant species composition is associated with cloud frequency (Sklenár/Bendix/Balslev 2008). Bush (2002) hypothesized that the elevation of cloud formation will increase due to climate warming, caused by an upslope movement of the cloud base and resulting in a narrower elevational interval supporting the montane rain forest. A prolonged uplifting of the cloud layer can, for example, induce a decline in canopy trees (dieback, see below), enhanced seedling recruitment, population growth among understory plant species, and an increase in fruit production. Increases in air temperature can be linked to community changes of birds and amphibians in montane rain forests (Pounds/Fogden/Campbell 1999). Projected distributions of novel and disappearing climates by 2100 AD show that disappearing climates are concentrated in tropical montane regions (Williams/Jackson/Kutzbach 2007), thus indicating an increased extinction risk for species with narrow geographic distributions. Climatic extremes, rather than mean values, have crucial impact on the dynamics of vegetation (Condit/Hubbell/Foster 1996). The increase of climatic extremes and fluctuations in recent decades has profound effects on the highly specific conditions outlined above, thus affecting species composition, structure, and dynamics of tropical montane rain forests. Climatic fluctuations such as extreme wet-dry oscillations can act as triggers of forest decline, particularly on marginal sites with shallow soils or poor drainage (e.g. Mueller-Dombois 2006; see below). Rapid changes in climate and atmospheric chemistry threaten sustainability of forests (Wargo/Auclair 2000). Droughts, for instance, have a severe impact on terrestrial vegetation dynamics, particularly in tropical rain forests.17 Fortunately, tropical mountains are characterized by a variety of local climates; therefore,

17 See Lowry, Lee and Stone (1973); Leighton and Wirawan 1986; Oren, Zimmermann and Terborgh (1996); Walsh and Newbery (1999); Nakagawa, Tanaka, Nakashikuza, Ohkubo, Kato, Maeda, Sato, Miguchi, Nagamasu, Ogino, Teo, Abang and Lee (2000); Potts (2003); Williamson and Ickes (2002); Werner (2003).

drought stress does not influence all parts of a montane rain forest in a given region in the same way. Climatic extremes can arise from the shifts in tropical precipitation patterns, as in the case of El Niño anomalies18 in the central and eastern Pacific, which have caused droughts in Indonesia (Werner 2003; Slik 2004) and inundated the South American west coast and the islands of the central Pacific with heavy rains (McPhaden/Zebiak/Glantz 2006). Remote effects of such changes are recorded in the North and South Pacific and beyond; strong El Niño and La Niña events have worldwide consequences (McPhaden/ Zebiak/Glantz 2006). Severe ENSO-droughts have been increasing in recent decades.19 On a local scale, extreme weather events like tropical storms can damage trees by floods, defoliation, and windfall, thereby causing tree fall gaps that affect neighbouring trees as well as understory plants. Landslides, a typical geomorphic process in humid high mountain regions, create large-scale disturbances in montane forests, and become more frequent after extreme precipitation events. These disturbances may lead to a decrease in the connectivity of populations by fostering forest fragmentation. Nevertheless, there is also evidence from palaeological records that tropical montane rain forests have been a continuous and constant feature, e.g., of the Andean environment throughout the Holocene (Bush/Hanselman/Hooghiemstra 2007). Here, rates of community change were low for thousands of years, despite considerable variability in the climate. Unfortunately, the warming expected in the near future has no physical and biotic analogues in the past (Hannah/Lovejoy 2007). 46.2.1

Climate-Induced Forest Decline

Future warming is expected to exacerbate regional tree species die-off 20 worldwide (Adams/Guardiola-

18 El Niño or El Niño-Southern Oscillation (ENSO): periodic change in the atmosphere and ocean temperature of the tropical Pacific; manifested in the atmosphere by changes in the pressure difference between Tahiti and Darwin, Australia, and in the ocean by warming or cooling of surface waters of the tropical Eastern Pacific Ocean 19 See McClure (1983); Loope and Giambelluca (1998); Boehmer and Niemand (2009); Boehmer, Niemand, Gerrish, Jacobi and Mueller-Dombois submitted for publication. 20 Die-off refers to a sudden sharp decline of a population (animals or plants).

Vulnerability of Tropical Montane Rain Forest Ecosystems due to Climate Change Claramonte/Barron-Gafford/Villegas/Breshears/Zou/ Troch/Huxman 2009) thereby affecting regional carbon budgets. The suppression or loss of foundation tree species in montane rain forests has acute and chronic impacts on fluxes of energy and nutrients, hydrology, food webs, and biodiversity.21 It is well known that climatic fluctuations like ENSO have strong effects on the dynamics of plant populations in a wide range of terrestrial ecosystems (Curran/Caniago/Paoli/Astianti/Kusneti/Leighton/Nirarita/Haeruman 1999; Holmgren/Scheffer/Ezcurra/ Gutiérrez/Mohren 2001). El Niño events bring drier, warmer, and sunnier conditions to the wet tropics, resulting, e.g., in a prolonged uplifting of the cloud layer, thus causing stress, e.g., in trees growing on waterlogged substrates. For example, as a result of the 1982–1983 super-El Niño event, which brought record temperatures and severe drought to Indonesia (Leighton/Wirawan 1986), 71 per cent of the canopy trees (>60 cm diameter) died in forest plots on dry ridges and slopes (dieback). ‘Dieback’ is a special form of forest decline, which affects only canopy trees and is known from several of the world’s forest ecosystems (Jane/Green 1983; White 1986; Lawesson 1988; Mueller-Dombois 1988; Stewart 1989; King/Neilson 1992; figure 46.3). The term has been defined as “progressive dying back from the tips of twigs, branches or tops” (Podger 1981; Manion 1981; Ciesla/Donaubauer 1994). As far back as 1986, it was shown that forests are highly susceptible to dieback under ongoing climate warming (Solomon 1986, Mueller-Dombois 1987, 1988). After a dwindling of interest in the late 1980’s, forest dieback resulting from climate change has started to become a major focus of ecological research again (Allen 2009). An important aspect is the assumption that in many forest ecosystems neither floods nor droughts lead to the death of canopy trees. Instead, the combination of both, due for example to more rapid precipitation and irradiation changes, seems to have lethal effects, especially on marginal sites (Auclair 1993; Mueller-Dombois 2006). The phenomenon of stand-level dieback in forests, formerly interpreted as a local symptom of a disease or a pest attack (Petteys/Burgan/Nelson 1972), is nowadays understood as a complicated combination of biotic and abiotic factors that includes generic levels as well as climatic

21 See Ellison, Bank, Clinton, Colburn, Elliott, Ford, Foster, Kloeppel, Knoepp, Lovett, Mohan, Orwig, Rodenhouse, Sobczak, Stinson, Stone, Swan, Thompson, Von Holle and Webster (2005).

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triggers (Mueller-Dombois 1988). According to Auclair (1991), region-wide, persistent episodes of forest dieback are incited by extreme moisture stresses in the Pacific area. While extreme precipitation events typically are local and short-lived phenomena, the dissolving of the cloud layer during ENSO anomalies may affect large areas over extended periods of time. It is thus assumed that large-scale dieback is triggered by a rapid succession of climatic extremes. For example, high radiation, followed by extreme precipitation and repeated high radiation (Auclair 1991, 1993), may cause lethal cavitation22 in forest stands. Four types of climatic fluctuations can cause cavitation injury in tropical montane rain forests: 1) wet-dry oscillations; 2) intermittent high pressure or blocking high; 3) extreme rainfall; and 4) extreme drought. Dry-wet oscillations should be considered as dieback triggers as well. The frequency of such fluctuations has been increasing in recent decades.23 Usually, forest stands do not die while in full vigour, except from catastrophic disturbances, diseases or insect pests. In the absence of such influences, an endogenous predisposition to decline can be assumed, which may be related to the structure and dynamics of the tree population itself (Mueller-Dombois 1987). For instance, after intensive research on the forest decline on the Island of Hawaii in the early 1970’s, it was shown that neither disease nor climatic stress actually caused the mass-dieback of the dominating tree species Metrosideros polymorpha (Papp/Kliejunas/Smith/Scharpf 1979; Hodges/Adee/Stein/Wood/ Doty 1986). Instead, attention focused on a complex array of demographic factors, namely a uniform age and stand structure (‘cohorts’; Mueller-Dombois 1983, Gerrish/Mueller-Dombois 1999), probably resulting from large-scale disturbances, e.g., through volcanic eruptions, leading to the simultaneous aging and finally death of the cohort (‘cohort senescence’; Mueller-Dombois 1987). The concept of cohort senescence, which claims that old cohorts of canopy trees are predisposed to die, due to their low energy potential for recovery from stress, has particular implications for climate 22 Cavitation refers to a catastrophic hydraulic failure in the xylem (water transport tissue) of vascular plants when the tension of water within the xylem becomes so great that dissolved air within the water expands (bubbles), thus blocking the transport of stem water. 23 See Loope and Giambelluca (1998); Boehmer and Niemand (2009); Boehmer, Niemand, Jacobi, Gerrish and Mueller-Dombois (submitted for publication).

798


Figure 46.3: Dieback of the foundation species Metrosideros polymorpha, montane rain forest, Mauna Kea, island of Hawaii. Sources: a) aerial photograph by Dieter Mueller-Dombois (1972); b) Jutta Pscherer (2002).

change. It suggests that aging forests are susceptible to synchronized decline under certain demographic conditions (Boehmer/Wagner/Mueller-Dombois submitted for publication). Such weak stages in stand

demography increase the vulnerability of forests. Extreme rainfall events or prolonged droughts, or combinations of these factors could generate a physiological shock that may trigger dieback among old,

Vulnerability of Tropical Montane Rain Forest Ecosystems due to Climate Change stressed trees (Mueller-Dombois 1987; Gerrish/Mueller-Dombois/Bridges 1988, Gerrish 1990). Santiago/Goldstein/Meinzer/Fownes/Mueller-Dombois (2000) propose that poor drainage causes roots to avoid hypoxic24 soil horizons. This in turn could reduce the leaf area. Dieback symptoms may, therefore, represent an adjustment in leaf area to available resources, primarily in large trees that have a limited rhizosphere, while saplings and smaller trees are frequently established on nurse logs that provide more oxygenated sites. Further research is needed to test the relationship between the age of cohorts and the effect of climatic anomalies on their vitality. 46.2.2

Natural Disturbances

In recent decades natural disturbance has increasingly been recognized as a principal factor controlling the structure and dynamics of ecosystems.25 Natural disturbances are a major driving factor behind vegetation dynamics in humid high mountain regions at different scales (Boehmer/Richter 1997; Boehmer 1999; Kammer/Moehl 2002). Tropical montane rain forests are subject to numerous natural disturbance factors (e.g. storms, landslides, fires, volcanic eruptions).26 These factors cause a pattern of coexisting successional stages including moss carpets, herbaceous stands, fern thickets, scrubs, and mature forest. This habitat heterogeneity, e.g., in areas of frequent landslides, explains the high species richness of many tropical montane rain forest ecosystems. This is especially true for mountain regions characterized by perhumid conditions and frequent earthquakes (Richter 2008). There is a critical threshold for inducing landslides that depends on the following factors in combination: mechanical and hydrological soil parameters and the destabilization of the watersoaked organic layer by the forest load (Asch/Deimel/Haak/Simon 1989). Extreme precipitation events destabilize slopes, particularly where montane forests are already fragmented by wind-induced tree fall or clearing. If larger areas are affected, a change in

24 Hypoxic soils have a lack of oxygen as a result of poor drainage or during short-term flooding. 25 See Levin and Paine (1974); Connell and Slatyer (1977); Connell (1978, 1979); Sousa (1984); Pickett and White (1985); Boehmer and Richter (1997); Turner, Baker, Peterson and Peet (1998); White and Jentsch (2001). 26 See Clark (1990); Waide, Zimmerman and Scatena (1998); Whitmore and Burslem (1998); Chazdon (2003); Cochrane (2003).

799

the local climate and an increase in the invasibility of the ecosystem can result. Climate change is expected to cause an increase in natural disturbance impact, e.g., from tropical storms (Markham 1996; Elsner/Kossin/Jagger 2008) although there is at present no clear trend in the number of tropical cyclones (IPCC 2007). Such events can alter forest composition and structure for many years. Usually, storms cause more damage on ridges than in valleys and reduce the vegetation in upper height intervals significantly. Average maximum canopy height, for instance, decreased as much as 50 per cent in Puerto Rico’s lower montane rainforest after Hurricane Hugo in 1989 (Brokaw/Grear 1991). Nevertheless, local variations in hurricane damage (or disturbance impact, respectively) can increase complexity of forest structure and, in doing so, provide habitats that are essential or favourable for a significant number of plant and animal species (Petraitis/Latham/Niesenbaum 1989). Coarse woody debris (CWD), produced when branches fall to the ground, provides important safe sites for seedling establishment on the forest floor (Santiago/Goldstein/Meinzer/Fownes/Mueller-Dombois 2000). In addition, the fine litter fall caused by tropical storms alters nutrient cycling in the forest system (Lodge/Scatena/Asbury/Sánchez 1991). 46.2.3

Range Retractions, Expansions and Biological Invasions

A warming of tropical mountains of 2 to 3°C (IPCC 2001) might cause an upslope migration of species,27 thus worsening conditions for the species in montane rain forests that are highly adapted to cloud water (Benzing 1998). Many plant species in the Amazon cloud forest, for instance, may not survive the climate changes forecast to occur within the next 100 years (Bush/Silman/Urrego 2004). Nadkarni and Solano (2002) showed that less cloud water generally has a negative effect on epiphyte growth and leaf production and radically alters the composition of canopy communities. As the mortality of montane rain forest species increases, the competition pressure induced by invasive indigenous lowland species is simultaneously heightened. Short migratory distances from low-

27 See Pounds, Fogden and Campbell (1999); Peterson, Ortega-Huerta, Bartley, Sánchez-Cordero, Soberón, Buddemeier and Stockwell (2002); Raxworthy, Pearson, Rabibisoa, Rakotondrazafy, Ramanamanjato, Raselimanana, Wu, Nussbaum and Stone (2008); Sekercioglu, Schneider, Fay and Loarie (2008).


800 lands to the upper forest line make montane rain forests sensitive monitors of climate change (Bush/ Flenley 2007). Such range expansions are widely expected, but have not yet been studied in detail (Colwell/Brehm/Cardelús/Gilman/Longino 2008). Invasion by non-indigenous species is one of the most important issues in today’s applied ecology (Convention on Biological Diversity; United Nations 1993). Some alien species are able to cause fundamental changes in indigenous ecosystems, including the local extinction of native species (SCBD 2001). The weakening of indigenous foundation tree species through climate change increases the invasibility of forests. Declining forests become more vulnerable to biological invasions, for instance, because dying trees can be replaced by invading alien tree species. These occupy available sites before the indigenous species can establish new individuals or a cohort (Boehmer 2005). The increased invasibility of montane rain forests by canopy decline makes invasive alien species capable of completely changing structure and species composition (Asner/Hughes/Vitousek/Knapp/KennedyBowdoin/Boardman/Martin/Eastwood/Green 2008). An impressive example of the negative potential for a synergetic effect presented by invasive species is Kahili Ginger (Hedychium gardnerianum), a 2 metre tall herb native to India’s Himalaya region. It was introduced to many tropical regions as an ornamental plant; today it is listed as invasive in most of those regions (CAB International 2005). In its natural range, Kahili Ginger prefers open, well-lighted environments. Nevertheless, in tropical montane rain forests, it also occurs in semi- and full shade beneath the forest canopy (Stone/Pratt 1994). For instance, in Hawaii’s montane rain forest, native species cannot compete in the infested sites; no regeneration of indigenous or endemic plant species within Kahili Ginger stands can be observed (Minden/Jacobi/Porembski/Boehmer in press). Furthermore, the colonization by Kahili Ginger provides favourable conditions for the spread of the invasive non-indigenous Strawberry Guava (Psidium cattleianum) tree. In the case of climate induced canopy decline, there will be no considerable regeneration of native rain forest plants; an almost total loss of native tree species (including the foundation species Metrosideros polymorpha) can be expected, and a new exotic rain forest type, dominated by two alien plant species, is likely to persist (Boehmer 2005; Minden/Hennenberg/Porembski/Boehmer 2009; Minden/Jacobi/Porembski/Boehmer in press).

46.2.4

Negative Synergetic Effects of Human Impact

Humans have inhabited tropical forests for thousands of years (Piperno/Bush/Colinvaux; Roosevelt 1999). Consequently, many tropical montane rain forests have witnessed multiple human impact too (Bruijnzeel/Hamilton 2000), although they did not become subject to intense human disturbance until the mid20th century (Kellman/Tackaberry 1997). Human disturbances like forest clearing and selective logging directly affect forest structure, leading to widespread fragmentation (Wilder/Brooks/Lens 1998; Laurance 2007; Hemp 2009), thereby entailing complex edge effects28 within the forest fragments.29 Recent analyses show a continuous decrease in the area covered by montane rain forests, with a simultaneous increase in fragmentation, e.g., in Mexico between 1976 and 2003 (Cayuela/Golicher/González-Espinosa/RamírezMarcial/Rey Benayas 2006a; Cayuela/Rey Benayas/ Echeverría 2006b) and East Africa (Schmitt/Denich/ Friis/Demissew/Boehmer in press; Hemp 2009). Even in areas of intact forest cover, indigenous people may have harvested forest products (figure 46.4), introduced useful species, or hunted, thus affecting abundance of important seed dispersers and seed predators.30 Considerable effects may also have resulted from changes in coffee cultivation from shaded to non-shaded systems (semi-forest coffee systems; Senbeta/Denich/Boehmer/Woldemariam/ Demel/Demissew 2007). The complex consequences of such human activities in montane tropical forest areas are still not well understood. Lawton/Nair/Pielke/Welch (2001) showed that regional deforestation of upwind lowland forest influences cloud patterns in nearby mountains. In a tropical mountain region of southern Mexico, for instance, the agricultural conversion from orchards to dry farming in the coastal lowlands during the 1970’s and 1980’s resulted in an upward flux of overheated air masses and less cloud frequency in the mountains, thus causing a regional climate change (Richter 2008). 28 Edge effects are environmental changes (greater desiccation stress, wind turbulence, litter fall, etc.) associated with the abrupt, artificial boundaries of forest fragments. 29 See Laurance, Nascimento, Henrique, Laurance, Andrade, Ribeiro, Giraldo, Lovejoy, Condit, Chave, Harms and D’Angelo (2006). 30 See Senbeta and Denich (2006); Dietz, Hölscher, Leuschner, Malik and Amir (2007); Gradstein, Kessler and Pitopang (2007).


801

Figure 46.4: Managed montane rain forest with indigenous coffee shrubs (Coffea arabica), Kayakela Forest, Kafa zone, south-west Ethiopia. Source: Photograph by Manfred Denich (2004).

Significant increases in stand precipitation, potential evaporation, and infiltration rates are contrasted by a significant decrease in water holding capacity. One of the consequences is a hundred- to thousand-fold rise in soil loss (compared to land under traditional cultivation) and a dramatic change in stream drainage. The most damaging result is a higher runoff rate during the wet period, with exuberant river discharges, as demonstrated by the floods during hurricane Stan in September 2005. This event led to extended sedimentations in the foreland of the mountain chain and is one of many examples of the consequences of the destruction of the water and soil reservoirs of tropical mountain forest ecosystems. Neither protecting the impacted lowland area by artificial dams and retention ponds, nor securing the slide areas by wire nets can solve the problem in the long term (Richter 2008). Disturbance and deforestation, in combination with climate change, threaten the regional distributions of endangered tree species in tropical montane

forests (Golicher/Cayuela/Alkemade/González-Espinosa/Ramírez-Marcial 2008). Strong forest fragmentation leads, for instance, to disruptions of mutual relationships between plants and animals, of seed disperser-seed interactions (reduced avian visitation of bird-dispersed trees), thus possibly posing long-term effects on tree species abundance and population dynamics (Lehouck/Spanhove/Colson/Adringa-Davis/ Cordeiro/Lens 2009). Conservation concepts should, therefore, consider a multi-site approach with several protected areas at different altitudinal levels to cover the diversity of species and communities (Kessler 2001; Schmitt/Denich/Friis/Demissew/Boehmer in press). On abandoned land, regeneration of montane rain forests and de-fragmentation is possible as long as primary forest fragments still exist close to the disturbed sites (Karlowski 2006; Muñiz-Castro/WilliamsLinera/Rey Benayas 2006).t


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46.3

Conclusions

Climate change increases the vulnerability of tropical montane rain forests e.g. by intensifying the frequency of climatic anomalies and extreme weather events, and increasing natural disturbances. A variety of human impacts additionally cause complex synergetic effects. The numerous consequences of these transformations include: changes in species abundances; canopy loss; reduction in wildlife habitat; increased invasibility; the displacement of native species by invasive species; the formation of novel plant and animal communities; biodiversity loss; alterations to the hydrologic cycle; a decrease in water holding capacity; and temporal disruptions or permanent loss of fundamental ecosystem goods and services like regional water and air quality regulation, carbon sequestration, genetic and pharmaceutical resources, and natural hazard regulation. In addition to having a direct impact on montane ecosystems, severe changes in foreland ecosystems (e.g. deforestation) can affect nearby montane rain forests. However, many questions regarding the potential effects of climate change on ecosystem dynamics and species ranges remain unanswered and require further research.31 Since several key concepts (e.g. of disaster, disturbance, resilience) are applied differently in the discussion on human and environmental insecurity from the way they are understood in ecology, a clarifying interdisciplinary discussion is urgently needed.

31 See Bruijnzeel (2001); Bush, Hanselman and Hooghiemstra (2007); Feeley and Silman (2008); Svenning, Kerr and Rahbek (2009); Boehmer, Niemand, Jacobi, Gerrish and Mueller-Dombois (submitted for publication).

Vulnerability of Tropical Montane Rain Forest Ecosystems Due to Climate Change

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VII

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