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EFFECTS OF HABITAT EXTENT AND FOREST DISTURBANCE ON BIRD

COMMUNITIES IN LOWLAND NEPAL

Bhagawan Raj Dahal

MSc (Ecology)

A thesis submitted for the degree of Doctor of Philosophy at

The University of Queensland in 2015

School of Geography, Planning and Environmental Management

ii

Abstract

Habitat loss and degradation are recognized as the major contributors to species decline and

extinction, and therefore represent a key conservation challenge for biodiversity conservation. Key

to the protection of biodiversity is acquisition of ecological knowledge about how anthropogenic

forest disturbances affect species and how species respond to emergent landscape characteristics.

Furthermore, it is also important to assess how different management approaches and land tenures

influence retention of the biota of particular sites and of landscapes. However, this crucial

ecological knowledge is yet to be obtained for the threatened lowland landscapes of Nepal.

Protected areas cover only a small proportion of forests in lowland Nepal; the majority of forests

outside the protected areas (off-reserve) have been managed by the state government. However, in

recent years, community forestry programs have been increasingly popular as attempts to protect

biodiversity while permitting consumptive forest use by people. It is therefore important to

understand effectiveness of different forest management tenures for avifaunal conservation. I

compared species richness, abundance, diversity and community composition of birds among sites

in community forests, state forests and protected areas. Although sites in protected areas had the

greatest richness of birds, community forests and state managed forests had complementary

assemblages, supporting species not represented in protected areas. Vegetation characteristics such

as large tree density, tree canopy cover and shrub density were also greater in community forests

than in state-managed forests. The findings suggest that the community forestry approach appears to

improve habitat quality compared to state-managed forests, and therefore can be an alternative

tenure type for conservation of off-reserve forests and avifauna in the region.

Subsistence forestry practices such as logging, lopping, and grazing are sources of forest

disturbance in lowland Nepal. Such activities do not reduce forest area, but change habitat

characteristics, potentially affecting biodiversity directly, and through interactions with landscape

characteristics. I examined effects of forest use practices on species richness and abundance of

forest birds, and whether landscape context such as the extent of forest cover moderates disturbance

effects on birds. I found that extraction of forest products reduced forest structural complexity and

altered the avifaunal community of a site. At the site level, large tree density, tree canopy cover and

shrub density were important habitat characteristics, while the extent of forest cover in the

landscape had the greatest influence on richness of birds. The effects of forest disturbance

(livestock grazing and logging) intensity on birds depended on the extent of forest in the

surrounding landscape, with strongest effects in sites with less surrounding forest. Thus, although

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site-level vegetation structure is important, maintenance of forest extent in the landscape is also key

for avifaunal conservation in the region.

Several recent studies have demonstrated that the extent of forest cover and other landscape

characteristics significantly influence bird species richness. However, different foraging guilds are

likely to respond to landscape characteristics in different ways. Therefore, I examined the strength

and magnitude of the relationships between the extent of forest cover and estimated species richness

for overall birds and for each foraging guild separately. I found that landscape-level species

richness of birds positively related to the extent of forest cover in the landscape. However, the

relationship varied among the foraging guilds, with strong effects for foliage-gleaning insectivores

and, to a lesser extent, frugivores, but only weak effects for sallying insectivores. The relationship

between estimated species richness and the extent of forest cover in the landscape was nonlinear,

with species richness decreasing more steeply below about 20-30% forest cover in the landscape.

Importantly, I found that the relationship between richness and forest extent varied among foraging

guilds and with landscape characteristics. Therefore, generalizing relationships between species

richness and the extent of forest across all species could potentially mask important relationships at

the functional level.

The findings of this thesis have important implications for the conservation of avifauna in multiple-

use forest landscapes. Although both site-and landscape-scale forest characteristics have important

influences on bird communities, the extent of forest in the landscape both directly and indirectly

affects persistence of birds in these landscapes. The extent of forest in the landscape can moderate

the effects of subsistence forest use practices on bird assemblages. Therefore, conservation benefits

for avifauna can be maximized by maintaining both site-level habitat structures such as large trees,

and the extent of forest cover at the landscape-level. This can be achieved with appropriate

protection measures through reducing human pressure on forests, and restoration of degraded forest

habitats, particularly those that are heavily exploited such the state-managed forests. Thus,

management approaches such as community forestry for management of off-reserve forests can

potentially complement protected areas and maximize conservation outcomes in the region. Such

measures will improve habitat quality and increase the chance of maintaining viable populations of

the full complement of avifaunal species in the lowland landscape of Nepal.

iv

Declaration by author

This thesis is composed of my original work, and contains no material previously published or

written by another person except where due reference has been made in the text. I have clearly

stated the contribution by others to jointly-authored works that I have included in my thesis.

I have clearly stated the contribution of others to my thesis as a whole, including statistical

assistance, survey design, data analysis, significant technical procedures, professional editorial

advice, and any other original research work used or reported in my thesis. The content of my thesis

is the result of work I have carried out since the commencement of my research higher degree

candidature and does not include a substantial part of work that has been submitted to qualify for

the award of any other degree or diploma in any university or other tertiary institution. I have

clearly stated which parts of my thesis, if any, have been submitted to qualify for another award.

I acknowledge that an electronic copy of my thesis must be lodged with the University Library and,

subject to the policy and procedures of The University of Queensland, the thesis be made available

for research and study in accordance with the Copyright Act 1968 unless a period of embargo has

been approved by the Dean of the Graduate School.

I acknowledge that copyright of all material contained in my thesis resides with the copyright

holder(s) of that material. Where appropriate I have obtained copyright permission from the

copyright holder to reproduce material in this thesis.

v

Publications during candidature

Dahal, B. R., C. A. McAlpine, and M. Maron. 2014. Bird conservation values of off-reserve forests

in lowland Nepal. Forest Ecology and Management 323:28-38.

Dahal, B. R., C. A. McAlpine, and M. Maron. 2015. ‘Impacts of extractive forest uses on bird

assemblages vary with landscape context in lowland Nepal’. Biological Conservation. 186: 167-

175.

Publications included in this thesis

This thesis contains three jointly authored manuscripts (two published and one has submitted for

peer-review). These papers are reproduced in full as chapters of this thesis (2-4). I conducted the

majority (90%) of the work contained within these manuscripts, including: original idea, data

collection, data analysis, interpretation, synthesis, drafting and writing. Co-author contributions are

indicated below the relevant citations.

vi

Dahal, B. R., C. A. McAlpine, and M. Maron. 2014. Bird conservation values of off-reserve forests

in lowland Nepal. Forest Ecology and Management 323:28-38

– incorporated as Chapter 2.

Contributor Statement of contribution

Bhagawan Raj Dahal, (Candidate) Research idea and design (80 %)

Data collection (100 %)

Wrote and edited the paper (85 %)

Statistical analysis of data (100 %)

Clive A. McAlpine Research design (10 %)

Edited paper (5%)

Martine Maron Research idea and design (10 %)


Dahal, B. R., C. A. McAlpine, and M. Maron. 2015. ‘Impacts of extractive forest uses on bird

assemblages vary with landscape context in lowland Nepal’. Biological Conservation 186: 167-175.

–incorporated as Chapter 3.

Contributor Statement of contribution

Bhagawan Raj Dahal, (Candidate) Research idea and design (80 %)

Data collection (100 %)


Statistical analysis of data (100 %)

Clive A. McAlpine Research design (10 %)

Edited paper (5 %)

Martine Maron Research idea and design (10 %)


vii

Contributions by others to the thesis

No contributions by others apart from those listed above

Statement of parts of the thesis submitted to qualify for the award of another degree

None

viii

Acknowledgments

I would like to take this opportunity to express my sincere gratitude to The University of

Queensland for awarding me an International Postgraduate Research Scholarship (IPRS) and

UQ Centennial Scholarship (UQCent) that enabled me to undertake PhD study. Without this

support, my PhD journey would not have commenced at this renowned University. Many

thanks to my enrolling school - the School of Geography, Planning and Environmental

Management for all the help and support I received for my research.

I am very much grateful to my supervisors Martine Maron and Clive McAlpine for their

consistent support throughout this research project. I am particularly truly indebted and

thankful to my primary advisor Martine Maron whose wonderful cooperation, consistent

encouragement and support made this thesis possible. Thank you so much for your advice,

guidance, ideas, time, help and patience to complete this thesis. I am very grateful and

honoured to have you as my principle supervisor.

I thank to the Ministry of Forest and Soil Conservation Nepal for granting research

permission. Many thanks to staffs of Himalayan Nature, Koshi Bird Observatory, Chitwan

National Park, Parsa Wildlife Reserve and Community forest user committees for their

cooperation during the field work.

Many thanks to Suman Acharay, Dinesh Ghimire, Kapil Pokhrel, Tika Sherpa and Arjun

Baral for your assistance in the field. I really enjoyed working with you all. I am thankful to

Jeevan Bikash Samaj, Arjun Chapagain and Tika Ram Giri for their support in logistic

management. I would like to thank to Rudra Kumal, Hira Shrestha and Geeta Ghimire who

generously offered me their house to live in during the field work. Many thanks for your all

cooperation and help. I am very much grateful to Wayne Heydon for his wonderful

hospitality during my stay in Brisbane for thesis correction.

I would like to express my sincere gratitude to Mr Douglas Michael and Mrs Margo Michael

for their care and love during my stay in Brisbane. They kindly provided me a free

accommodation for two years at their home. Thanks to Meredith Gray for your help in

organizing all these things. I am grateful to Dr Roger Jones for his time in brushing up my

writing skills and Jessie Wells for her input in statistical modelling.

ix

I would like to thank to GPEM family particularly Judy, Alan, Suhan for their cooperation in

administrative work. I am grateful to my colleagues for their excellent company here in

University. Thanks to Emma, Kiran, Alvaro, Will, Justus, Andrew, Danielle, Jeremy, Payal,

Rowan, and all LEC friends. Special thanks to Kiran for his help in GIS. I thank to Pradeep

Gyawali, Gokarn Junga Thapa, Biplav Pokhrel, Rajiv Paudel, Uttam Babu Shrestha and

Salma Baral for their support in different stages of my PhD study.

I am grateful to Dr Hem Sagar Baral and Carol Inskipp for their inspiration to work in the

field of bird conservation. Their contribution to ornithological research and conservation in

Nepal is exceptional. I am truly indebted to Hem Dai for his consistent cooperation and

support to build my career in conservation science. I would like to thank to World Pheasant

Association (WPA), Dr Narayan Dhakal, Dr Phillip McGowan and Prof John Carroll for their

support in my research career. I am also thankful to my thesis reviewers’ Dr Peter Garson and

Dr Toby Gardner for their constructive comments, which greatly helped improve this thesis.

Finally, I am thankful to my family for their continuous support and encouragement. I am

very grateful with my mother Gaura Devi, late father, all my brothers and sisters for their

love, care and consistent support for my studies. They taught me the value of education in my

life and society from the beginning of my formal education. Many thanks to my wife Amita

and daughter Abha for their love and understanding which was important to accomplish this

thesis.

x

Key words

Avian community, conservation value, extractive forest disturbances, forest management,

habitat thresholds, interactive effect, landscape context, landscape-level forest cover, multi-

use forest landscapes, tenure arrangement

Australian and New Zealand Standard Research Classifications (ANZSRC)

ANZSRC code: 050104, Landscape Ecology (50%)

ANZSRC code: 050202, Conservation and Biodiversity (30%)

ANZSRC code: 050211, Wildlife and Habitat Management (20%)

Fields of Research (FoR) Classification

FoR code: 0602, Ecology, (50%)

FoR code: 0502, Environmental Science and Management, (30%)

FoR code: 0501, Ecological Applications, (20%)

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Table of contents

Abstract ii

Declaration by author iv

Acknowledgments viii

Table of contents xi

List of Tables xiv

List of Figures xv

List of Plates xvii

List of Appendices xviii

CHAPTER 1: INTRODUCTION AND LITERATURE REVIEW 19

1.1 Background to the problem 20

1.2 Aims and objectives 21

1.3 Theory and concepts underlying this thesis 21

1.3.1 Human modification of landscapes 21

1.3.2 Effects of landscape change on biodiversity 22

1.3.3 Scale and species richness 22

1.3.4 Extent of forest and species richness 23

1.3.5 Conservation of remnant forests 24

1.3.6 Effects of forest use practices on avifauna 26

1.3.7 Forest management in Nepal 27

1.3.8 Threats to lowland forest and birds 29

1.4 Thesis outline 31

CHAPTER 2: BIRD CONSERVATION VALUES OF OFF-FOREST RESERVE

FORESTS IN LOWLAND NEPAL 34

2.1 Abstract 35

2.2 Introduction 36

2.3 Material and methods 39

2.3.1 Study area 39

2.3.2 Study sites 40

2.3.3 Bird surveys 41

2.3.4 Explanatory variables 42

2.3.5 Statistical analyses 43

2.4 Results 45

2.4.1 Species richness, abundance and diversity 45

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2.4.2 Habitat preference groupings 46

2.4.3 Foraging guilds 47

2.4.4 Bird assemblage structure 48

2.4.5 Habitat characteristics 50

2.5 Discussion 51

2.6 Management implications 53

CHAPTER 3: IMPACTS OF EXTRACTIVE FOREST USES ON BIRD

ASSEMBLAGES VARY WITH LANDSCAPE CONTEXT IN LOWLAND NEPAL 55

3.1 Abstract 56

3.2 Introduction 57

3.3 Material and methods 59

3.3.1 Study area 59

3.3.2 Study sites and landscapes 59


3.3.4 Explanatory variables 60

3.3.5 Statistical analysis 62

3.4 Results 65

3.4.1 Vegetation and disturbance 65

3.4.2 Bird communities 65

3.4.3 Effects of site- and landscape-level factors 66

3.4.4. Interactions between forest extent and disturbance intensity 69

3.5 Discussion 72

3.5.1 Effects of forest disturbances on bird communities 72

3.5.2 Moderating effects of landscape context 73


CHAPTER 4: RELATIONSHIPS’ BETWEEN LANDSCAPE-LEVEL SPECIES

RICHNESS AND FOREST EXTENT VARY AMONG BIRD GUILDS 76

4.1 Abstract 77

4.2 Introduction 78

4.3 Material and Methods 80

4.3.1 Study area 80

4.3.2 Study design 81


4.3.4 Landscape variables 81

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4.3.5 Data analysis 82

4.4 Results 83

4.4.1 Relationships between landscape-level species richness and forest cover 83

4.4.2 Relative importance of landscape variables 84

4.4.3 Interactions between landscape characteristics 86

4.5 Discussion 87

4.5.1 Species richness and forest cover 87

4.5.2 Relative importance of landscape characteristics 89

4.5.3 Interactions between landscape characteristics 89


CHAPTER 5: SYNTHESIS AND RECOMMENDATIONS 91

5.1 Overview 92

5.2 Off- reserve forests provide complementary habitats for bird conservation 93

5.3 Forest use practices can have detrimental effects on vegetation and associated birds 94

5.4 Effects of forest-use practices on bird assemblages vary with the landscape context 96

5.5 Relationships between species richness and forest extent vary among bird guilds 98


5.7 Limitations of my research 107

5.8 Future research 107

5.9 Conclusion 108

REFERENCES 110

APPENDICES 133

Appendix A 133

Appendix B 138

Appendix C 156

xiv

List of Tables

Table 2.1 Summary of explanatory variables used to assess influence of stand and landscape-

scale variables on bird communities 43

Table 2.2 Mean species richness, average abundance per survey, Shannon diversity index

and Pielou evenness index of all birds in three forest management tenures 46


and Pielou evenness index (±SE) of birds of different habitat groups in community managed

forests, state managed forests and protected areas 46


and Pielou evenness index of different foraging guilds of birds in three forest management

tenures 47

Table 2.5 Results of PERMANOVA and PERMDISP pairwise comparison tests of bird

community composition in response to different forest management tenures 48

Table 3.1 Summary of explanatory variables used to assess the influence of stand and

landscape-scale variables on bird communities 62

Table 5.1 Summary of study themes and key findings of this thesis 103

xv

List of Figures

Figure 2.1 The location of the study area and bird survey sites: (a) Chitwan forests (b) Parsa

forests and (c) Eastern Terai forests 41

Figure 2.2 Bird assemblage NMDS (2D stress = 0.24) plot fitted with environmental vectors.

Vectors represent the mean direction and strength of correlation of different environmental

variables (Fextent = Forest extent, Scv = Shrub cover, Sden = Shrub density, Mtdn = Mature

tree density, Tcv = Tree canopy cover). Vectors with significance P < 0.05 only are shown49

Figure 2.3 Distribution of tree size classes (±S.E.) in sites in community forests, state forests

and protected areas 50

Figure 3.1 Location of the three study regions in Nepal: a) Chitwan forest, b) Parsa-Bara

forest and c) Eastern forest 59

Figure 3.2 Mean (± S.E.) species richness (dark bars) and average abundance (light bars) of

(a) all birds (b) bark-gleaning insectivores and (c) foliage-gleaning insectivores 66

Figure 3.3 Model averaged coefficient estimates (± S.E.) across the 95% confidence set of

models for all explanatory variables 67

Figure 3.4 Summed Akaike weights (∑ωі) of final subset of the explanatory variables for (a)

overall species richness, (b) average abundance, (c) bark-gleaner species richness (d) bark-

gleaner abundance (e) foliage-gleaner species richness, and (f) foliage-gleaner abundance69

Figure 3.5 Relationship between overall bird species richness (a),overall bird abundance (b),

bark-gleaning abundance (c), and foliage-gleaning abundance and percent of forest cover in

2.5 km radius of survey site in landscape (heavily disturbed sites (filled circles and heavy

dashed line) and lightly disturbed sites (open circles and fine dashed line). For the purposes

of displaying the interaction effect, sites were divided into heavily and lightly disturbed based

on the value of the disturbance index (from three disturbance types) 71

Figure 4.1 The study landscapes in lowland Nepal (grey shading indicates forest cover) and

histogram showing the different extent of forest cover in landscape 80

Figure 4.2 Models of the relationship between estimated species richness and forest extent in

the landscape. AIC: Akaike information criterion for each model 84


models for all explanatory variables: (a) overall bird, (b) foliage-gleaning insectivore, (c)

frugivore and (d) sallying insectivore 85

Figure 4.4 Summed Akaike weight (∑ωі) from model averaging of the environmental

variables for landscape-level richness of (a) all species, (b) foliage-gleaning insectivores, (c)

Frugivores, and (d) sallying insectivores 86

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Figure 4.5 Relationship between overall estimated richness and percent of forest cover (a) in

more- disturbed landscapes (filled circles and full line) and less-disturbed landscapes (open

circles and heavy dashed line) and (b) higher-rainfall landscapes (filled circles and full line)

and lower-rainfall landscapes (open circles and heavy dashed line) 87

xvii

List of Plates

Plate 1: An example of resource extraction in lowland Terai forests (anti-clockwise from top

left: logging for timber, removal of standing tree for firewood, cattle grazing, and tree canopy

collection for fodder). 19

Plate 2: Regeneration of forest after implementing a community forestry program in

Barandabhar Community forest, lowland Nepal 34

Plate 3: Creekbed in large patch of forest adjacent to an agricultural area in Parsa Wildlife

Reserve, lowland Nepal 55

Plate 4: A patch of native forest in Chitwan National Park in lowland Nepal 76

Plate 5: Agricultural intensification in adjacent to a protected area, lowland Nepal 91

xviii

List of Appendices

Appendix A 133

Table A.1 Total number of individuals and their respective food guilds and habitat groups, and

relative percentage of species observation across sites in lowland tropical landscapes: CF =

Community forest, SF = State forest, PA = Protected area 133

Table A.2 Top ten contributing species to dissimilarities between management tenures using

SIMPER analysis based on Bray-Curtis dissimilarity 137

Appendix B 138

Table B.1 Summary statistics of habitat characteristics (± SE) across sites in grazing, logging and

lopping disturbances in lowland tropical landscape 138

Table B.2 Results of analysis of variance (ANOVA) for the vegetation characteristics across the

different disturbance types 139

Table B.3 Results of analysis of variance (ANOVA) for overall species richness, total average

abundance, species richness and abundance of bark-gleaning and foliage-gleaning insectivore 142

Table B.4 Model averaged coefficient and sum of Akaike weights (∑ωі) (fixed effects) for each

variable based on AIC. Significant estimates are in bold 144

Table B.5 AIC, ∆value and Akaike weights (ωі) for models of overall species richness, total

average abundance, species richness and abundance of bark-gleaning and foliage-gleaning

insectivore 146

Table B.6 AIC, ∆value and Akaike weights (ωі) for interaction models of overall species

richness, total average abundance, species richness and abundance of bark-gleaning and foliage-

gleaning insectivore 152

Table B.7 Mean values of different disturbance types between sites classified as lightly and

heavily disturbed in lowland Terai forests 154

Table B.8 Correlation matrix of explanatory variables measured. Coefficients in bold shows pairs

highly correlated variables 155

Appendix C

Table C.1 AIC, Δ value and Akaike weights (ωі) for models of overall estimated species,

frugivore, foliage-gleaning insectivore and sallying insectivore 156

Table C.2 Model averaged coefficients across the 95% confidence set of models for all

explanatory variables 160

Figure C.1 Species accumulation curves of all 28 studied landscapes based on Chao2/ICE

estimated richness 161

19

CHAPTER 1

INTRODUCTION AND LITERATURE REVIEW

Plate 1: An example of resource extraction in lowland Terai forests (anti-clockwise from top

left: logging for timber, removal of standing tree for firewood, cattle grazing, and tree canopy

collection for fodder).

20

1.1 Background to the problem

In the face of growing demand for food and resources for an ever-increasing human

population, managing forests for biodiversity conservation is a challenging task. Rapidly

growing demands for food and resources have often been met at the cost of forests and

woodlands across the globe (FAO 2005, Gibbs et al. 2010). Approximately 80% of the

world’s original forest cover has been cleared, fragmented and degraded (World Resources

Institute 1997). Anthropogenic habitat degradation is greater in areas of high population

density and poverty, particularly in developing countries (Laurance 2010) where a significant

proportion of the population lives near the forests (Hegde and Enters 2000, Millennium

Ecosystem Assessment 2005).

Such widespread anthropogenic habitat loss and degradation has significantly changed the

pattern and function of the landscape and consequently altered the habitat characteristics in

the landscape (Christensen et al. 2009). Loss of habitat is the key threat to biodiversity and is

the major contributor to global species extinction (Fahrig 2001, Pereira et al. 2004).

Landscape characteristics such as the extent of forest cover in the landscape and forest

disturbance type and intensity may be important determinants of species richness in the

landscape. Therefore, understanding the response of species to the emergent properties of

landscape is crucial for their effective conservation.

Forest loss and degradation as a result of anthropogenic pressure pose the principal threats to

forest-dependent birds (Schmiegelow and Mönkkönen 2002). However, effects of

disturbance on forest birds may vary among species, depending on their sensitivity to habitat

disturbances. Anthropogenic disturbance often has the most adverse effects on forest habitat

specialists (Sodhi and Ehrlich 2010) and species of birds with specialised diets (Cleary et al.

2007, Greve et al. 2011). Forest loss and degradation are often associated with

disproportional loss of critical habitat components such as large trees, fallen logs and woody

debris (Inskipp et al. 2013). Persistence of many species depends on availability of such

critical habitat features in the landscape (Kumar et al. 2011). Thus, as habitat loss changes the

distribution of habitat resources in the landscape (Andren 1994, Coulson and Tchakerian

2010), understanding how bird communities respond to habitat loss and associated forest

disturbance is crucial for making sound conservation management decisions.

21

1.2 Aims and objectives

The overall aim of this study was to investigate the effects of habitat characteristics, forest

extent and forest use disturbances on forest bird assemblages in human-dominated landscapes

in lowland Nepal. A secondary aim was to explore conservation values of off-reserve forests

for bird communities. In this thesis, I examined effects of anthropogenic forest use practice

such as logging, grazing and lopping on vegetation structure and associated bird

communities, and the role of landscape context in moderating the effects of disturbance on

birds. Further, I investigated whether the effects of landscape-level forest extent on

landscape-level species richness of birds are universal across foraging guilds, or if a

particular guild contributes disproportionately to observed patterns between species richness

and forest cover in the landscape. The specific research objectives investigated in thesis are

to:

Determine conservation values of off-reserve forests for bird communities in

comparison to those of protected areas in lowland Nepal.

Investigate effects of site and landscape characteristics on bird species richness and

abundance, and explore interactions between disturbance intensity and landscape

context in their effects on bird assemblages.

Investigate whether the effects of remnant forest extent in a landscape on the species

richness of birds in that landscape are consistent across foraging guilds, or if a

particular guild contributes disproportionately to observed relationship between

species richness and forest cover in the landscape.

1.3 Theory and concepts underlying this thesis

1.3.1 Human modification of landscapes

Native forests are invaluable habitats for the conservation of globally significant biodiversity.

These forests encompass important characteristics of natural ecosystems and support

important terrestrial ecosystems (World Commission on Forests and Sustainable

Development 1999). However, despite the global biological importance of forest ecosystems,

anthropogenic activities such as forestry, subsistence and commercial agriculture, and urban

development have caused significant loss and degradation of forest landscapes. These

processes of habitat loss and modification reduce the total area of suitable habitat, increase

the number of patches and increase the isolation of patches from one another (Saunders et al.

1991, Fahrig 2003, Fischer and Lindenmayer 2007). Such fragmentation contributes directly

22

and indirectly to changes in the structure and function of ecological mosaics and their

constituent biota (Franklin et al. 2002, Morrison et al. 2006, Fischer and Lindenmayer 2007).

1.3.2 Effects of landscape change on biodiversity

Landscape change alters ecological processes and threatens the survival of species and

populations in several ways. For example, landscape modification involves the removal of

critical habitat components and alters habitat structure and composition, thereby affecting

long-term persistence of faunal populations (Flynn et al. 2009, Halley and Iwasa 2011,

Lindenmayer and Cunningham 2013). Anthropogenic modification of habitat can

disproportionately affect particular vegeation types, which reduces native plant species

richness, in turn reducing habitat diversity in the landscape (Fischer and Lindenmayer 2007).

The reduction in habitat extent and diversity significantly alters the faunal assemblages that

landscapes can support (Fahrig 2003, Bennett et al. 2006, Morrison et al. 2006). As a

consequence, many species of fauna are severely threatened and some are on the verge of

local or even global extinction (Jetz et al. 2008, IUCN 2009). Therefore, understanding the

relationships between patterns of species occurrence and landscape properties in modified

landscapes is critically important for developing effective conservation strategies (Radford

and Bennett 2007, Gardner et al. 2009, Balmford et al. 2012).

Landscape modification not only reduces the type and amount of habitat but also changes the

configuration of remaining habitat patches (van den Berg et al. 2001, Bennett et al. 2006).

Such changes affect resource availability for a wide range of faunal species. Because isolated

habitat patches may not contain all of a species’ resource requirements (Bennett and Saunders

2010), many species that require different habitat attributes for different purposes (such as

foraging and breeding) must travel between patches to acquire resources (Dunning et al.

1992, Law and Dickman 1998). Increasing isloation of habitat attributes could potentially

impede this movement of fauna, affecting the persistence of faunal communties in human-

modified landscapes (Hinsley 2000, Fahrig 2003, Ewers and Didham 2006). Thus, the

protection of a complementary habitat within close proximity can be important for

maintaining viable populations of many species (Dunning et al. 1992).

1.3.3 Scale and species richness

Understanding how the landscape affects the species that can occupy a given site does not

necessarily reveal the patterns of species diversity expected across different spatial scales.

23

The number of species in an individual site/or patch (alpha diversity) is only part of what

generates regional species richness (Whittaker 1972). As the number of sites sampled across

an area increases, variation in species composition (beta diversity) across sites/or habitat

types in a landscape is also likely to increase (Whittaker 1977, Koleff et al. 2003, Kessler et

al. 2009). Because a landscape is composed of a mosaic of different habitat types, the

contribution of landscape heterogeneity to species diversity can be important. Thus species

richness at the site level might not necessarily represent the pattern of species diversity at

larger scales (Kettle and Koh 2014). The rate of change of species composition among sites

or habitat patches (also called species turnover) contributes significantly to the regional

diversity (γ-diversity) Tscharntke et al. 2005). Thus, a multi-scale approach to understanding

the relationships between landscape properties and species richness is important for

conservation planning for regional diversity (Bennett et al. 2006, Radford and Bennett 2007).

1.3.4 Extent of forest and species richness

There is likely to be a minimum extent of habitat in the landscape required to support the

persistence of full species diversity (Andren 1994, Fahrig 2001). Landscapes with more forest

cover support a larger species pool (Hu et al. 2012, Taylor et al. 2012) through both sampling

effects (Wiens 1992, Whittaker and Fernández-Palacios 2007) and because a greater forest

extent offers habitat diversity (Radford et al. 2005, Maron et al. 2012). Several empirical

studies have shown that sites in landscapes with more forest support higher densities of

reptiles (McAlpine et al. 2015), greater species richness and abundance of birds (Villard et al.

1999, Mortelliti et al. 2010, Martensen et al. 2012, Taylor et al. 2012), and greater richness of

small mammals (McAlpine et al. 2006, Estavillo et al. 2013). These findings suggest that a

wide range of forest-dependent faunal communities can persist even in modified landscapes,

as long as adequate forest cover is retained (Radford et al. 2005, Ochoa‐Quintero et al. 2015).

Moreover, the extent of forest in the landscape not only affects faunal communities directly,

but may also have a potential role in moderating the effects of anthropogenic disturbance. For

example, the intensity of disturbance (livestock grazing, logging and extraction of firewood

and fodder) may interact with the amount of forest in the landscape to drive landscape-level

species richness. Thus, understanding of interactive effects between forest extent and

disturbance intensity is important for biodiversity conservation in multiple-use forest

landscapes.

24

The understanding of the nature and magnitude of the relationship between species richness

and forest extent is important for conservation of faunal communities in human-modified

landscapes. As habitat extent decreases, so do ecological functions and associated

biodiversity in the landscape (Swift and Hannon 2010). Species richness may respond non-

linearly to habitat loss, and decline sharply when the extent of forest exceeds certain

threshold levels (Maron et al. 2012, Ochoa‐Quintero et al. 2015). For example, in the eastern

Australia, Maron et al. (2012) found a non-linear relationship between woodland birds and

landscape-level forest extent. A sharp decline in species richness occurred when forest extent

fell below 40% in the landscape. In southern Australia, Radford et al. (2005) found that the

species richness of woodland birds declined sharply in landscapes with less than 10% of

habitat cover. Similarly, Martensen et al. (2012) found evidence of thresholds for understory

birds at 30-50% of forest cover in Atlantic Forest landscapes, below which species richness

declined more steeply. A similar trend in the relationship between species richness and

landscape-level forest cover was reported by Ochoa‐Quintero et al. (2015) in the Brazilian

Amazon forest. They found evidence of a threshold at 30-40% of forest cover, below which

species richness of both mammals and birds declined aburptly.

Although these relationships have typically been considered as general phenomena, the

response of species to habitat extent may vary depending on land-use or soil type (Maron et

al. 2012), species sensitivity to forest cover change (Martensen et al. 2012), and diversity of

food resoures available in the landscape. For example, species richness of frugivorous birds

in particular landscapes is likely to be driven by the diversity and availability of fruit

resources (Kissling et al. 2007), while, although at site level, abundance of insectivorous

birds is affected by the abundance of prey resources such as arthropods in forest patches

(Capinera 2011). Therefore, it is likely that the relationship between the extent of forest cover

and landscape-level species richness of birds varies among different foraging groups,

although this has not hitherto been explored.

1.3.5 Conservation of remnant forests

In response to continued loss of habitat and biodiversity, protected areas have been

established around the globe. A large body of research has shown that protected areas have

been important in maintaining larger extent of forests and their constituent biota (e.g.

Rodrigues et al. 2004, Jenkins and Joppa 2009). Protected areas are often targeted at

protecting species of high conservation value (e.g. Lee et al. 2007), and therefore reduce the

25

risk of species extinction (Brooks et al. 2004, Jenkins and Joppa 2009, Joppa and Pfaff 2011).

However, the protected area network also suffers from lack of representation of certain

species and habitats (Tewksbury et al. 2002, Hoekstra et al. 2005, Barr et al. 2011, Butchart

et al. 2012). Protected areas may not necessarily contain the full range of habitats required to

support all species across the landscape. For example, approximately 43% of Asian Important

Bird Areas are unprotected (BirdLife International 2004). In addition to this, protected areas

have not been able to prevent habitat loss and associated faunal diversity in many parts of the

world (e.g. Clark et al. 2013). These findings indicate that protected areas alone are

inadequate for conservation.

In response to the inadequacy of the current protected area network, there has been increasing

interest in conservation of off-reserve forests for biodiversity conservation (Bhagwat et al.

2005, Ellis and Porter-Bolland 2008, Persha et al. 2010). As more than 80% of the world

terrestrial habitats are outside protected areas (Barr et al. 2011, Bastian et al. 2012), these off-

reserve forests can be important complementary habitats to the existing protected area

network. For example, off-reserve forests often include patches of remnant native forest of

different types to those within protected areas, and regrowth (secondary) forests (Dudley and

Phillips 2006). These provide habitat resources for many species that are poorly represented

within protected areas (Lindenmayer and Burgman 2005, Mathur and Singh 2008, Cox and

Underwood 2011). Thus, appropriate policies for conservation management of off-reserve

forests could help maintain a diversity of habitat resources for a range of species across the

landscape (Lindenmayer 2009).

Community forestry has emerged as an alternative forest management strategy to commercial

forestry or complete reservation, particularly in developing countries (Ellis and Porter-

Bolland 2008, Porter-Bolland et al. 2012, Kitamura and Clapp 2013). Community forests

offer legal rights to local community and establish a sense of ownership in forest resource

management (Hayes and Ostrom 2005, Singh and Chapagain 2006). In developing countries,

about 27% of the total forest area is either community-managed or owned (Molnar et al.

2011). Studies have shown that community forest initiatives have lower deforestation and

improve forest condition in the landscapes due to restoration of degraded forests (Nagendra

and Gokhale 2008, Porter-Bolland et al. 2012, Lambrick et al. 2014). They are also important

sources of livelihoods for forest-dependent human communities (Lawrence et al. 2006,

Charnley and Poe 2007). For example, forest resources such as fuel wood, fodder, and timber

26

are the primary source of livelihood for local communities (Kumar and Shahabuddin 2005).

Thus, managing forests outside protected areas through community participation can be

beneficial to both the biodiversity and local community.

1.3.6 Effects of forest use practices on avifauna

Anthropogenic forest use practices such as timber and firewood extraction and livestock

grazing are key drivers of forest loss and degradation (Millennium Ecosystem Assessment

2005). Yet in many developing countries, these practices constitute the most important

aspects of a subsistence livelihood (Chao 2012, World Bank 2012). For example, globally,

particularly in developing countries, between one-third and one-half of people rely on wood

and other biomass fuels for energy (Bailis et al. 2012). Such widespread exploitation of forest

severely alters habitat characteristics, with changes affecting trees, shrubs and associated

structures such as hollows (Chettri et al. 2002, Kumar et al. 2011). This structural

simplification of habitat can result in significant declines of fauna (Lindenmayer et al. 2012,

Lee and Carroll 2014).

In developing countries, forest resource extraction such as timber and firewood, lopping of

tree branches for fodder and firewood, and cattle grazing are major forms of disturbance in

remnant forests (Chettri et al. 2005, Shahabuddin and Kumar 2007, Thapa and Chapman

2010). Excessive extraction of these resources can lead to significant habitat loss and

degradation. This may disproportionally reduce the key suitable habitats for faunal

communities in the landscape. Harvesting of resources for human use potentially affects

vegetation structure and composition (Sagar and Singh 2004, Kumar and Shahabuddin 2005,

Gil-Tena et al. 2008). The removal of standing trees changes stand structure (Marzluff et al.

2000, Moktan et al. 2009) and alters the abundance and density of trees (Lindenmayer 2009).

For example, Chettri et al. (2005) found that large tree density, tree basal area and woody

biomass significantly lower in sites in heavily degraded forests in Sikkim, India. Therefore,

the simplification of vegetation structure as a result of forest resource extraction may pose a

serious threat to the persistence of avifauna.

Studies have shown that structural complexity of habitat strongly influences bird

communities (Sekercioglu et al. 2002, Maron and Kennedy 2007), especially forest interior

specialists and insectivores (Lee et al. 2007, Greve et al. 2011). These groups of birds use

large and dead or rotting trees as foraging and nesting habitats (Shahabuddin and Kumar

27

2007). For example, species richness and abundance of bark-gleaning insectivores can be

reduced by the reduced density of larger trees (Adams and Morrison 1993), while removal of

tree branches and canopy cover may have caused lower richness and abundance of foliage-

gleaning insectivores (Leal et al. 2013). However, most studies of the effects of extractive

forest practices have focused on the effects of large-scale logging on avifaunal communities

(Sekercioglu et al. 2002, Politi et al. 2012, Thinh et al. 2012). The role of repeated extraction

of smaller amounts of woody biomass for timber and firewood in affecting birds and bird

functional groups is less well-understood.

Livestock grazing is another common and detrimental form of anthropogenic forest

disturbance. The disturbance caused by grazing significantly affects forest ecosystem

processes and reduces plant diversity (Ludwig et al. 2000, Mayer et al. 2006). Changes in

plant species composition and foliage density may adversely affect the resource availability

to birds (Alexander et al. 2008), and thus affect their richness and abundance. Grazing often

perturb bird species assemblages, with effects varying with grazing intensity (Eyre et al.

2009, Whitehorne et al. 2011). Grazing changes understory composition through reduction or

removal of shrubs and herbaceous perennials (Buffum et al. 2009); therefore species that

depend on understory vegetation for foraging and nesting may be most negatively affected by

livestock grazing (Martin and McIntyre et al. 2007). Grazing is a common practice in

community-managed forests, and so understanding its effects in those forests is important for

recommendations around forest management prescriptions and tenure.

1.3.7 Forest management in Nepal

In Nepal, forests cover nearly 36% of the land area and include many important ecosystems

(BCN and DNPWC 2012). The country’s forests are designated as national or private forests,

with five sub-categories on the basis of management regimes: state managed forests (also

called national forest), community forests, protected forests, leasehold forests and religious

forests. However, most forest is managed under three of these management regimes: state

managed forests, community managed forests, and protected forests. An estimated total of

2.45 million hectares remains off reserve (Shyamsundar and Ghate 2011). Of the forest

outside the protected area, 1.2 million hectares are currently managed by community forest

user groups (DoF 2010, BCN and DNPWC 2012), while ~1 million hectares of forests is

under the national government management (Shyamsundar and Ghate 2011).

28

Although Nepal has designated about one quarter of its land mass as protected areas, the

majority of its protected land is concentrated in the high Himalayas and throughout the less-

productive landscapes (HMG/MFSC 2002, Heinen and Shrestha 2006). For example, about

48% of high Himalayas are protected, whereas only 0.8% of the Mild Hills and 5.5% of the

Terai zone are protected (Shrestha et al. 2010). Thus, only a small proportion of Nepal’s most

productive lowland forest is represented in the current protected area network. Despite the

small proportion in size, lowland protected areas are important in conservation of country’s

last remaining natural habitats and their constituent biota. They are critical for the

conservation of many endangered and globally significant mammalian species (e.g. Greater

One-horned Rhinoceros Rhinoceros unicornis, Royal Bengal Tiger Panthera tigris) as well

globally threatened avifaunal species such as Great Hornbill Buceros bicornis and Kashmir

Flycatcher Ficedula suvrubra.

However, protected areas do not represent large tracts of primary forests in the lowland of

Nepal. Several important habitats for birds are still located outside the protected areas and

have received little conservation attention despite their great bird conservation values. For

example, in recent years, community-managed forests have become increasingly recognised

for their conservation values for biodiversity in Nepal. Nepal offers some of the best

examples of community-based forest management in the world (Pokharel et al. 2007,

Nagendra and Gokhale 2008). The transfer of use and management rights to locally-formed

community groups has improved effectiveness in management of resources (Ojha et al. 2009,

Nagendra 2007, Shyamsundar and Ghate 2011). Since the implemenation of the Forests Act

of 1993 and the Forest Rules and Regulations of 1995, more than one quarter of country’s

national forest has become managed by community forest user groups (CFUGs) (MOF 2012).

Local communities perceive community forest regime as a secure tenure generating

sustainable sources of income from forest products (Gautam 2007, Kanel and Dahal 2008).

For example, the annual income of community forests was double the total revenue of the

Department of Forests for the fiscal year of 2002 (Kanel et al. 2003). This ownership in

resource management and utilization is one of the key drivers of community forestry success.

Studies have shown that rates of habitat loss and degradation are reduced in community

managed areas compared with state managed forests (Nagendra 2007, Kanel and Dahal

2008). In addition to improving forest health, community forestry has played a crucial role in

boosting local livelihoods (Sapkota et al. 2009).

29

Although both protected areas and off-reserve forests comprise a Shorea robusta-dominated

forest mosaic, habitat characteristics among these differently-managed forest types vary. This

is primarily due to different forest management practices. For example, the strict reserve

systems are predominantly relatively undisturbed primary forests (Nagendra et al. 2008). Off-

reserve forests support both primary forest and regenerating forest (Kanel and Dahal 2008).

Such structural and compositional differences in habitat characteristics among tenure types

may offer habitat for different assemblages of faunal species in the landscapes. The old

growth primary forest in protected areas, for example, may support an assemblage of forest-

sensitive species and forest specialists (Inskipp and Inskipp 1991, BCN and DNPWC 2011)

while the silviculturally-treated and successional forest habitat within community forests may

be expected to support a range of bird species including forest specialists (BCN and DNPWC

2011), open country species (Roman 2001) and mixed-habitat species (Anand et al. 1997,

Blake and Loiselle 2001). Thus, the effective conservation of these forests may play an

important complementary role in safeguarding regional diversity of birds.

However, the response of faunal communities to community management of forests is poorly

known in the region. In Nepal, the majority of faunal studies have been within the protected

areas (Inskipp and Inskipp 2001, Baral et al. 2012, Kapfer et al 2012) and are largely

focussed at species level (Poudyal et al 2008, Dahal et al 2009, Baral 2012). More recently,

there have been some studies in off-reserve forests, in particular on forest regeneration and

plant communities (Gautam et al. 2002, Timilsina et al. 2007, Sapkota et al. 2009, Thapa

2010). Studies reported a significant increase in plant diversity and decreased deforestation

within the community-managed forests following the tenure change (Nagendra et al. 2008,

Sapkota et al. 2009). Yet, to my knowledge, the effect of this vegetation change on fauna is

poorly understood in Nepal. There is therefore a need to quantify the contribution of these

areas to biodiversity conservation, and in particular, the extent to which the biota they support

is complementary to that within formal protected areas.

1.3.8 Threats to lowland forest and birds

Forests in lowland landscapes are generally degraded due to anthropogenic forest extraction

activities (Webb and Sah 2003, DoF 2009). Before 1950, the region supported continuous

dense tropical forest. With the eradication of malaria in the early 1950s, large tracts of the

highly productive lowland forests were converted to agriculture (Hrabovszky and Miyan

1987). Consequently, most of the forest was destroyed and the remaining forest areas were

30

subjected to intense human exploitation. Nearly half of Nepal’s population now lives in the

17% of the country that is lowland (Central Bureau of Statistics 2011). Therefore the majority

of forests in this region are exploited mainly for subsistence livelihood.

Forest use practices such as logging, grazing and the excessive extraction of fuel wood and

fodder are the major activities in these forests (Webb and Sah 2003). Similarly, extraction of

litter, grass and removal of dead and logged trees are other examples of extraction of forest-

based products in the region. Livestock farming is a major profession and is an essential

component of the rural farming system in Nepal. For example, about 41% of the country’s

cattle, 39% of buffaloes and about 37% of goats are all farmed in the lowlands (MoAC 2004),

that has contributed to the rapid depletion of forest land (HMG/MFSC 2002).

Despite the prevailing threats, these forests can be important habitats for a wide range of

avifaunal diversity. Recent assessment study by BirdLife Nepal and Wildlife Department-

Nepal for the region reported that these forests harbour more than 50% of nationally

threatened species (BCN and DNPWC 2010), over three quarters of the country’s breeding

species and 67% of wintering species (Inskipp 1989). Of the total species recorded in Nepal,

over 70% of the forest bird species are recorded in tropical and subtropical forests of the

lowland (Inskipp and Inskipp 1991, HMG/MFSC 2002). Furthermore, 53% of the country’s

nationally threatened bird species inhabit the lowland tropical forest (Bird Life International

2012), of which 56% are only found in the lowland forest (BCN and DNPWC 2010).

Ecological information about the effects of anthropogenic habitat disturbances on avifauna,

particularly at multiple scales, is poor in lowland Nepal. This knowledge gap is a challenge

when devising strategies for avifaunal conservation for the region. To achieve effective

conservation of forests and associated biodiversity in these areas, empirical information about

how the remaining extent of forest and other landscape attributes affects to species richness,

abundance and community structure of birds across landscapes is required. Thus,

understanding the drivers of species richness and assemblages at the landscape-level is

critical for appropriate conservation of birds in the region.

31

1.4 Thesis outline

As I discussed in the previous sections, the increasing extent and intensity of land-use is a

significant cause of biodiversity declines world-wide. In developing countries, forest-use

practices in the form of logging, grazing, and lopping are the major forms of habitat

disturbances in the remnant forests. However, effects of subsistence forest disturbances on

bird assemblages and the role of landscape context in moderating effects of disturbance on

birds are poorly understand in lowland Terai forests of Nepal. The effects of properties of

landscapes on forest bird assemblages in this region with its particular disturbance regimes

are not known. As identified in the previous section, the current understanding of patterns of

species richness and forest extent relationships at the functional levels of birds is limited.

This thesis aims to address these knowledge gaps. This section describes how the remainder

of this thesis is structured to answer the three research questions related to the overall aim of

this study. The next three chapters (2-4) are presented as a set of stand-alone journal articles,

each of which is either published or in review. Each chapter addresses a specific research

question that relates to the broader research aim of this thesis.

Chapter 1: Introduction and literature review

This chapter includes the thesis overview and general introduction and overview of the

problem, with aims and specific objectives of the study. I summarise the literature from a

broad range of topics relevant to this thesis. I concluded this chapter with a brief overview of

threats to birds, knowledge gaps and the need to acquire ecological information for

conservation of avifauna in the region.

Chapter 2: Bird conservation values of off-reserve forests in lowland Nepal

In this chapter, I evaluated whether off-reserve forest supports bird assemblages

complementary to those of protected area. I compared species richness, abundance, diversity

and community composition of birds among three management tenures. The findings from

this chapter support the hypothesis that off-reserve forests in particular community forests

have complementary bird assemblages to that of protected areas. This chapter has been

published in Forest Ecology and Management in 2014 (Volume 323).

Chapter 3: Impacts of extractive forest uses on bird communities vary with landscape

context in lowland Nepal

32

In Chapter 2, I used site-scale habitat characteristics to evaluate the conservation values of

off-reserve forests and protected areas for bird communities. As I highlighted in previous

sections, lowland forests are subjected to subsistence forestry practices, which, in turn, may

significantly change the habitat structure and associated avifauna. Therefore in Chapter 3, I

investigated effects of forest disturbances on bird communities, and whether landscape

context—specifically, forest extent within a 500 m and a 2500 m radius of survey sites—can

moderate these effects on forest bird assemblages. I then examined the relative importance of

site-and landscape-scale forest habitat characteristics on bird species richness and abundance.

This study demonstrated that structural features of forest stands such as canopy cover, tree

sizes and shrub density were important influences on avifaunal assemblages. I found that

while forest use practices significantly affected the avifaunal community of sites, the

intensity of these effects can be moderated by maintaining the forest cover in the landscape.

This chapter has been accepted for publication in Biological Conservation (accepted 11

February, 2015).

Chapter 4: Relationships between landscape-level bird species richness and forest extent

vary among guilds.

In Chapter 3, I found strong positive relationships between forest bird assemblages and both

site and landscape characteristics. I also found that extent of forest within a 500 m and a 2500

m radius of survey sites was important to moderate impacts of forest disturbance on birds at

the site-level. However in Chapter 4, I extended my research approach beyond the

site/landscape context design. Here, I investigated the relative importance of forest extent and

other landscape characteristics on estimated richness at the landscape-level of all birds,

frugivores, foliage-gleaners and sallying insectivores. In this Chapter, I found that although

the relationship between species richness and forest extent was strong, the strength and

magnitude of the relationship varied considerably among foraging guilds. The relationship

between species richness and extent of forest cover in the landscape was non-linear, with

landscape-level species richness declining more steeply when forest cover in the landscape

fell below 20-30%. This chapter will be submitted to Diversity and Distributions in March

2015.

Chapter 5: Synthesis and conclusions

This chapter synthesises the findings of the previous chapters and explores how each

contributes to our understandings of multi-scale effects of habitat characteristics and forest

33

disturbances on bird assemblages in human dominated forested landscapes. I have also

highlighted future research priorities for the region, and for relevant areas of the field of

landscape ecology.

34

CHAPTER 2

BIRD CONSERVATION VALUES OF OFF-RESERVE FORESTS IN LOWLAND

NEPAL

Plate 2: Regeneration of forest after implementing a community forestry program in

Barandabhar Community forest, lowland Nepal.

Published as: Dahal, B. R., C. A. McAlpine, and M. Maron. 2014. Bird conservation values

of off-reserve forests in lowland Nepal. Forest Ecology and Management 323:28-38.

35

2.1 Abstract

Although protected areas are central to global biodiversity conservation, off-reserve forests

are increasingly recognized as potentially important for the long term conservation of biota,

particularly in less-developed countries where communities rely directly on resources from

natural areas. We assessed the conservation value of differently managed forests for birds in

lowland tropical forests of Nepal. In particular, we explored whether their conservation value

was additional or complementary to those of formal protected areas. Using data collected

from 112 sites in protected areas (n = 31), state managed forests (n = 37) and community

managed forests (n = 44), we assessed how bird species richness, abundance, diversity and

community composition varied among tenures. Although sites in protected areas had the

greatest species diversity, community managed forests supported a complementary

assemblage. Of 124 species recorded, only 45% were common to all management tenures.

Overall, the distinctiveness and richness of species in sites in forests outside of protected

areas contributed substantially to regional avifaunal diversity. These results highlight the

potentially critical role of appropriately managed community forests. The maintenance of

diverse bird assemblages in forest regions depends on complementary management of forests

both outside and inside the established protected areas.

Key word: Bird community; forest management; conservation value; tenure arrangement;

community participation

36

2.2 Introduction

In the face of growing pressure on global biological diversity, the protected area network is

increasingly important for biodiversity conservation worldwide (Joppa et al. 2008, Jenkins

and Joppa 2009). Protected areas have been important in maintaining extensive primary

forests and protecting species of high conservation value (Lee et al. 2007). However, there

are concerns regarding the adequacy of protected areas in terms of representation of species

and their habitats (Rodrigues et al. 2004). Recent work has highlighted limitations of

protected areas in maintaining key biodiversity features in landscapes (Laurance et al. 2012,

Clark et al. 2013). With the conservation focus primarily on particular areas, biodiversity

conservation in surrounding landscapes can be neglected (Bhagwat et al. 2005, Hansen and

DeFries 2007).

There has been increasing interest in the importance of forests outside the protected areas for

biodiversity conservation (Bhagwat et al. 2005, Persha et al. 2010). Off-reserve forests can be

important reservoirs of biodiversity that are complementary to the existing protected area

network in several ways. For example, off-reserve forests are often in different vegetation

types to those within protected areas, providing habitat resources that are poorly represented

within protected areas (Cox and Underwood 2011). For instance, tropical moist deciduous

and semi-evergreen forest in south Asia (Persha et al. 2010), evergreen mixed deciduous

forests in Thailand (Tantipisanuh and Gale 2013), and natural sacred forests in India

(Bhagwat and Rutte 2006) are predominantly represented outside of reserves, where

community management initiatives appear important for biodiversity conservation. Such

forests can also have greater habitat heterogeneity due to different disturbance regimes,

therefore supporting species that use various successional stages of habitat (Brawn et al.

2001, Chandler et al. 2012). As no single habitat necessarily provides all the required

resources for a given species’ persistence (Saunders et al. 1991, Becker et al. 2007),

conservation management of off-reserve forests can be essential for the persistence of many

species (Sodhi and Ehrlich 2010). Thus, effective off-reserve conservation policies help

ensure a diversity of habitat resources across the landscapes in which protected areas are

embedded.

In developing countries, about 22% of the total forest area is either community-managed or

owned, compared with only three percent in developed countries (White and Martin 2002).

Community forest initiatives have been increasingly successful in preventing deforestation

37

and restoration of forest condition in the landscapes (Klooster and Masera 2000; Nagendra

and Gokhale 2008, Porter-Bolland et al. 2012). Nepal offers some of the best examples of

community-based forest management in the world (Pokharel et al. 2007, Nagendra and

Gokhale, 2008). About one-fourth of forests in Nepal are currently managed by community

forest user groups (Kanel and Dahal 2008, Ojha et al. 2009). Rates of habitat loss and

degradation are reduced in community managed areas compared with state managed forests

(Nagendra 2007, Kanel and Dahal 2008). This success is primarily driven by effective

implementation of a decentralized forest management regime (Ojha et al. 2009). This

approach has increased active participation of local communities in conservation of resources

as they perceive community forest regime as a secure tenure for sustainable sources of forest

products (Gautam 2007, Kanel and Dahal 2008).

Protected areas and off-reserve forests differ in their habitat features due to different

management approaches, so it is likely that species richness and composition of birds may

vary among sites in different forest tenures. For example, landscapes with relatively

undisturbed and structurally complex forest habitat within the formal protected area systems

may support forest specialists and disturbance-intolerant species (Inskipp and Inskipp 1991,

BCN and DNPWC 2011). However, successional habitats of different stages within the

community managed forests may support more open country specialists (Roman 2001) and

mixed habitat species (Anand et al. 1997, Blake and Loiselle 2001). As differently-managed

forests offer habitat heterogeneity, appropriate conservation of these habitats can optimize

regional bird diversity. It is therefore important to quantify the contribution of these areas to

biodiversity conservation, and in particular, the extent to which the biota they support is

complementary to that within formal protected areas.

In this study, we examine the contribution of differently-managed forests to the conservation

of forest bird communities in lowland Nepal. The role of alternative forest management

tenures in biodiversity conservation is often neglected. In particular, while state-centric forest

management approaches tend to have spatially uniform management approaches, community

management approaches can be diverse, while also securing the right to resources and

embracing a participatory approach to the management of forest resources. Several studies

have investigated the effectiveness of the community forestry approach in reducing

deforestation and improving plant species richness and density (e.g. Agrawal et al. 2008,

Sapkota et al. 2009, Persha et al. 2010), However, it is important to examine whether

38

community managed forests in particular can play an important conservation role for forest

bird communities, complementing that of protected areas. We specifically aimed to: 1)

determine whether species richness, abundance and diversity of forest bird assemblages

varied among sites in community forests, state forests and protected areas, and 2) compare

the composition of forest bird assemblages among different management regimes to assess

conservation values of variously managed off-reserve forests for avian biodiversity in

Nepal’s lowland landscapes.

39

2.3 Material and methods

2.3.1 Study area

The study was conducted in the eastern and central Terai of Nepal. Nepal has a total

landmass of 147,181 km2 divided among three main geographical regions: the Himalayan

region, mid hill region and the Terai region. The lowland Terai encompasses most of the

country’s tropical moist forest from the Mechi River in the east to the Narayani River in the

centre. The area is characterized by a tropical climate, with average precipitation of

approximately 1,800 mm (Springate-Baginski et al. 2003) and mean maximum temperatures

of 15-400C (Sah et al. 2002). Before 1950, the region was an uninterrupted patch of dense

tropical forest. With the eradication of malaria in the early 1950s, the highly productive

lowland zone of the country was settled and subsequently agricultural expansion occurred

(Hrabovszky and Miyan 1987). Consequently, most of the forest was destroyed and

remaining forest areas were subjected to intense human exploitation. Nearly half of the

country’s population lives in the 17% of the country that is lowland (Central Bureau of

Statistics 2011).

The government of Nepal introduced and implemented forest legislation in 1978 with the aim

of diversifying the management tenures and reducing large-scale clearance of forest

(Department of Forest 2009). Thus, forests in Nepal are now managed under three major

regimes: as state managed forests (forests managed by the central government), community

managed forests (forests managed by local forest user groups), and protected forests (IUCN

management categories I-IV). The state-managed forests are those that are managed by the

Department of Forests as production forests, and protected areas are managed by Department

of National Parks and Wildlife Conservation for conservation. Protected areas in this study

corresponded to IUCN protected area management categories II (National Park) and IV

(Wildlife Reserve). The main aim of category II is to maintain ecological integrity at

ecosystem-scale, while the category IV is aimed to protecting habitats and individual species.

Approximately 1.2 million hectares of forests are currently managed by the community forest

user groups (>15000 community forest user groups) (Kanel and Dahal 2008, BCN and

DNPWC 2012) while ~1 million hectares of forests is directly managed by the central

government (Shyamsundar and Ghate 2011).

40

2.3.2 Study sites

A total of 112 sites were selected within lowland tropical forest within an elevational range of

90 – 300 m asl. These sites were allocated among three management tenures in approximate

proportion to the available area within each. We randomly allocated survey sites within

forests of each tenure type using digital vegetation mapping data. Initially, we chose 128 sites

using a GIS, but based on accessibility, we ended up with 44 sites within community

managed forests, 37 within the state managed forests and 31 within protected areas, including

in Chitwan National Park (IUCN category II protected area) and its buffer zone forest of

Barandabhar corridor, Parsa Wildlife Reserve (IUCN category IV protected area) which have

been managed for conservation for more than twenty-five years (Baral and Inskipp 2005).

The southern part of the Barandabhar core forest is managed by the park authority; its

peripheral areas are community-managed forests. Geographically, 60 sites were located in the

eastern landscapes (Eastern Terai forests) and 52 sites were located in the central lowland

landscapes (Parsa and Chitwan forests). The vegetation of the lowland Terai is mainly

consisted of Shorea robusta mixed forest. Therefore, all sites were located within the same

vegetation type. All sites were located at least 500 m from roads to minimize any road

induced variation on bird assemblages. The minimum distance between sites was at least

1000 m so as to reduce the chance of spatial dependence.

41

Figure 2.1 The location of the study area and bird survey sites: (a) Chitwan forests (b) Parsa

forests and (c) Eastern Terai forests

2.3.3 Bird surveys

Each study site comprised a fixed-width belt transect measuring 200 m x 50 m. Study sites

were demarcated by placing visible markers at each site and taking GPS coordinates. Each

transect was surveyed on three occasions between November 2012 and May 2013, across two

seasons. This arrangement of bird survey allowed us to capture winter visitors, winter

migrants and early summer visitors. Each site was visited on two occasions in winter and one

in summer. On each visit, the observer (BRD) recorded all birds seen or heard within 25 m of

the centreline of the transect while walking along its length over a 10-min period.

A variety of techniques have been employed to describe the characteristics of bird

populations. These include radio-telemetry (Powell et al. 2005), colour banding (Powlesland

et al. 2000, Rodewald et al. 2013), distance sampling (Buckland 2001), fixed-radius point-

counts (Gregory et al. 2004, Buckland 2006) and fixed-width transect-counts (Bibby et al.

2000, Westbrooke et al. 2003, Maron and Kennedy 2007). Both fixed-radius point-count and

fixed-width transect-count methods are widely used to describe the species richness, relative

42

abundance and densities of birds (Manuwal and Carey 1991, Buckland 2006Gregory et al.

2004). We therefore used fixed-width transect counts in order to compare the relative

abundance and richness of birds per unit area among different sites (Manuawal and Carey

1991, Bibby et al. 2000, Hostetler and Main 2011).

Surveys were conducted only between 0600 and 1100 hours in the morning and 1400 to 1745

in the afternoon. Although we did not test for effects of time of day on bird observation prior

to actual field survey; several other studies reported that the detection rate of most bird

species is greater in morning (Bried et al. 2011) with another peak in activity in the late

afternoon, 2-3 h before sunset (Kessler and Milne 1982). Generally birds tend to avoid the

midday heat (Pizo et al. 1997), therefore we surveyed birds within 4 h after sunrise and

within 3.45 h before sunset. To avoid possible bias, we standardized the survey protocol in

such a way that although not all sites had afternoon surveys, this occurred equally among site

categories, and so no bias was introduced due to this. All surveys were conducted by the

same observer during fair weather at no heavy rain and wind.

2.3.4 Explanatory variables

Data on vegetation and habitat structure were collected at each bird survey transect. Using

four randomly-located 20 m x 20 m quadrats, the percentage of tree canopy cover was

estimated, the number of trees counted, and their diameters measured within the 20 m x 20 m

quadrat. Tree cover was estimated visually (Pattison et al. 2011). We divided the quadrat into

quarters, and assessment of tree canopy cover was determined by two observers for each

quarter. The cover values for each quarter were then averaged and the four mean values for

each quadrat averaged, before a grand mean was calculated for the site. Nested within each of

the 20 m x 20 m tree quadrats was a 5 m x 5m quadrat, used to collect understorey vegetation

data. The shrub cover and number of individual shrubs were collected within each of these

nested quadrat and the grand mean taken for each transect. We calculated the total area of

forest habitat (in ha/km2) within a 500 m buffer distance from each bird survey sites in a GIS

(using ArcGIS 9.3). We included both primary and old growth regenerating forests to classify

forest habitat based on land-cover data provided by WWF Nepal (WWF 2005). Stand and

landscape-scale explanatory variables used in analyses are summarised in Table 2.1.

43

Table 2.1

Summary of explanatory variables used to assess influence of stand and landscape-scale

variables on bird communities.

Variables Unit Description

Forest extent Hectares/km2

Amount of forest area in 500 m radius of survey site.

Tree canopy cover Percent Mean percentage cover of all tree crowns in 200 m x 50

m line transects.

All trees density Count/m2

Number of trees with >10 cm diameter at breast height

(DBH) per square metre.

Mature large trees

density

Count/m2

Number of trees with >100 cm DBH per square metre.

Shrub cover Percent Mean percentage cover of all shrubs <2 m tall, in 200 m

x 50 m line transect.

Shrub density Count/m2

Number of individual shrubs <2 m tall, per square

metre.

2.3.5 Statistical analyses

2.3.5.1 Univariate analysis

We compared average abundance per survey, species richness, evenness (as measured by

Pielou evenness index) and species diversity (as measured by Shannon diversity index) of

bird assemblages among sites within the three management tenures using a one-way analysis

of variance (ANOVA) followed by Tukey’s Honestly Significant Difference (HSD) test. We

also compared abundance, species richness, evenness and diversity of birds with different

habitat associations (forest specialists, forest generalists and forest edge specialists) and

within different foraging guilds among the management tenures. Due to large number of zero

values for diversity and evenness of forest specialists and bark gleaning insectivore, we only

included species richness and abundance of these groups in statistical comparisons.

Membership of habitat groups and foraging guilds was identified based on primary habitat

specialization and diet information compiled from Ali and Ripley (1983) and Grimmett et al.

(2009). In addition, we compared habitat characteristics among different management

tenures. All analyses were carried out using the R statistical package version 3.1 (R

Development Team 2012).

44

2.3.5.2 Multivariate analyses

Multivariate data analyses were employed to examine variation in bird assemblage

composition among forest management regimes. We used nonparametric, permutational

multivariate analysis of variance PERMANOVA (Anderson 2001) based on Bray-Curtis

similarity values from species abundance matrix to test whether bird community composition

differed across management tenures. In this analysis, management tenures were considered

factors and sites were samples. Bird abundances were square-root transformed prior to

analysis to reduce the influence of highly abundant species. The test statistic for

PERMANOVA is the pseudo F-ratio, where a large pseudo F-ratio indicates that sites in

different management tenures are differed in bird community composition in multivariate

space. We performed additional pairwise PERMANOVA tests (Anderson et al. 2008) to

explore the extent of differences in species composition between the management tenures.

A significant pseudo-F ratio with P-values from the PERMANOVA can indicate a difference

in community composition between treatments due either to differences in the location of the

treatment communities in multivariate space or to differences in dispersion of communities in

multivariate space within the treatments (Anderson et al. 2008). Thus, we used a

complementary analysis, the permutational analysis of multivariate dispersions (PERMDISP)

(Anderson et al. 2006) to test for differences in the homogeneity of multivariate dispersion

among sites in different management tenures. Following a finding of significant differences

in dispersion, we performed pairwise tests. We also calculated mean species turnover (i.e.

beta diversity) using the vegan package in R (Oksanen et al. 2011) for each forest

management tenure. Beta diversity is the rate of change of species composition among sites

or habitat patches (Whittaker 1972). Understanding beta diversity among sites across

different management tenures allowed us to quantify their contribution to the regional

avifaunal diversity. We also evaluated which species were most responsible for

differentiating communities using similarity percentage (SIMPER) analysis. SIMPER

evaluates the contribution of each species to the Bray-Curtis dissimilarity of all pairs of

samples between groups (Clarke and Warwick 2001). These analyses were carried out using

PRIMER v6 with PERMANOVA+ add-on software (Anderson et al. 2008).

Non-metric multidimensional scaling (NMDS) ordination was conducted using Bray-Curtis

similarities to visualize pattern of bird assemblage among management tenures (Clarke

1993). Ordination serves to summarize community data by producing a low-dimensional

45

ordination space in which distance between species and samples sites reflect the ecological

differences between them (Gauch 1982). We performed a vector-fitting routine using the

vegan package in R (Oksanen et al. 2011) to examine the relationship between bird

communities with environmental variables. A vector-fitting protocol allows us to visualize

the most contributing variables to the pattern of bird community composition (Johnson et al.

2007). The length of the fitted arrow is proportional to the correlation between the ordination

axis and the environmental variable. Environmental variables were standardized to lie

between 0 and 1 prior to analysis.

2.4 Results

2.4.1 Species richness, abundance and diversity

A total of 124 bird species from 28 families were recorded across all sites over three seasons

(Appendix A: Table A.1). Of all recorded species, 68% were local residents, 16% winter

visitors, 2% summer visitors and 4% winter passage migrants. In all three surveys, the most

widespread species was the black-hooded oriole Oriolus xanthornus, which occurred at 110

of the 112 survey sites. The grey-headed canary flycatcher Culicicapa ceylonensis, spangled

drongo Dicrurus hottentottus, and jungle babbler Turdoides striata, were the next most

widespread species, each being recorded 80%, 79%, and 78% of sites, respectively

(Appendix A: Table A.1).

Ninety-three species were recorded in community managed forests, 87 species were recorded

in the state managed forests, and 80 species were recorded in protected areas. Some species

of birds were distinct to each of the forest management regimes. Seventeen species were

recorded only in sites in community managed forests, 13 species were in sites in state

managed forests, and 15 species were in sites in protected areas. Only 45% of the recorded

species were found in all three management tenures.

Among the three forest management tenures, neither the total bird abundance (F2, 109 = 0.83,

P > 0.05), total species number (F2, 109 = 2.44, P > 0.05) nor Pielou evenness index (F2, 109 =

0.26, P > 0.05) differed significantly. However, there was a statistically significant difference

in species diversity (Shannon diversity index; F2, 109 = 4.79, P < 0.05) among management

tenures (Table 2.2). A post-hoc Tukey test showed that differences in species diversity were

between state forests and community forests (P < 0.01), and between state forests and

46

protected areas (P < 0.05). There was no significant difference in species diversity between

sites in community forest and protected areas (P > 0.05).

Table 2.2

Mean species richness, average abundance per survey, Shannon diversity index and Pielou

evenness index (±SE) of all birds, in community managed forests (n = 44), state managed

forests (n = 37) and protected areas (n = 31).

Variables Community forest State forest Protected area F value

Species richness 19.2 ± 0.5 17.5 ± 1.1 20.0 ± 0.3 2.47

Abundance 23.3 ± 1.3 23.0 ± 1.6 25.2 ± 1.3 0.83

Shannon diversity index 2.6 ± 0.0 2.4 ± 0.1 2.7 ± 0.1 4.79*

Pielou evenness index 0.9 ± 0.0 0.9 ± 0.1 0.9 ± 0.1 0.26

*Significant at P < 0.05

2.4.2 Habitat preference groupings

We recorded a total of 14 species of birds classified as forest specialists, 30 forest edge

species and 75 forest generalist species. Overall, significantly higher mean species richness

and abundance of forest specialist birds were observed in sites in protected areas followed by

community managed forests and state managed forests (Table 2.3).

Table 2.3


evenness index (±SE) of all birds in different habitat in community managed forests (n = 44),

state managed forests (n = 37) and protected areas (n = 31). Degrees of freedom between

groups (management types) = 2 and within groups = 109 in all cases.

Habitat

group Variables

Community

forest

State

forest

Protected

area F value

Forest

generalists

Species richness 11.5 ± 0.4 10.2 ± 0.4 11.3 ± 0.5 3.22*

Abundance 15.3 ± 0.8 15.5 ± 1.0 16.2 ± 0.2 0.58

Shannon diversity index 2.2 ± 0.0 1.9 ± 0.0 2.1 ± 0.1 4.32*


Forest edge

species

Species richness 6.7 ± 0.3 5.9 ± 0.5 6.4 ± 0.4 0.84

Abundance 7.5 ± 0.7 6.6 ± 0.7 7.2 ± 1.7 0.31

Shannon diversity index 1.6 ± 0.0 1.4 ± 0.1 1.6 ± 0.1 3.02

47

Pielou evenness index 0.9 ± 0.0 0.8 ± 0.0 0.9 ± 0.0 5.28**

Forest

specialists

Species richness 2.2 ± 0.2 2.1 ± 0.3 2.7 ± 0.3 5.91**

Abundance 1.3 ± 1.0 0.8 ± 0.2 1.9 ± 0.2 6.64**

Shannon diversity index 0.4 ± 0.0 0.3 ± 0.0 0.7 ± 0.1 -

Pielou evenness index 0.9 ± 0.0 0.8 ± 0.0 0.9 ± 0.0 -

Significant at *P < 0.05; **P < 0.01

2.4.3 Foraging guilds

The sites in protected areas had the greatest abundance, species richness and diversity of bark

gleaning insectivore, followed by community forests and then state forests (Table 2.4).

However, other foraging guilds varied little among management tenures (Table 2.4). While

we identified seven main foraging guilds of birds, we excluded those groups that have large

number of zero values in the analyses, leaving only four that could be analysed (Table 2.4).

Table 2.4


evenness index (±SE) of different foraging guilds of birds in community managed forests (n

= 44), state managed forests (n = 37), and protected areas (n = 31). Degrees of freedom

between groups (management types) are 2 and within groups are 109 in all cases for food

guild.

Food guild Variables Community

forest

State

forest

Protected

area F value

Bark

gleaning

insectivore

Species richness 2.9 ± 0.3 2.5 ± 0.2 3.8 ± 0.3 7.58***

Abundance 3.5 ± 0.4 2.2 ± 0.3 4.5 ± 0.4 10.37***

Shannon diversity index 0.7 ± 0.1 0.7 ± 0.0 1.1 ± 0.0 -

Pielou evenness index 0.9 ± 0.0 0.8 ± 0.1 0.8 ± 0.0 -

Foliage

gleaning

insectivore

Species richness 2.7 ± 0 .2 3.3 ± 0.3 3.3 ± 0.2 1.93

Abundance 3.2 ± 0.4 4.7 ± 0.7 4.5 ± 0.5 2.51



Sallying

insectivore

Species richness 4.0 ± 0.2 3.3 ± 0.3 3.3 ± 0.2 3.90*

Abundance 3.5 ± 0.4 2.8 ± 0.3 3.0 ± 0.3 2.25



Frugivore Species richness 5.9 ± 0.3 5.4 ± 0.3 5.3 ± 0.3 1.22

48

Abundance 9.5 ± 0.7 8.8 ± 0.8 8.1 ± 0.6 0.8



Significance at*P < 0.05; **P < 0.01;***P < 0.001

2.4.4 Bird assemblage structure

Significant differences in community composition of birds were observed among sites in all

three forest management tenures (PERMANOVA: F = 3.56, P < 0.001). The PERMANOVA

pairwise a posteriori comparison tests showed that community composition differed between

each pair of tenures (P < 0.05, Table 2.5). The greatest difference in community composition

was between sites in community forests and those in protected areas (P < 0.001), and

protected areas and state forests (P < 0.001). PERMDISP analyses for the homogeneity of

multivariate dispersions also showed significant differences in community composition

among the forest management tenures (F = 5.56, P < 0.01). However, no significant

difference in multivariate dispersion was observed between protected areas and community

forests. Thus, the significant difference in community composition in sites between

community forests and protected areas from the PERMANOVA was due to differences in the

location of sites of community forests and protected areas in multivariate space rather than

differences in dispersion around the mean composition within sites in the community forests

and protected areas (Table 2.5). The multivariate dispersion in community composition in

sites in state managed forests was significantly larger, having an average Bray-Curtis

distance-to-centroid over 44% greater than sites in protected areas (39%) and community

forests (39%). A separate beta diversity (i.e. species turnover) analysis showed that the mean

variation in species composition was highest among sites in state managed forests (mean β

diversity = 0.55 ± 0.1) than among sites in sites in protected areas (0.48 ± 0.1) and

community managed forests (0.48 ± 0.1).

Table 2.5

Results of PERMANOVA and PERMDISP pairwise comparison tests of bird community

composition in response to different forest management tenures.

PARMANOVA test PERMDISP test

Management tenures t-value t-value

Community forest - Protected area 2.22*** 0.24

49

Community forest - State forest 1.32* 3.23***

Protected area - State forest 2.07*** 2.56***

Significance at*P < 0.05; **P < 0.01;***P < 0.001

Non-metric multidimensional scaling (NMDS) ordination of species composition showed

strong clustering of sites according to forest management tenures, although stress was

relatively high (Fig. 2.2). Vector fitting of environmental variables to the bird assemblage

NMDS ordination showed significant correlation of environmental variables (Fig. 2.2).

Among the environmental variables, the amount of forest within a 500 m radius had a

significant influence on bird community structure (P < 0.001). Significant correlations were

found with mature tree density (P < 0.01), tree canopy cover (P < 0.01), mean shrub cover (P

< 0.01) and shrub density (P < 0.5) in the study area.

Figure 2.2 Bird assemblage NMDS (2D stress = 0.24) plot fitted with environmental vectors.

Vectors represent the mean direction and strength of correlation of different environmental

50

variables (Fextent = Forest extent, Scv = Shrub cover, Sden = Shrub density, Mtdn = Mature

tree density, Tcv = Tree canopy cover). Vectors with significance P < 0.05 only are shown.

The SIMPER analyses identified the species responsible for distinguishing community

composition between management tenures. The rose-ringed parakeet Psittacula krameri

contributed most to the dissimilarity between sites in protected areas and community forests,

being more common in sites in community forests, while the jungle babbler Turdoides striata

contributed most to the dissimilarities between sites in state forests and community forests,

and between protected areas and state forests, as it was commonest in sites in protected areas

and community forests. The top ten bird species that contributed to differences between

management tenures are presented in Appendix A: Table A.2.

2.4.5 Habitat characteristics

There were significant differences in habitat characteristics among different management

tenures. Habitat characteristics such as tree density (F2, 109 = 6.47, P < 0.01), shrub cover (F2,

109 = 4.96, P < 0.01), tree canopy cover (F2, 109 = 3.58, P < 0.05) and mature tree density (F2,

109 = 2.91, P < 0.05) differed significantly among sites in all three management regimes.

However, we found no significant difference in shrub density (F2, 109 = 2.05, P > 0.05) among

management tenures. Sites in state managed forests had the highest average number of small

trees (<20 cm diameter DBH), while community forests and protected areas had highest

average number of large mature trees (>100 cm DBH) (Fig. 2.3).

Figure 2.3 Distribution of tree size classes (±S.E.) in sites in community forests, state forests

and protected areas.

0.00

0.01

0.01

0.02

0.02

0.03

0.03

0.04

0.04

0.05

0.05

>10 dbh 20-49 dbh 50-99 dbh >100 dbh

Tre

e d

ensi

ty (

stem

s/m

2)

Tree size classes (cm)

Community forest

Protected area

State forest

51

2.5 Discussion

While sites in protected areas had the greatest richness of birds, community forests and state

managed forests had complementary assemblages, supporting species not represented in

formal conservation reserves. In this study 17 species of birds were recorded only in

community forests, among which Abbott’s babbler Malacocincla abbotti and blue-eared

barbet Megalaima australis are nationally threatened (Inskipp 1989, Baral and Inskipp 2004;

Inskipp et al. 2013). In total, 24% of species were found only in forests outside the protected

areas, and only 45% of species were common to all forest tenures. This distinctness of bird

assemblages in off-reserve sites contributes to diversity in Nepal’s lowlands.

Overall bird community composition differed among the three management tenures,

indicating that differently managed off-reserve forests support distinctive bird assemblages.

While sites in community forests and state forests had complementary bird assemblages,

protected areas emerged as particularly valuable, with significantly greater species richness

and diversity of forest specialists and bark-gleaners. The greater community indices of forest

specialists and bark-gleaners sites in protected areas are likely to reflect the larger extent of

mature forests within the Terai protected areas (Wikramanayake et al. 2004). Furthermore,

the higher richness and diversity of forest specialists in sites in protected areas may be related

to the fact that anthropogenic disturbance is limited in such areas (Baral and Inskipp 2005).

Several studies in the region show that extraction of fodder, firewood and non-timber forest

product can negatively influence bird communities (Shahabuddin and Kumar 2007, Dahal et

al. 2009, Kumar et al. 2011; Inskipp et al. 2013). Disturbance-intolerant species may

therefore be benefited by strict forest management that restricts the removal of standing dead

trees, fallen timber for firewood and canopies by pruning.

On several measures, including overall Shannon diversity and the species richness and

abundance of bark-gleaners and forest specialists, sites in community forests were most

similar to those in protected areas. This similarity is likely to reflect the relatively more

similar habitat characteristics of sites in protected areas and community forests, with no

differences in tree canopy cover, large mature tree density and shrub density between sites in

community managed forests and protected areas. Similarities in habitat structure between

community forests and protected areas have also been noted by Timilsina et al. (2007), who

found that community forests and protected areas had similar tree density in Nepal’s western

lowlands, and Nagendra (2002) who reported similar tree and sapling biomass between

52

protected areas and community forests. A recent study by Persha et al. (2010) in three south

Asian countries; Nepal, India and Bhutan found that tree species richness in community

forests was equivalent to that of protected areas. This demonstrates the potentially valuable

contribution of community forests to provision of habitat of similar quality to that within

protected areas.

Although sites in community forests and protected areas had higher bird diversity, state

managed forests had the greatest multivariate dispersion. This reflects higher species turnover

among sites in state managed forests. A comparatively high species turnover among sites in

state managed forests may be due to number of factors. First, forest management practices

such as selective logging, thinning and pruning are commonly employed in much of the state

managed forests. Such silvicultural practices cause temporary increases in habitat

heterogeneity and this can increase spatial and temporal variation in species richness and

abundance of birds (De La Montana et al. 2006, Forsman et al. 2010). Second, of 80 species

of birds recorded in sites in state managed forests, 28% were singleton records (i.e. species

represented by detection of a single individual during sampling) compared to 15% each for

protected areas community forests (Appendix A: Table A.1). These results suggest that,

though state managed forests had complementary bird assemblages, these forests may not

necessarily provide suitable habitats for long-term persistence of populations of several

species we detected.

While overall species richness and abundance of birds in sites in community forests and state

managed forests were similar, bird assemblage structure was significantly different. In this

study only 57% of species were common to sites in community managed forests and state

management forests. In particular, forest specialists and bark gleaners were less abundant in

sites in state managed forests than those in community managed forests. This is likely to

reflect the altered habitat conditions in state managed forests (Kanel and Dahal 2008). Large

mature trees are removed from such forests for their timber, leaving a high density of small

trees in sites (Fig. 2.3). In contrast, community-managed forests often include arrangements

to reforest and restore degraded forest, which can contribute to maintaining habitat for

specialist and rare bird species in Terai forests (Baral and Inskipp 2005).

Among the dietary groups, bark gleaners differed most markedly in their richness, abundance

and diversity among the management tenures. The greater abundance of bark-gleaning

53

insectivores in sites in protected areas and community forests may be linked to both the

greater extent of forest surrounding such sites, and the higher density of large trees. Several

other studies have found that bark gleaning birds are particularly sensitive to habitat

alteration (Adams and Morrison 1993, Zurita and Bellocq 2012, Inskipp et al. 2013), and are

strongly correlated with large tree density (Cleary et al. 2007, Greve et al. 2011).

Furthermore, forest resource extraction like harvesting of standing dead trees and fallen

timber for firewood may reduce foraging and nesting sites for many bark-foraging species.

Fallen timber and dead trees are linked to nesting success of bark foraging birds including the

greater flameback Chrysocolaptes lucidus, grey-headed woodpecker Picus canus and greater

yellownape Picus flavinucha (Kumar et al. 2011). Bark-foraging bird communities may

therefore be benefited by forest management that retains higher densities of large and dead

trees as foraging and nesting habitats.

2.6 Management implications

While the lowlands of the Terai in Nepal contain about 28% of country’s forested areas

(Sinha 2011), less than 10% of the lowlands forests are formally protected. Nearly half of the

country’s population lives in the lowlands (Central Bureau of Statistics 2011) and the

majority of them rely directly on resources from their surrounding natural areas. For instance,

about 80% of the total energy consumed in the country is produced by fuel wood, extracted

from these forests (Gurung et al. 2011).

Although protected areas serve as essential refugia for most species of forest birds, the further

extension of protected areas in lowland landscapes is likely to be limited due to economic,

political and social factors. Instead, participation of a wide spectrum of stakeholders and

institutions in forest management, such as through community forest arrangements, can be

used to complement the contribution of protected areas to biodiversity conservation. The

community forestry program in Nepal has demonstrated its value for improving forest

conservation outside the protected areas (Kanel and Dahal 2008, Nagendra et al. 2008, this

study). Yet, there has been reluctance to transfer management rights to local communities in

the lowland Terai region due to the prevalence of forests with high economic value (Bhattarai

2006). Thus, only 10% of the Terai forests have transferred to community management

(Kanel and Dahal 2008), compared to 24% in the hill regions of Nepal (Bhattarai 2006).

Strengthening the community forestry programs across the off-reserve forests in lowland

landscapes can not only ameliorate habitat loss and degradation (Gautam et al. 2004, Kanel

54

and Dahal 2008) but also generate livelihood opportunities for surrounding communities and

reduce pressure on protected areas (Straede et al. 2002). Hence, the maintaining regional bird

diversity in lowland forest landscapes critically depends on complementary management of

forests both outside and inside the established protected areas.

55

CHAPTER 3

IMPACTS OF EXTRACTIVE FOREST USES ON BIRD ASSEMBLAGES VARY

WITH LANDSCAPE CONTEXT IN LOWLAND NEPAL

Plate 3: Creekbed in a large patch of forest adjacent to an agricultural area in Parsa Wildlife

Reserve, lowland Nepal.

Published as: Dahal, B. R., C. A. McAlpine, and M. Maron. 2015. ‘Impacts of extractive

forest uses on bird assemblages vary with landscape context in lowland Nepal’. Biological

Conservation. 186: 167-175.

56

3.1 Abstract

Forest use practices such as logging, lopping of tree branches for fodder, and grazing do not

reduce forest area but disturb forest structure and impact biodiversity. Although such forest

disturbances can be key determinants of the biota occupying a site, rarely is the interaction

between disturbance intensity and landscape context considered, despite its relevance to

conservation management. We investigated the influence of site-and landscape-level habitat

characteristics on birds, and explored whether the effects of site-level disturbance on bird

richness varied with forest extent in lowland landscapes in Nepal. While extractive uses

reduced forest structural complexity and altered the avifaunal community of a site, the

intensity of such effects depended on the extent of forest in the surrounding landscape (19.6

km2). The extent of forest, large tree density, and tree canopy cover were important predictors

for all bird response groups. However, the effect of forest extent on bird richness was

stronger for sites with greater disturbance intensity. Managing and restoring landscapes to

support greater forest cover may not only have a positive direct effect on bird conservation,

but may also help to compensate for site-level disturbance, such as characterises multiple-use

forests worldwide.

Key word: Avian community, multi-use forest landscapes, extractive forest disturbances,

landscape context, interactive effect, landscape-level forest cover

57

3.2 Introduction

In recent decades, anthropogenic activities have been the principal cause of habitat loss and

degradation worldwide (Millennium Ecosystem Assessment 2005, Foley et al. 2005, Ellis and

Ramankutty 2008). However, anthropogenic habitat degradation is greater in areas of high

population density and poverty. Such areas are mainly in developing countries (Laurance

2010) where a significant proportion of the population live near the forests (Hegde and Enters

2000, Millennium Ecosystem Assessment 2005). About one billion people living in

developing countries rely on forest-based products, primarily for subsistence livelihoods

(Chao 2012). This has resulted in extensive use of forest resources, for timber and firewood,

cutting of tree canopy for fodder, livestock grazing and collection of non-timber forest

products (Chettri et al. 2002, Shahabuddin and Kumar 2007, Christensen et al. 2009). Thus,

managing forests for biodiversity conservation while satisfying human demands for forest

products is a major global conservation challenge (Chappell and LaValle 2011).

Anthropogenic activities can reduce the forest area and also cause significant changes in

forest structure and composition (Chettri et al. 2002, Sagar and Singh 2004, Shahabuddin and

Kumar 2006). Repeated extraction of timber resources reduces tree basal area, tree height,

canopy closure, and regeneration capacity (Sundriyal and Sharma 1996, Mishra et al. 2004,

Sapkota et al. 2010). For example, removal of live trees increases light levels in the forest,

thereby modifying canopy structure (Sekercioglu et al. 2002, Villela et al. 2006), altering tree

density and diversity (Moktan et al. 2009), and changing understorey characteristics (Aleixo

1999, Moktan et al. 2009). The logging and forest extraction practices also change tree

diversity and composition (Kumar and Shahabuddin 2005, Berry et al. 2008, Sapkota et al.

2009, Borah et al. 2014). Other forms of extraction of woody biomass such as lopping of tree

branches affects canopy structure. Livestock grazing also simplifies the understory forest

structure and reduces regeneration, foliage density, canopy height, and vegetation cover

(Tasker and Bradstock 2006, Piana and Marsden 2014). Such loss of structural components

and alteration of floristic diversity in forests ultimately affects populations of many species

reliant on forest habitat (Díaz et al. 2005, Berry et al. 2008, Lee and Carroll 2014).

Habitat variables measured at the site-scale (<1 ha), however, may not be sufficient for

meaningful prediction of species responses to disturbance type and intensity. Rather, the local

effects of such anthropogenic disturbances on forest fauna may also depend on the landscape

context (100s-1000s ha) in which a site is embedded. Sites within more forest cover may

58

offer greater habitat heterogeneity (Heikkinen et al. 2004, Kallimanis et al. 2008), which may

potentially govern species richness in the landscape (Tews et al. 2004). Furthermore,

increasingly, faunal communities are being shown to be affected strongly by the proportion of

forest habitat in a landscape (McGarigal and McComb 1995,Trzcinski et al. 1999, Radford et

al. 2005, Ewers and Didham 2006 , Smith et al. 2011). Studies in fragmented landscapes

suggest that landscape context - in particular, surrounding forest extent - mediates the effects

of fragmentation on faunal communities (e.g Graham and Blake 2001, Deconchat et al.

2009). It is therefore plausible that landscape-scale variables, such as the extent of forest,

may actually moderate the effects of site-scale anthropogenic impacts such as subsistence

forestry practices on faunal communities, and vice-versa. Yet knowledge of such interactions

between the extent of forest in the landscape and the impact of forest disturbances remains

limited.

In this study, we first investigated the effects of anthropogenic disturbance on vegetation

structure and consequences for bird communities in the lowland Terai forests of Nepal. This

region is dominated by the highly productive Sal forests, which are facing significant

anthropogenic pressure from extractive and grazing uses. The economy of rural communities

in the region is based largely on subsistence agriculture, livestock rearing, and selling of

firewood and non-timber forest products (Sharma 1990). Such activities have contributed

elsewhere to a decline of many forest bird species (Inskipp et al. 2013, Baral et al. 2014),

particularly species with small home range and/or other ecological requirements (Inskipp

1989).

Second, we modelled the effect of the interaction between landscape context and disturbance

intensity on the bird assemblages of these forests. We hypothesized that forest disturbances

will negatively affect vegetation structure and bird communities, but that disturbance

intensity will interact with the extent of forest in the landscape to affect the avifauna of a site.

Specifically, our main objectives were to: (1) determine whether the vegetation

characteristics and species richness and abundance of forest bird assemblages varied with

logging, grazing, and lopping intensity; and (2) assess the relative importance of site-and

landscape-scale forest habitat characteristics on bird species richness and abundance and the

existence of any interaction effects between disturbance intensity and landscape context.

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3.3 Material and methods

3.3.1 Study area

The study was conducted in southern Nepal, also called ‘Terai’ (80° 4’ 30” to 88° 10’ 19” E

26° 21’ 53” to 29° 7’ 43” N, elevation 63 - 330 m ASL). The Terai encompasses most of the

country’s tropical moist forest from the Mechi River in the east to the Narayani River in the

centre. The annual rainfall decreases from 2,680 mm to 1,138 mm from east to west, and the

mean monthly rainfall ranges from 8 mm in November to 535 mm in July (FRA/DFRS

2014). The area is characterized by a tropical climate, with the maximum monthly mean

temperature of 35-40°C in April/May and the minimum, 14-16 °C, in January (Jackson et al.

1994). Before 1950, the region supported continuous dense tropical forest. With the

eradication of malaria in the early 1950s, large tracts of the highly productive lowland forests

were converted to agriculture (Hrabovszky and Miyan 1987). Consequently, most of the

forest was destroyed and the remaining forest areas were subjected to intense human

exploitation. Nearly half of Nepal’s population now lives in the 17% of the country that is

lowland (Central Bureau of Statistics 2011).

Figure 3.1 Location of the three study regions in Nepal: a) Chitwan forest, b) Parsa-Bara

forest and c) Eastern forest.

3.3.2 Study sites and landscapes

Twenty-eight landscapes, each 5 km x 5 km, and supporting different amounts of forest cover

(7.9% - 95.3%), were selected across south-central (Bara-Parsa forest and Chitwan forest).

and south-eastern lowland Terai forests (eastern forests) among three tenure types.

60

Geographically, 15 landscapes were located in eastern Terai forests and 13 landscapes were

located in central lowland Terai forests. Four survey sites, each measuring 200 m x 50 m,

were randomly located in each landscape, resulting in a total of 112 sites (28 landscapes x 4

sites). Of the total 112 sites, 44 sites were in community manage forests, 37 were in state-

managed forests and 31 were in protected areas. The non-forested part of landscapes in this

region is mainly comprised of a mixed land–use type that includes rural towns, agriculture

and agro-forestry. All sites were located at least 500 m from roads to minimize any road-

induced variation in bird assemblages. The minimum distance between sites was at least 1000

m to reduce the chance of spatial dependence.

3.3.3 Bird surveys

At each study site, birds were surveyed on three occasions between November 2012 and May

2013 allowing us to capture winter visitors, winter migrants and early summer visitors. Each

site was visited on two occasions in winter and one in summer. On each visit, the observer

(BRD) recorded all birds seen or heard within 25 m of the centreline of the transect while

walking along its length over a 10-min period. Prior to the data collection, we tested for

visibility of birds within 50 m and 25 m of the transect, and found that visibility beyond 25 m

was challenging. To reduce the risk of sampling bias (Järvinen and Väisänen 1975), we used

a rangefinder to help ensure all birds counted were within the fixed belt transect.

Surveys were conducted only between 0600 and 1100 hours in the morning and 1400 to 1745

in the afternoon. Although the effects of time of day on bird observation were not tested prior

to the actual field survey; several other studies have reported that the detection rate of most

bird species is greater in morning (Bried et al. 2011) with another peak in activity in the late

afternoon, 2-3 h before sunset (Kessler and Milne 1982). In general, birds tend to reduce

activity during the midday heat (Pizo et al., 1997). To avoid possible bias, we standardized

the survey protocol in such a way that, although not all individual sites had an afternoon

survey, afternoon surveys occurred equally among sites. All surveys were conducted during

fair weather when there was no heavy rain and the wind speed was low.

3.3.4 Explanatory variables

Data on vegetation and habitat structure were collected at each bird survey transect (Table

3.1). Four 20 m x 20 m quadrats were randomly located on each transect. For each quadrat,

we measured the percentage of tree canopy cover, the number of trees, and their diameter at

61

breast height. Tree canopy cover was estimated following Pattison et al. (2011) by dividing

the quadrat into quarters, and visually assessing canopy cover for each quarter. The values for

each quarter were then averaged and the four mean values for each quadrat averaged to

provide an overall mean for the site. Nested within each of the 20 m x 20 m quadrats was a 5

m x 5 m quadrat, used to collect understorey vegetation data. Shrub cover estimates and the

number of individual shrubs were collected within each nested quadrats and an overall mean

calculated for each transect. Similarly, herbaceous cover was estimated from a 1 m x 1 m

quadrat, nested within each of the 20 m x 20 m tree quadrats. The extent of forest habitat (in

ha/km2) within a 0.5 km and a 2.5 km buffer distance from each bird survey sites was

calculated using GIS (using ArcGIS 9.3). We included both primary and old growth

regenerating forests to classify forest habitat based on land-cover data provided by WWF

Nepal (WWF 2005). These spatial extents, while arbitrary, were chosen areas as likely to

represent the likely daily area of use for many individual birds and the extent over which a

species might range throughout the landscape over a year. The density of paved roads within

each landscape was also calculated using ArcGIS.

As one of our objectives was to investigate the effects of forest disturbance on birds,

indicators of disturbance due to forest-use practices were recorded to reflect the intensity of

disturbances for each site. Forest-use practices such as livestock grazing (cattle), logging and

lopping are the major forms of anthropogenic disturbance in the lowland Terai forests.

Livestock grazing tends to result in changes to understorey species composition and structure;

logging involves the removal of trees >20 cm diameter for timber production, house

construction and fuelwood, and lopping is usually for fodder and small fuelwood and

involves removal of tree branches 5 – 20 cm diameter. In each quadrat, all the logging

stumps, lopping trees, and dung piles were counted. The values of each disturbance variable

across the four quadrats were averaged for each site.

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Table 3.1

Summary of explanatory variables used to assess the influence of site and landscape-scale

variables on bird communities.

Variables Unit Description

Site-scale

Tree canopy cover Percent Mean percentage cover of all tree crowns in 200

m x 50 m line transects.

Large trees density Count/m2

Number of trees with >100 cm DBH per square

metre.

Shrub density Count/m2

Number of individual shrubs <2 m tall, per

square metre.

Landscape-scale

Forest extent (0.79

km2)

Hectares/km2

Amount of forest area in a 0.5 km radius of

survey site.

Forest extent (19.6

km2)

Hectares/km2

Amount of forest area in a 2.5 km radius of

survey site.

Road density Meters/hectare Total length of paved roads divided by the area

(ha) of each 5 x 5 km study landscape

3.3.5 Statistical analysis

We compared overall species richness, average bird abundance per survey, and species

richness and mean abundance of bark-gleaning and foliage-gleaning birds among sites that

differed in logging, grazing and lopping intensity using a three-way analysis of variance

(ANOVA). To do this, we classified sites into heavily logged or lightly logged, heavily

grazed or lightly grazed and heavily lopped or lightly lopped based on the extent of

disturbance intensity (Appendix B: Table B. 7).. As the livelihoods of the local population

partly depend on extraction of adjacent forest resources, these subsistence activities are

common forms of forest disturbances and often occur together in the region.

Sites with ≤ 1 cut stumps per quadrat on average were categorized as lightly logged and >1 as

heavily logged; sites with ≤ 2 dung clusters on average were categorized as lightly grazed and

> 2 as heavily grazed; sites with ≤ 5 lopped tree branches on average were categorized as

lightly lopped and > 5 as heavily lopped. We used only two categories for each disturbance

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type (i.e. lightly vs. heavily disturbed) because the level of disturbance among the categories

were very different, indicating distinctness among categories in terms of disturbance

intensity. We used Multidimensional Scaling (MDS) ordination to visualize the pattern of

disturbances among different disturbance categories. As we included three tenure types in

this study, the majority of sites in protected areas and community managed forests

experienced less human disturbance and were separately clustered from those in state-

managed forests. The state-managed forests were most heavily disturbed. In addition to the

comparison of bird responses, we also compared all vegetation variables (Appendix B: Table

B.1 and B.2) among these site categories, also using a three-way analysis of variance

(ANOVA). These analyses were conducted using the ‘lm’ function in package “car” (Fox et

al. 2012) in the R environment (R Core Team, 2014).

We used generalised linear mixed models to evaluate the influence of vegetation covariates,

landscape-scale habitat characteristics and road density on the estimates of bird species

richness and the abundance (all birds, bark-gleaning and foliage-gleaning birds). For this

modelling, we selected a subset of vegetation variables deemed most likely to affect bird

assemblages (Table 3.1). Prior to modelling, we tested for collinearity among explanatory

variables using Pearson’s correlation coefficient and excluded one of each pair of variables

that had coefficients of correlation >|0.5| from the further analyses (Appendix B: Table B.8

for correlation matrix). All explanatory variables were standardised (mean = 0, standard

deviation = 1) to allow comparison of model parameter estimates.

Mixed-effects models are a robust statistical method for a nested study design (Pinheiro and

Bates, 2000; Zuur et al., 2010). In a hierarchically nested study, where data are clustered at

different spatial scales, there is a different error variance associated with each scale (Crawley

2007). As we collected data from sites (1 ha) embedded within larger landscapes (5 km x 5

km), fitting a standard regression may lead to poor model fit (Beck and Katzy 1995). We

therefore used generalized linear mixed models to handle random effects as well as non-

normally distributed data (Bolker et al. 2009). In this analysis, fixed effects included

vegetation variables, forest extent and road density. Because sites were clustered by 5 km x 5

km landscape, landscape identity was used as random factor in the mixed model, with sites

nested within landscape. The mixed-effects modelling was performed using the ‘‘lme4’’

package with Poisson error distributions (Bates et al. 2014) in R (R Core Team 2014).

64

Model averaging of the explanatory variables was then conducted for all response groups

using the package MuMIn (Bartoń 2012) in R statistical software (R Core Team, 2014). The

model averaging approach determines the strength of effects of the subset of explanatory

variables of species richness and abundance of each bird group (Burnham and Anderson

2002). Models were ranked according to their Akaike’s Information Criterion (AIC) value

and Akaike weight (ωі). The relative importance of all explanatory variables was calculated

by summing the Akaike weights (∑ωі) of variables across all the models where the variable

occurred. The larger the ∑ωі value the more important the variable (Burnham and Anderson

2002, Symonds and Moussalli 2011). To evaluate the goodness-of-fit of the model to the

data, we calculated conditional and marginal R2 values for the best models using the method

described by Nakagawa and Schielzeth (2013). The conditional R2 values show the

proportion of the variance explained by the global models (i.e. variance explained by fixed

and random factors), while the marginal R2 values show the proportion of the variance

explained by the fixed factors only (Nakagawa and Schielzeth 2013). Calculations for both R2

values were done using the ‘arm’ package in R statistical software (Gelman et al. 2012). We

tested for spatial autocorrelation by constructing spline correlograms of the model residuals

of full models for all response variables using functions “spline.correlog” in the “ncf” R

package to plot correlograms with 1000 permutations (Bjornstad 2013).

We used generalized linear models to investigate interactive effects between disturbance

intensity and forest extent measured within a 0.5 km and 2.5 km radii of the sampling site on

richness and abundance of all birds and bird within each foraging guilds. For this, we

classified study sites into two broader disturbance categories: lightly disturbed and heavily

disturbed (Appendix B: Table B. 7). To generate the classification, we used the mean value of

each of the three disturbance types and standardized the range to lie between 0 – 1. We then

took the average value for each site across all three disturbance factors. This composite

variable for each site was then used as an index of disturbance in the modelling of

interactions between different extents of forest cover and disturbance intensity. We ranked

the models based on their AIC values. We then compared the models by highest ranking AIC

value and calculated the Akaike weight the explanatory variables for each response variable.

65

3.4 Results

3.4.1 Vegetation and disturbance

Large tree density, large tree basal area and tree canopy cover were significantly lower in

heavily logged than lightly logged sites (large tree density: F2, 102 = 22.36, P < 0.001; large

tree basal area: F2, 102 = 9.70, P < 0.01); tree canopy cover: F2, 102 = 4.69, P < 0.05). Total

basal area (F1, 102 = 9.04, P < 0.01), percentage of shrub cover (F1, 102 = 7.94, P < 0.01), and

shrub density (F1, 102 = 4.02, P < 0.05) were significantly lower in heavily lopped sites

compared to lightly lopped sites. Grazing significantly reduced the ground herbaceous cover

(F1, 102 = 8.55, P < 0.01). The detailed results of effects of different disturbance types on

vegetation characteristics are presented in Appendix B: Table B.1 and B.2.

3.4.2 Bird communities

A total of 124 bird species was recorded across the 112 sites, comprising 68% local residents,

16% winter visitors, 2% summer visitors and 4% winter passage migrants. Overall mean

species richness was 19 ± 0.5 (±SE) and average abundance was 23 ± 0.69 (±SE) across all

sites. The overall species richness of birds in sites in heavily grazed and heavily lopped areas

was significantly lower (grazed: F1, 102 = 10.29, P < 0.001; lopped: F1, 102 = 27.6, P < 0.001)

(Fig. 3.2). The average abundance of birds (all species combined) was significantly lower in

heavily logged sites (F2, 102 = 3.5, P < 0.05). Heavy lopping had significant negative effects

on foliage-gleaning species richness and abundance (richness: F1, 102 = 7.13, P < 0.01; and

abundance: F1, 102 = 6.9, P < 0.01). Similarly, heavy logging adversely affected bark-gleaning

bird communities (richness: F2, 102 = 3.48, P < 0.05; abundance: F2, 102 = 1.46, P < 0.05)

(Appendix B: Table B.3).

66

Figure 3.2 Mean (± S.E.) species richness (dark bars) and average abundance (light bars) of

(a) all birds (b) bark-gleaning insectivores and (c) foliage-gleaning insectivores.

3.4.3 Effects of site- and landscape-level factors

The extent of forest cover within a 2.5 km radius of each survey site, large tree density, and

tree canopy cover had a strong influence on species richness and abundance for all bird

response groups (Fig. 3.3 and Appendix B: Table B.4). A high density of large trees was

found to be particularly important for bark-gleaning insectivores, while tree canopy cover and

shrub density positively influenced richness and abundance of foliage-gleaning insectivores

(Fig. 3.3).

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models for all explanatory variables.

There were between 6 and 9 models in the 95% confidence set (∑ωі = 0.95) for all response

groups (Appendix B: Table B.5). The goodness of fit statistics based on R2 values showed

that fit of the best models was sound for overall species richness (marginal R2 = 0.49 and

conditional R2 = 0.51) and total abundance (marginal R2 = 0.45 and conditional R2 = 0.66).

Model fit was also good for models of richness and abundance of bark-gleaning insectivores

(highest marginal R2 = 0.77 and 0.53 for richness and abundance; conditional R2 = 0.75 and

0.60 for richness and abundance) and richness and abundance of foliage-gleaning insectivores

(marginal R2 = 0.74 and 0.86 for richness and abundance; conditional R2 = 0.71 and 0.85 for

richness and abundance). Although both conditional and marginal R2 values were fairly

68

similar for all response groups, model performance was better for overall species richness and

abundance when the random factor was included in the global models. The extent of forest

cover within a 2.5 km radius of the survey site was the predictor with the highest rank

importance, and was positively correlated with all response groups. Similarly, tree canopy

cover and large tree density had a high rank importance and a positive relationship with all

response groups (Fig. 3.4).

69

Figure 3.4 Summed Akaike weights (∑ωі) of final subset of the explanatory variables for (a)

overall species richness, (b) average abundance, (c) bark-gleaner species richness (d) bark-

gleaner abundance (e) foliage-gleaner species richness, and (f) foliage-gleaner abundance.

3.4.4. Interactions between forest extent and disturbance intensity

The interaction between the extent of forest cover within a 2.5 km radius of each survey site

and disturbance intensity was significant in models of overall species richness (P < 0.01),

average abundance (P < 0.01), and abundance of bark-gleaning birds (P < 0.05) and foliage-

gleaning birds (P < 0.01). However, interaction effects of forest cover within a 0.5 km radius

70

of survey site were evident only for abundance of bark-gleaning and foliage-gleaning birds.

Interaction terms were included amongst the best models for all response groups (<2 ΔAIC)

(Appendix B: Table B.6). Thus, the effects of disturbance on bird communities at the site-

scale depended on the extent of forest cover in the landscape. The bird assemblages of more-

disturbed sites responded more strongly to the percent of forest cover in the landscapes than

did those of undisturbed sites (Fig. 3.5). Both bark-gleaning and foliage-gleaning species

responded more strongly to landscape-level forest cover in more-disturbed sites. For all two

response groups, the higher richness and/or abundance in less-disturbed sites was only

evident when there were low to moderate levels of forest cover in the landscape; in the most

forested landscapes, bird responses were similar among heavily-and lightly-disturbed sites.

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Figure 3.5 Relationship between overall bird species richness (a),overall bird abundance (b),

bark-gleaning abundance (c), and foliage-gleaning abundance and percent of forest cover in

2.5 km radius of survey site in landscape (heavily disturbed sites (filled circles and heavy

dashed line) and lightly disturbed sites (open circles and fine dashed line). For the purposes

of displaying the interaction effect, sites were divided into heavily and lightly disturbed based

on the value of the disturbance index (from three disturbance types).

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3.5 Discussion

Heavily-extracted sites had less-complex forest structure and depauperate avifaunal

communities, indicating potentially widespread deleterious effects of forest disturbances on

avifaunal communities in multi-use forests. However, while forest use practices significantly

affected the avifaunal community of sites, the intensity of these effects was dependent on

landscape context. Similarly, the effect of the extent of forest in the surrounding landscape

was weaker when the site was less-disturbed. This study therefore underscores the

importance of understanding the potentially interactive effects of disturbances at multiple

scales.

3.5.1 Effects of forest disturbances on bird communities

All three types of forest disturbance had deleterious effects on bird communities when they

occurred at higher intensities. Our results are largely consistent with past findings that

reported lower richness and abundance of birds in more disturbed sites (Peh et al. 2005,

Shahabuddin and Kumar 2007). Selective removal of large mature trees reduces habitat

suitability for many species of birds (Eyre et al. 2009, Touihri et al. 2014) that rely on them

for foraging and nesting (Sekercioglu 2002, Vergara and Marquet 2007). For example, bark-

gleaning insectivores are often strongly associated with large tree density (Cleary et al. 2007,

Greve et al. 2011, Dahal et al. 2014), and so are highly sensitive to habitat alteration (Adams

and Morrison 1993, Zurita and Bellocq 2012, Inskipp et al. 2013). Therefore, density of large

trees in a forest stand was the most important predictor of the distribution and abundance for

many species of birds at the site-scale.

Lopping and logging also had negative impacts on foliage-gleaning bird communities. These

species use the canopy layer for foraging, so removing parts of the canopy can have a

negative impact on species richness, abundance and composition. In this study, lightly lopped

and lightly logged sites had three times more foliage-gleaning insectivores than did heavily

lopped and heavily logged sites. Similar patterns of species distribution have also been noted

by Shahabuddin and Kumar (2006), who found that foliage-gleaning species such as great tit

Parus major, Hume’s warbler Phylloscopus humei and small minivet Pericrocotus

cinnamomeus were significantly more common where canopy cover was closed in an Indian

reserve. A recent study by Leal et al. (2013) in a Mediterranean region, found that foliage-

gleaning species such as great tit and chiffchaff Phylloscopus collybita were most affected by

canopy pruning activities.

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In this study, heavily grazed sites had only half the species richness of lightly grazed sites,

indicating a strongly detrimental effect of grazing practices on bird communities. Heavy

grazing significantly affects the population of many species of birds (Martin and McIntyre

2007). Grazing and trampling reduces understorey vegetation (Whitehorne et al. 2011, Piana

and Marsden 2014), thereby affecting understory-foraging species (Martin and McIntyre

2007) and ground-dwelling bird species (Maron and Lill 2005, Inskipp and Baral 2013).

Grazing can also have indirect effects on bird communities through changes in vegetation

characteristics such as nest site suitability and food supply (Dennis et al. 2008). For example,

in our study area, Abbot’s babbler Malacocincla abbotti, a terrestrial insectivore that nests

close to ground or in shrubs (Grimmett et al. 1999), was only recorded in lightly grazed sites,

presumably due to loss of understorey vegetation cover.

3.5.2 Moderating effects of landscape context

Although both site-and landscape-scale forest characteristics were important predictors in

determining species richness and the abundance, landscape characteristics had a consistently

positive and strong effect on all groups of birds. At the 19.6 km2 (2.5 km radius of survey

site) landscape scale, the forest extent had an important influence on bird species richness and

abundance. Strong effects of the proportion of habitat in landscapes surrounding sites have

been reported for forest breeding birds (e.g. Trzcinski et al. 1999) and woodland birds (e.g.

Maron et al. 2012). The scale at which landscape context is important varies with taxon and

habitat type. For example, reptiles respond most strongly to habitat context measured at 0.5

km2 scale (Bruton 2014), koalas at 1 km2 (McAlpine et al. 2006) and land birds (100 km2)

(e.g. Radford et al. 2005). However, in our study, the extent of forest within 0.79 km2 of a

site was less important, suggesting a greater role for landscape context at larger scales in

influencing the structure and composition of avian assemblages.

As we predicted, the impact of site-scale forest disturbance on bird communities depended on

the extent of forest cover within the surrounding landscape. Overall bird species richness,

average abundance and the abundance of bark-gleaners increased most rapidly with

landscape-level forest cover in the most heavily-disturbed sites, which indicates a higher

importance of forest extent in human-dominated landscapes where site-level habitat quality is

poor (Vergara and Armesto 2009). The benefits to the avifauna of less-disturbed sites were

strong at low to moderate levels of forest cover, but in landscapes with higher forest cover

(60% - 90%), the avifauna was less sensitive to disturbance. We found that disturbance

74

intensity was negatively related to the extent of forest cover (r = -0.67), with more intense

effects in lower-cover landscape. Therefore, our results suggest that the local-scale impact of

disturbances on bird communities may be moderated, at least partly, by maintaining a high

proportion of habitat surrounding such sites.

The response of bird communities to the interaction between landscape-level forest cover and

disturbance intensity varied depending on levels of species mobility. The differences may be

due to the different movement strategies that characterise each of the foraging guilds. For

example, many species of bark-gleaners (e.g. greater flameback Chrysocolaptes lucidus,

lesser yellownape Picus chlorolophus) have specialised niches and small ranges (Kumar et al.

2014), and this group responded strongly to the amount of forest cover within 0.5 km in the

highly disturbed sites. This indicates that forest cover close to disturbed sites may buffer

negative effects of disturbance by offering different habitat resources. For highly mobile

birds, forest cover within this small extent may be less important, as they range over much

larger areas. Furthermore, the highly mobile and widespread generalists such as sallying

insectivores (e.g. collared falconet Microhierax caerulescens, black drongo Dicrurus

macrocercus) had little response to the extent of forest cover in the landscape at either scale.

Landscapes with more forest cover support a larger species pool (Radford and Bennett 2007,

Haslem and Bennett 2008, Taylor et al. 2012) through both sampling effects, and because a

greater extent of forest cover in landscapes offers habitat diversity (Radford et al. 2005,

Maron et al. 2012), and can serve source habitats for range of species (Pulliam 1988). Birds

are a mobile taxon, and their presence at a particular site does not necessarily mean they are

resident, nor that they solely use the resources within that site. Thus, complementary

resources may be more widely distributed within the occupied landscape. More habitat within

the surrounding landscape increases the chance that suitable refugial or complementary

habitat exists (Dunning et al. 1992), increasing the likelihood of occupancy within the

landscape and thus the chance of detection at a site, even a degraded one. Therefore,

maintaining more forest cover has important ecological consequences for the ability of a

wide-range of avifaunal species to persist.

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Our results have important management implications in terms of sustainable forest

management and biodiversity conservation at the landscape scale. Our study demonstrates the

maintenance of larger areas of mature forest in the landscape should be a high conservation

priority for bird conservation in highly-disturbed landscapes. This can be achieved with: a)

appropriate conservation and restoration of degraded landscapes particularly for those forests

that are heavily degraded such as most of the state forests and b) maintenance of existing

forest cover in protected areas as protected areas are important natural habitats in the region.

Similarly, it is also important to reduce human pressures on forests to maintain vegetation

complexity at the site-scale. Since most lowland landscapes in Nepal are multiple-use and are

subject to a high degree of anthropogenic pressures, the development and implementation of

sustainable forest management plans are urgent to prevent further degradation of habitat and

their avifauna in lowland landscapes of Nepal.

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CHAPTER 4

RELATIONSHIPS BETWEEN LANDSCAPE-LEVEL SPECIES RICHNESS AND

FOREST EXTENT VARY AMONG BIRD

GUILDS

Plate 4: A patch of native forest in Chitwan National Park in lowland Nepal.

Submitted to Journal of Applied Ecology as: Dahal, B. R., C. A. McAlpine, and M. Maron.

2015. ‘Effects of landscape characteristics and habitat extent on bird communities in lowland

Nepal’.

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4.1 Abstract

Anthropogenic habitat modification has dramatically altered the spatial patterning of different

habitats, which affects the richness of species that landscapes support. A particular focus of

research has been thresholds in extent of preferred habitat, below which richness declines

rapidly. However, there is likely to be variation among functional groups in the strength and

shape of the relationship between the extent of habitat and landscape-level species richness. I

surveyed birds across 28 landscapes (each 5 × 5 km) that differed in the extent of remnant

forest in southern lowland Nepal. The estimated species richness of all birds and bird within

several foraging guilds at the landscape-level were modelled as a function of forest extent in

the study landscapes, and factors that potentially modify the relationship (such as degree of

disturbance) were explored. The landscape-level species richness of forest birds was

consistently positively related to the extent of forest cover in the landscape. However, the

strength of the relationship varied substantially among foraging guilds, with effects strongest

for foliage-gleaning and frugivores. As with previous studies, species richness increased most

steeply with forest extent in less-forested landscapes, but my findings differed in that richness

continued to increase as forest extent approached 100%. The strongest effects of forest cover

on overall bird richness occurred in landscapes with a greater extent of disturbance.

Managing and restoring forests to maintain forest extent, particularly in more-degraded

landscapes, should be a key strategy for landscape-level conservation of birds in the region.

Keywords: Forest extent, forest birds, habitat thresholds, habitat disturbance, landscape

change, Nepal

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4.2 Introduction

Anthropogenic land-cover change has dramatically altered the spatial patterning of different

habitats (DeFries et al. 2004, Foley et al. 2005) which in turn has affected the distribution and

abundance of many species. More than 50% of the world’s animal species have declined in

population size over the last four decades due principally to the loss and degradation of

natural landscapes (WWF 2014). As pressure from human land-uses continues and landscape

patterns and ecological processes are disrupted (Franklin et al. 2002, Fischer and

Lindenmayer 2007), understanding how these novel landscape patterns affects species

persistence at large spatial scales is increasingly important for biodiversity conservation

(Fahrig 2003, Bennett et al. 2006).

The total extent of forest cover is a key driver of the occurrence and abundance of species,

ultimately affecting the total number of species a landscape can support. Landscapes with

more habitat support a larger species pool (McGarigal and McComb 1995, Trzcinski et al.

1999, Fahrig 2003, Maron et al. 2012) through both sampling effects (Wiens 1992, Whittaker

and Fernández-Palacios 2007), and because habitat extent correlates with habitat diversity

(Radford et al. 2005, Kallimanis et al. 2008). Different types of habitat across the landscapes

can be refugial or complementary for many species of fauna. For example, habitat diversity

can provide critical complementary resources for different activities such as breeding,

foraging, and nesting, allowing persistence of more species in the landscape (Dunning et al.

1992, Law and Dickman 1998). Similarly, certain citical habitats and habitat features (e.g.

riparian forest, large mature trees) required for particular species are more likely to occurr in

landscapes with more habitat. Therefore, the larger extent of habitat is likey to offer both

primary and complementary habitats for the peristence of faunal communties within a

landscape.

Empirical studies suggest that sites in landscapes with more forest can have higher densities

of reptiles (McAlpine et al. 2015), greater species occurrence and abundance of birds (e.g.

Villard et al. 1999, Mortelliti et al. 2010, Martensen et al. 2011, Taylor et al. 2012), and

greater richness of small mammals (e.g. McAlpine et al. 2006, Estavillo et al. 2013). In

particular, studies of both individual species (e.g. Betts et al. 2007, Suarez-Rubio et al. 2013)

and landscape-level species richness of birds (e.g. Radford et al. 2005, Hu et al. 2012, Maron

et al. 2012) have reported non-linear responses to forest extent. A particular focus has been

habitat extent thresholds, whereby if forest extent falls below the threshold, sharp declines in

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population sizes and species diversity occur within that landscape (Andren 1994). Radford et

al. (2005) showed that steep declines of landscape-level species richness of woodland birds

below a threshold of 10% remnant habitat cover, and Maron et al. (2012) also reported

nonlinearities in this relationship. Since such thresholds would have important consequences

for species persistence in modified landscapes (Ewers and Didham 2006), understanding

where and for which species groups such thresholds exist is key to developing conservation

and restoration strategies (Lindenmayer and Luck 2005, Huggett 2005).

Although several studies have now reported a nonlinear relationship between landscape-level

species richness and extent of suitable habitat (Taylor et al. 2012, Maron et al. 2012), the

shape and magnitude of bird species response to forest cover may not be universal across the

different foraging guilds. It is likely that different groups of species vary in their response to

forest cover in the landscape. For example, frugivore richness might be strongly driven by the

variety of fruiting plants which diminishes as habitat loss proceeds, where as raptorial

carnivores with more generalist diets might be less-affected. Thus, strongly nonlinear

relationships with forest cover might be driven by particular groups of species that are most

sensitive to forest loss. However, variation in the nature of the relationship between richness

of different foraging guilds of birds and the extent of forest habitat in the landscapes remains

unexplored.

Further, both species persistence and the effect of forest cover are likely to be influenced by

other landscape characteristics, such as levels of anthropogenic disturbance. Many types of

forest disturbance, such as logging and livestock grazing, negatively affect species

persistence at the site-level (Shahabuddin and Kumar 2007) and such relationships could

cloud understanding of how species respond to loss of forest cover. Thus, the relationship

between landscape-level species richness and forest cover may be altered by interactions

among landscape characteristics. Knowledge of such interactions remains limited.

Here I investigate relationships between landscape characteristics and landscape-level species

richness of forest birds across lowland Nepal. Firstly, I examine whether the effects of

remnant forest extent in a landscape on the species richness of birds in that landscape are

consistent across foraging guilds, or if a particular guild contributes disproportionately to

observed patterns between species and forest cover in the landscape. Secondly, I determine

whether non-linear relationships exist between landscape-level species richness of birds and

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forest cover, and examine how different landscape characteristics influence these

relationships. Finally, I assess the existence of interaction effects between disturbance

intensity and landscape-level forest extent on species richness of birds.

4.3 Material and Methods

4.3.1 Study area

The study was conducted in southern Nepal, in the region called ‘Terai’ (80° 4’ 30” to 88°

10’ 19” E 26° 21’ 53” to 29° 7’ 43” N, elevation 63 - 330 m ASL). The Terai encompasses

most of the country’s tropical moist forest from the Mechi River in the east to the Narayani

River in the centre. The mean annual rainfall decreases from 2,680 mm to 1,138 mm from

east to west, and the mean monthly rainfall ranges from 8 mm in November to 535 mm in

July (FRA/DFRS 2014). The area is characterized by a tropical climate, with the maximum

monthly mean temperature of 35-40°C in April/May and the minimum, 14-16 °C, in January

(Jackson et al. 1994). Before 1950, the region supported continuous dense tropical forest.

With the eradication of malaria in the early 1950s, large tracts of the highly productive

lowland forests were converted to agriculture (Hrabovszky and Miyan 1987). Consequently,

most of the forest was destroyed and the remaining forest areas were subjected to intense

human exploitation. Nearly half of Nepal’s population now lives in the 17% of the country

that is lowland (Central Bureau of Statistics 2011).

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Figure 4.1 The study landscapes in lowland Nepal (grey shading indicates forest cover) and

histogram showing the different extent of forest cover in landscape.

4.3.2 Study design

Twenty-eight landscapes, each 5 km x 5 km, and supporting different amounts of forest cover

(7.9%–95.3%), were selected across south-central (Bara-Parsa forest and Chitwan forest) and

south-eastern lowland forests (eastern forests) (Fig. 4.1). Geographically, 15 landscapes were

located in eastern Terai forests and 13 landscapes were located in central Terai forests. Four

survey sites, each measuring 200 m x 50 m, were randomly located in each landscape,

resulting in a total of 112 sites (28 landscapes x 4 sites). All sites were located at least 500 m

from roads to minimize any road-induced variation in bird assemblages. The minimum

distance between sites was at least 1000 m to reduce the chance of spatial dependence.

4.3.3 Bird surveys

At each study site, birds were surveyed on three occasions between November 2012 and May

2013On each visit, the observer (BRD) recorded all birds seen or heard within 25 m of the

centreline of the transect while walking along its length over a 10-min period. Surveys were

conducted only between 0600 and 1100 h in the morning and 1400 to 1745 h in the afternoon.

Although the effects of time of day on bird observation were not tested prior to the actual

field survey, several other studies have reported that the detection rate of most bird species is

greater in morning (Bried et al. 2011) with another peak in activity in the late afternoon, 2–3

h before sunset (Kessler and Milne 1982). In general, birds tend to reduce activity during the

midday heat (Pizo et al. 1997). To avoid possible bias, we standardized the survey protocol in

such a way that, although not all individual sites had an afternoon survey, afternoon surveys

occurred equally among site types and landscapes. All surveys were conducted during fair

weather when there was no rain and the wind speed was low.

4.3.4 Landscape variables

The total area of forest and water (river, permanent lakes) in each study landscape was

calculated using GIS (using ArcGIS 9.3). I measured the total area of forest and open water in

ha/km2 from land cover data of the region (WWF 2005). The land-cover data were made

available by WWF Nepal. Mean rainfall (mm/yr) data for each landscape was obtained from

the closest weather station (Department of Hydrology and Meteorology, Nepal). The total

numbers of trees were counted within four 20 m x 20 m quadrats to measure the species

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richness of trees at each site. Similarly, disturbance intensity was assessed at each site using

four 20 m x 20 m randomly located quadrats. In each quadrat, indicators of disturbance due to

forest-use practices were recorded to reflect the intensity of disturbances in terms of grazing,

fodder collection and fuel wood extraction (Shahabuddin and Kumar 2007). These included

the proportion of trees showing signs of lopping, number of cut stumps and number of

livestock dung piles. The density of livestock dung piles indicates the degree of usage of

forest habitat by grazing livestock, stump density reflects logging intensity, while signs of

lopping indicate the amount of fodder and fuelwood extraction for each site. The values of

each disturbance variable across the four quadrats were averaged for each site. The mean

value of lopping (1.7 ± 0.4), grazing (2.0 ± 0.4) and logging (0.5 ± 0.1) in the lightly

disturbed sites were significantly lower than mean values of lopping (46.5 ± 3.4), grazing

(12.5 ± 0.9) and logging (11.1± 0.9) in heavily disturbed sites. The values from the four

survey sites per landscape were then averaged to derive an average disturbance value. In

addition, to display the interactions, I classified the landscapes into lightly disturbed and

heavily disturbed landscapes. Landscapes with ≤ 5 disturbance index on average were

categorized as lightly disturbed, while ≥ 5 as heavily disturbed (Appendix B: Table B.7).

4.3.5 Data analysis

Bird survey data were pooled across all sites within each study landscape. Membership of

foraging guilds was identified based on primary habitat specialization and diet information

compiled from Ali and Reply (1987) and Grimmett et al. (2009). Although several foraging

guilds were identified, only those foraging guilds which had more than five species present in

each landscape were included in the analyses. These include: foliage-gleaning insectivores,

sallying insectivores and frugivores. To account for variation in sample completeness among

landscapes, I calculated the estimated species richness for each landscape for all birds and

different foraging guilds using the nonparametric species richness estimator Chao2 in the

programme Estimates 9.1.0 (Colwell 2013).

I fitted a series of regression models to visualize the relationships between estimated species

richness (Chao2) of each species group and forest cover. I compared the performance of

linear, exponential, loess and discontinuous piecewise regressions using the Akaike

Information Criterion (AIC) and adjusted R2 values. The loess (locally weighted, non-

parametric regression) is useful for examining nonlinear relationships (Cleveland and Devlin

1988, Jacoby 2000). For the piecewise regression model, we used the “segmented” package

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in R (Muggeo 2012) to identify the break-point from the data. When using the segmented

package to test for break-points, an initial estimate of the break-point from the data is

required as a starting estimate, which we tested for a priori using the Devies test (Davies

1987, Muggeo 2008).

I also modelled the relationship between landscape-level bird species richness and all

landscape variables using GLMs with a Poisson distribution in the lme4 package v.1.1-5 in R

(R Core Team 2014). All explanatory variables were standardised (mean = 0, standard

deviation = 1) to allow comparison of model parameter estimates (Burnham and Anderson

2002). Model averaging of the explanatory variables was then conducted for all response

groups using the package MuMIn (Bartoń 2012) in R statistical software (R Core Team,

2014). The model averaging approach determines the strength of effects of the subset of

explanatory variables of species richness and abundance of each bird group (Burnham and

Anderson 2002). Models were ranked according to their AIC value and Akaike weight (ωі).

A 95% confidence set of models, used for model averaging, was constructed by starting with

the model with the highest Akaike weight and repeatedly adding the model with the next

highest weight until the cumulative sum of weights exceeded 0.95 (Burnham and Anderson

2002). The Akaike weights in which each predictor variable occurred were summed; the

larger the ∑ωі value the more important the variable (Burnham and Anderson 2002, Symonds

and Moussalli 2011). As an indication of goodness-of-fit, I calculated R2 values for the global

models in R statistical environment (R Core Team 2014).

4.4 Results

4.4.1 Relationships between landscape-level species richness and forest cover

The landscape-level species richness of all birds increased with the increase of forest extent

in the landscape, with 42.1 % of the variance explained by the best model. However, the

strength of the relationship varied substantially among the foraging guilds. Sallying

insectivores were only weakly associated with forest cover, but foliage-gleaning insectivore

and frugivore richness were strongly related to forest extent, with 63.2 % and 36.2 % of the

variance explained by the best models, respectively. The shape of the response to forest cover

also varied among foraging guilds (Fig. 4.2). There was limited evidence to support the

threshold models. The break points (thresholds) as a function of forest cover ranged from 8.3

% of forest cover (sallying insectivores) to 14.1 % (foliage-gleaning insectivores). However,

inspection of the loess models for overall bird richness and foliage-gleaning species richness

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suggested initial steep increases with forest cover at low levels of cover, before a less-steep,

but continuing, increase in landscapes beyond 20-30% forest cover. This was most

pronounced for richness of foliage-gleaning species.

Figure 4.2 Models of the relationship between estimated species richness and forest extent in

each landscape. AIC: Akaike information criterion for each model

4.4.2 Relative importance of landscape variables

The strong positive effect of forest area on estimated species richness was evident from the

model averaging, with the summed Akaike weight revealing that forest cover was the most

influential parameter in each of the models for all response groups (Fig.4.4). The importance

of landscape-level forest cover was greatest for all species (∑ωі = 1.0) and foliage-gleaning

insectivores (∑ωі = 1.0) and weakest for sallying insectivores (∑ωі = 0.33). Other important

landscape characteristics included the extent of water body, which was the most reliable

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predictor for frugivorous and all species, while annual rainfall was an important influence on

foliage-gleaning insectivores. However, effects of all parameters were weak for sallying

insectivores (Fig. 4.3).


models for all explanatory variables for each of: (a) all species, (b) foliage-gleaning

insectivore, (c) frugivore and (d) sallying insectivore.

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Figure 4.4 Summed Akaike weights (∑ωі) from model averaging of the environmental

variables for landscape-level richness of (a) all species, (b) foliage-gleaning insectivores, (c)

Frugivores, and (d) sallying insectivores.

4.4.3 Interactions between landscape characteristics

The interaction between the extent of forest cover and disturbance intensity was significant in

models of estimated richness of all species. The bird assemblages of less-disturbed

landscapes responded more weakly to the extent of forest cover in the landscape than did

those of more-disturbed landscapes. This interaction effect was not evident when individual

guilds were considered. However, for foliage-gleaning insectivores, rainfall interacted

significantly with forest cover, such that the response of foliage-gleaning insectivores to the

extent of forest was weaker in higher-rainfall landscapes (Fig. 4.5).

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Figure 4.5 Relationship between overall estimated richness and percent of forest cover (a) in

more-disturbed landscapes (filled circles and full line) and less-disturbed landscapes (open

circles and heavy dashed line) and (b) higher-rainfall landscapes (filled circles and full line)

and lower-rainfall landscapes (open circles and heavy dashed line).

4.5 Discussion

Strong relationships between landscape-level forest extent and estimated species richness of

birds were evident, and these relationships varied among foraging guilds, with richness of

foliage-gleaning species contributing most strongly to the positive relationship. Further, the

effect of landscape-level forest extent on bird richness was influenced by disturbance levels

and rainfall. The relationship between estimated species richness and the extent of forest

cover in the landscape was nonlinear, supporting the hypothesis that the relationship between

forest extent and richness is steepest at low levels of forest extent significantly reduces.

However, as the relationship between richness and forest extent varied among foraging guilds

and with landscape characteristics, generalizing such relationships may mask important

elements of the consequences of landscape change.

4.5.1 Species richness and forest cover

Most studies of avian responses to landscape change have focused on response variables

measured at the site or patch-level, using characteristics of the site or the landscape context

surrounding site as explanatory variables (McGarigal and Cushman 2002, Guenette and

Villard 2005). However, response variables measured at small scales cannot necessarily

reveal emergent properties of whole landscapes (Bennett et al. 2006). More recently, key

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empirical studies (e.g. Radford and Bennett 2007, Pardini et al. 2010, Maron et al. 2012,

Taylor et al. 2012, Bennett et al. 2014, Ochoa‐Quintero et al. 2015) have begun to focus on

landscape-scale sampling, examining how the assemblages of entire landscapes respond to

landscape characteristics in Australia. My results are consistent with several of these recent

studies for birds that showed strong positive relationships between species richness and the

extent of forest cover in the landscape (Radford et al. 2005, Maron et al. 2012,

Ochoa‐Quintero et al. 2015). However, there was variation among foraging guilds, indicating

that not all groups of species respond in the same way to forest cover. The richness of

foliage-gleaning insectivores and, to a lesser extent, frugivores, was responsible for the

pattern, whereas richness of sallying insectivores were only weakly related to forest cover.

The strong positive relationships between foliage-gleaning insectivores and frugivores and

forest cover in the landscape may be due to increased diversity of trees and therefore foraging

substrates and fruit sources in more-forested landscape. Haddad et al. (2009), for example,

revealed that the species richness and abundance of arthropods was positively related to plant

species richness. In my study landscapes, forest cover was positively related to the richness of

tree species recorded in the transects (r = 0.54) and tree richness was also related to richness

of foliage-gleaners (r = 0.31) and frugivores (r = 0.19). A similar positive correlation—

although at a site level—between tree diversity and species richness of avian insectivores and

frugivores birds was found in agricultural landscapes (Harvey et al. 2006). Thus, a greater

variety of potential food sources is more likely to support a correspondingly large suite of

species, whereas diversity of food sources for the more generalist group of aerial insectivores

may be less likely to relate to forest extent.

The relationship between species richness and extent of forest cover was somewhat non-

linear, with species richness decreasing more steeply below about 20-30% forest cover in the

landscape. My findings are largely consistent with the results of other studies that have

shown non-linearity in the relationship between species richness and habitat extent (Radford

et al. 2005, Maron et al. 2012). However, in contrast to the findings of these studies, my

results also showed continued, but less-steep, increase of species richness above the

threshold. I also found that the observed responses of birds to forest loss were not consistent

across the forest bird community. Instead, several underlying factors such as landscape

productivity, vegetation or soil type (Maron et al. 2012) and species sensitivity to forest cover

change (Martensen et al. 2012) can affect the response of species richness to forest cover.

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Considering total species richness alone could potentially obscure variation in relationships at

the functional level of birds, particularly when forest cover is confounded with underlying

factors (Maron et al. 2012).

4.5.2 Relative importance of landscape characteristics

In addition to forest extent, several other factors affected the species richness of forest birds

in the landscape. Annual rainfall and extent of water were also positively related to

landscape-level species richness of birds. Richness of all birds and of frugivores both

increased with the extent of water in the landscape. This association is likely to be driven by

differences in primary productivity and resource availability in riparian habitats. Riparian

habitats generally support distinct vegetation (Palmer and Bennett 2006) and provide more

resources particularly emergent aquatic and aerial terrestrial insects for many insectivores

birds (Iwata et al. 2003). A higher species richness of birds in riparian habitats has been

reported in other studies (Knopf and Samson 1994, Schneider and Griesser 2009, Bennett et

al. 2014). Furthermore, the presence of permanent water bodies increases landscape

heterogeneity, generating habitat diversity (Tews et al. 2004). Habitat diversity in the

landscape is also important for persistence of mobile species which can require different

habitats for their survival (Saunders 1990, Law and Dickman 1998). Foliage-gleaning

insectivore richness was also higher in landscapes with greater annual rainfall. Rainfall

influences arthropod abundance and diversity in forest landscapes (Sofaer et al. 2012) and

increases carrying capacity for insectivorous birds (Williams and Middleton 2008).

4.5.3 Interactions between landscape characteristics

My study revealed that the effects of extent of forest cover on bird species richness depended

on the degree of disturbances in the landscapes. Positive effects of forest extent on bird

communities in the landscape are often reported (e.g. Andren 1994, Cushman and McGarigal

2003) but the way that intensity of disturbances (livestock grazing and logging) interacts with

landscape-level forest extent to drive landscape-level richness has not previously been

examined. The strongest effect of forest cover on all species richness occurred in landscapes

with a greater extent of disturbance, indicating a role of forest extent in moderating the

effects of disturbance. Accordingly, disturbance effects on avifauna of less-disturbed

landscapes were strongest at low to moderate levels of forest cover, but in landscapes with

greater forest cover (60% – 90%), the avifauna was less affected by disturbance. The species

richness of foliage-gleaning insectivores across the landscape was also dependent on the

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interaction between rainfall and forest cover, suggesting that the extent of forest cover has

less effect on foliage-gleaning insectivores in higher-rainfall landscapes, which are likely to

be more stable in food availability (Williams and Middleton 2008).


This study provides strong evidence of positive association of birds with the extent of forest

cover at the landscape level, and therefore has important management implications. The non-

linearity of relationships between species richness and forest cover suggest that species

richness increases more rapidly with forest cover at low (< 20%) levels of cover, but the

positive relationship between richness and forest cover is still evident at high levels of cover.

Nepal’s lowland forest landscapes face a high risk of collapse in avifaunal richness if further

forest loss and modification occur, particularly in landscapes of lower vegetation cover. For

example, the state managed forests in Parsa and Eastern landscapes are particularly

vulnerable due to low forest cover. Furthermore, logging, over-extraction of forest resources

and cattle grazing are ongoing disturbances in these forests. It is therefore important to place

a regulatory mechanism that helps reduce human pressure and maintain vegetation

complexity at both site-and landscape-scales. As more than 70% of the country’s forest bird

species (of which more than 50% are nationally threatened) inhabit the lowland forests

(Inskipp et al. 2013, Baral et al. 2014), maintenance of existing forest cover and interventions

to restore degraded habitats to prevent further loss of avifaunal populations in lowland

landscape of Nepal are critical.

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CHAPTER 5

SYNTHESIS AND RECOMMENDATIONS

Plate 5: Agricultural intensification in adjacent to the protected area, lowland Nepal.

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5.1 Overview

The conservation of faunal communities in human-dominated landscapes is a challenging

task, particularly in developing countries. About 1.6 billion people living in poverty depend

on forests for their livelihoods (World Resource Institute 2005, Chao 2012) and this

dependency is likely to increase in the future. This burgeoning demand for food and fibres

puts enormous pressure on remnant forests, and has significantly contributed to degrading

landscapes and changing habitats for both fauna and flora. Thus, effective conservation of

avifaunal assemblages in multiple-use landscapes, for example, requires an understanding of

how habitat characteristics at individual sites and also across landscapes affect species

composition and persistence (Ewers and Didham 2006, Lindenmayer and Fischer 2006).

In this thesis, I investigated the effects of habitat characteristics on forest bird assemblages at

multiple spatial scales in order to draw inferences for conservation of avifauna in the lowland

Terai forest of Nepal. In Chapter 2, I examined relationships between site characteristics and

bird richness and abundance, with both response and predictor variables measured at the

same scale. This site-level study was used to compare the conservation value for birds of

differently-managed forests. In Chapter 3, I examined the relationships between forest bird

assemblages and both site and landscape characteristics, including the extent of forest within

both a 500 m and a 2500 m radius of survey sites. Here, I hypothesized that the occurrence of

species and assemblages depends not only on the properties of sites at which birds were

sampled, but also on the proportion of forest in the landscape, and its interaction with site-

level factors. In Chapter 4, I extended my research approach beyond the site/landscape

context, and adopted a whole-of-landscape approach in which both the response variables and

predictor variables were measured at the scale of the whole landscape (Bennett et al. 2006).

Such an approach is useful in understanding the influence of emergent properties of entire

landscapes on faunal assemblages (Mortelliti et al. 2010, Maron et al. 2012, Taylor et al.

2012).

In this final chapter, I summarize the main outcomes of my research in relation to the

questions I posed, and discuss the implications for the management of human-dominated

landscapes for avifaunal conservation in the lowland Terai forests of Nepal. I also discuss the

limitations to this study and propose directions for future research.

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5.2 Off- reserve forests provide complementary habitats for bird conservation

In recent years, forest outside of formal conservation reserves has been increasingly

recognized for its potential role in conservation of biota (Persha et al. 2010, Porter-Bolland et

al. 2012, Baral et al. 2014). As existing protected area coverage is often biased in terms of the

species and habitats that are protected (Tewksbury et al. 2002, Hoekstra et al. 2005), the

conservation of off-reserve forests can provide complementary habitats that help to support a

larger suite of species across the landscape. Such complementary habitats in the landscape

provide critical resources, particularly for species that require a variety of habitats for their

persistence (Dunning et al. 1992, Law and Dickman 1998). Thus, off-reserve forests can be

critical to maximizing representation of biodiversity features in a landscape. It is therefore

important to understand conservation values of off-reserve forests and how such spatially-

distributed forest habitats complement existing protected area networks in achieving


In Chapter 2, I investigated whether off-reserve forests (state-managed and community-

managed forests) support bird assemblages that complement those of protected areas. I

compared the habitat attributes and bird assemblages among sites in each of these three forest

tenure categories. Protected area sites had the greatest richness and diversity of birds

compared to sites in community forests and state forests. They also had significantly greater

species richness and diversity of forest specialists and bark-gleaning insectivores. However,

off-reserve forests supported bird assemblages that complement those of protected areas.

Many species of birds that were not recorded in sites in protected areas were recorded in sites

in off-reserve forests. Only 45% of species detected were common to all three forest

management tenures. This distinctness of bird species in the sites in off-reserve forests

contributes to maintenance of species diversity across landscapes.

Habitat features such as tree canopy cover and large tree density were similar between sites in

community-managed forests and those of protected areas. This indicates that aspects of

habitat condition in sites in community forest are relatively good, and potentially provide

critical resources for many species of birds at levels similar to protected areas. Furthermore,

as off-reserve forests provide structural links among different critical resources in the

landscape, species of birds that exploit a variety of habitats, for example, frugivores that must

follow the shifting pattern of fruit availability over time, can benefit. Such connectivity is of

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primary importance to the distribution and abundance of biota (Lindenmayer et al. 2008,

Chisholm et al. 2011).

Having a complementary habitat resource within a landscape is likely to increase beta

diversity. Beta diversity (species differentiation across sites, also referred as spatial turnover)

has important implications for conservation planning (e.g. Condit et al. 2002, Wiersma and

Urban 2005). In this study, I found that beta diversity (i.e. species turnover) among sites in

off-reserve forest types was higher than among sites in protected areas (Chapter 2). Thus,

although alpha diversity is lower in off-reserve forests, richness at larger scales is likely to be

relatively similar.

However, these forests outside the protected areas, in particular state-managed forests, are

subject to heavy anthropogenic pressure for subsistence livelihood activities. As such,

anthropogenic activities have detrimental impacts on forest structure and associated avifauna

(Chapter 3), these state-owned forests may not necessarily support suitable habitats for long-

term persistence of populations of several species including forest specialists. Thus, Chapter

2 concludes that including state-owned forest into appropriate conservation measures such as

habitat restoration and preventing over exploitation to maximize regional avifaunal diversity.

5.3 Forest use practices can have detrimental effects on vegetation and associated birds

Forest use practices such as logging, livestock grazing, and lopping of tree branches for

fodder and fuelwood are the major forms of forest disturbance in multiple-use landscapes.

These extractive activities, mainly for subsistence, may significantly change the forest

structure and diversity (Sagar and Singh 2004, Kumar and Shahabuddin 2005). For example,

livestock grazing tends to result in changes to species composition and structure in the

understorey (Tasker and Bradstock 2006, Whitehorne et al. 2011), and logging for timber

production and fuelwood and lopping for fodder and fuelwood simplifies the stand structure

(Shahabuddin and Kumar 2007, Thapa and Chapman 2010). However, effects of subsistence

forest disturbance on bird assemblages have received little attention despite the fact that such

information is essential for effective conservation planning for anthropogenically-disturbed

landscapes.

In this study, I found that logging, grazing and lopping activities had deleterious effects on

multiple aspects of habitat condition (Chapter 3). As expected, the density and basal area of

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large trees and tree canopy cover were significantly lower in heavily-logged sites than in

lightly-logged sites. Lopping activities also significantly affected shrub density and shrub

cover, while grazing activities reduced herbaceous cover. Studies have shown negative

effects of logging (e.g.Moktan et al. 2009, Sapkota et al. 2010) and grazing (e.g. Whitehorne

et al. 2011) on vegetation structure; this study also underscores negative lopping activities as

substantial drivers of simplified vegetation structure. In lowland Terai forests, lopping that

involves removal of tree branches 5 – 20 cm diameter is widely practiced usually for fodder

and fuelwood (Chapter 3, Thapa and Chapman 2010). Thus, I concluded that as the

livelihoods of the local population partly depend on extraction of adjacent forest resources,

these subsistence activities are reducing the stand-scale habitat complexity in the lowland

Terai forests.

Mirroring the results for vegetation structure, species richness and abundance of birds were

also negatively affected by the intensity of logging, lopping and grazing practices in the

lowland landscape. The overall species richness of birds in sites in heavily grazed and heavily

lopped areas was significantly lower than in lightly grazed and lightly lopped sites. Similarly,

the average abundance of birds (all species combined) was significantly lower in heavily

logged sites than lightly logged sites. I found that the effects of forest disturbances on forest

birds varied with foraging guild (Chapter 3). For example, the average species richness and

abundance of bark-gleaning insectivores were significantly lower in heavily logged sites,

indicating that this group of birds was strongly affected by the removal of large trees.

Likewise, the average species richness and abundance of foliage-gleaning insectivores were

significantly lower in heavily lopped sites. This shows that lopping activities that involve the

removal of tree branches for fuelwood and the canopy layer for fodder heavily affected

foliage-gleaning insectivores (Chapter 3, Shahabuddin and Kumar 2006, Leal et al. 2013).

Similarly, birds that forage on the canopy layer were affected by excessive pruning of mid-

and upper-storey vegetation, probably because of the loss of foraging substrates. I concluded

that the several species including forest-specialist species tended to be more severely affected

by forest disturbance.

In this study, I found that overall richness and abundance of birds positively related to the

large tree density, tree canopy cover and shrub density. However, the relationship between

bird species and local-scale habitat characteristics differed among foraging guilds (Chapter 2,

Chapter 3). For example, species of birds such as bark-gleaning and foliage-gleaning

96

insectivores that require large trees for their feeding and nesting were strongly related to the

density of large tress and tree canopy cover for their survival (Chapter 2, Chapter 3, Adams

and Morrison 1993, Laiolo et al. 2004, Galitsky and Lawler 2015). The tree canopy cover and

shrub density was positively influenced the richness and abundance of foliage-gleaning

insectivores (Chapter 3).

As I outlined in Chapter 3, excessive resource extraction is the major cause of habitat

degradation in the lowland Terai forests. As richness and abundance of forest birds are

related to local-scale habitat structure (e.g Chettri et al. 2002, Guenette and Villard 2005), the

simplification of habitat as a result of forest disturbance is likely to affect the bird

assemblages within a site (Cleary et al. 2007, Maron and Kennedy 2007, Greve et al. 2011,

Chapter 3). However, in this study, I found that the effects of forest disturbance on birds were

not restricted only to the site level, as higher intensities of disturbance also negatively

affected the landscape-level species richness of birds (Chapter 4). This may have a significant

effect on the population of many threatened species which are already in decline in the region

(Inskipp et al. 2013). Simplification of habitat structure further threatens the last remaining

population of species such as the blue–eared barbet Megalaima australis, Abbott’s babbler

Malacocincla abbotti and the greater flameback Chrysocolaptes lucidus in the lowland Terai

forests. These bird species were not recorded in sites in heavily-disturbed areas during my

survey period. It is therefore critical to introduce effective conservation measures that help

reduce the current rate of exploitation of forest resources in the lowland Terai forests.

5.4 Effects of forest-use practices on bird assemblages vary with the landscape context

A large body of research has focused on the effects of site-scale habitat characteristics on

species richness and abundance (Chapter 2, Moktan et al. 2009, Khanaposhtani et al. 2012).

These studies demonstrated the importance of local habitat features for species and

abundance of bird communities. However, in recent years, increasing numbers of studies

have shown that species occurrence at a particular site depends not only on site

characteristics, but also on characteristics of the landscape in which the site is located

(Radford et al. 2005, Haslem and Bennett 2008, Döbert et al. 2014). In this thesis (Chapter 3),

I therefore examined the relative importance of site-and landscape characteristics on forest

birds in the lowland Terai forest of Nepal, and whether the effects of forest use practices on

the forest bird community depend on the extent of forest cover surrounding the sites.

97

As outlined in 5.3 above, the structural features of forest stands such as canopy structure, tree

sizes, and shrub density are important influences on avifaunal assemblages (Chapter 2,

Chapter 3, Guenette and Villard 2005, Bakermans et al. 2012). However, the extent of forest

cover on species richness and abundance of birds was a still more important predictor and

had a positive effect on all response groups (Chapter 3 and Chapter 4). However, the forest

extent was found strongly important to the forest sensitive species. For example, I found that

species richness and abundance of bark-gleaners and foliage-gleaner insectivores were

significantly related to the extent of forest cover (Chapter 3 and Chapter 4).

Among the most important findings of my thesis is that the extent of forest cover in the

landscape not only has direct effects on birds, but also has a significant role in moderating the

effects of disturbance on birds at the site-level (Chapter 3) as well as at the landscape-level

(Chapter 4). To my knowledge, this is the first study to investigate interactive effects between

intensity of disturbances and species richness at the site scale as well as at the landscape

scale. I found that effects of disturbance on site-level species richness and abundance of all

birds, and site-level species richness and abundance of bark-gleaning and foliage-gleaning

insectivores, depended on the extent of forest cover in the surrounding landscape.

Furthermore, analysis of landscape-level bird species richness indicated a role of forest extent

in moderating the effects of disturbance at the landscape level (Chapter 4). For example, I

found that landscape-level species richness of birds was less-affected by disturbance in sites

with a greater extent of forest cover in the landscape (Chapter 4). Thus, the richness of birds

across landscapes depends on both the extent of forest cover and the level of disturbance in

that landscape.

These findings are of significance for conservation management of avifauna, particularly in

multiple-use forest landscapes where disturbance levels vary across space. The remnant

forests of the region not only support the flora and fauna, but also meet the subsistence

demands of food and shelter for people residing near forests. My results show that the effects

of subsistence forest resources extraction on avifauna assemblages can be compensated for by

maintaining the extent of forests in the landscape (Chapter 3, Chapter 4) which would be a

potential win-win outcome for forest-dependent human communities. Therefore, the focus

should be on the restoration of forest habitats through a participatory approach to forest

conservation that seeks to benefit both biodiversity and the interests of the local people. I

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believe that these findings point a way forward for resolving conflicts among different

stakeholders in forest resources management in lowland Terai forests.

5.5 Relationships between species richness and forest extent vary among bird guilds

There has been an increasing interest in understanding the relationship between landscape

properties and faunal assemblages in human-dominated landscapes (Radford and Bennett

2007, Maron et al. 2012, Taylor et al. 2012). Studies that focus on site-level responses may

not sufficient to characterise the influences on avifauna across the landscape. In Chapter 4, I

adopted a whole-of-landscape approach in characterising the relationships between forest

birds and extent of forest cover, measuring both response and predictor variables at the

landscape-scale (Bennett et al. 2006). I investigated the relative importance of forest extent

and other landscape characteristics such as disturbance levels, rainfall and the extent of water

bodies in the landscape on estimated richness of all birds, frugivores, foliage-gleaners and

sallying insectivores.

The extent of forest cover in a landscape is an important predictor of species-rich

assemblages in the landscape (Haslem and Bennett 2008, Zuckerberg and Porter 2010, Maron

et al. 2012, Moura et al. 2013, Ochoa‐Quintero et al. 2015, Chapter 3, Chapter 4). In the

lowland forests, I found a consistent positive response of landscape-level species richness of

birds to the extent of forest cover in the landscape. Moreover, the relationships between

landscape-level species richness and the extent of forest cover were not uniform across the

foraging guilds. For example, I found that the richness of foliage-gleaning insectivores and,

to a lesser extent, frugivores, were strongly related to the extent of forest cover, whereas

richness of sallying insectivores was only weakly related to forest cover. The strong positive

relationships between foliage-gleaning insectivores and frugivores, and forest cover in the

landscape, may be due to the greater diversity of tree species in landscapes with more forest

cover, which in turn means more diverse foraging substrates and fruit sources.

The nonlinearity of the response of landscape-scale species richness of bird to the extent of

forest cover in the landscape was another important finding of this study (Chapter 4). As the

extent of forest cover in the landscape decreased, species richness also decreased in the

landscape, but this was steepest where forest cover in the landscape was below 20-30%.

However, in contrast to other empirical studies (e.g. Martensen et al., Maron et al. 2012), I

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found that species richness of birds continued to steadily increase up to 100% forest cover,

rather than plateauing (Chapter 4).

Chapter 4 provides clear evidence of the role of forest extent on persistence of forest bird in

lowland Terai forest. As I found, a major change in species richness occurred when forest

cover in the landscape declined to approximately 20-30% of the landscape. Although I did

not find evidence of thresholds in relationships between species richness and forest cover, the

non-linearity response of birds to loss of forest cover in the landscapes suggests that lowland

forest landscapes face a high risk of collapse in avifaunal richness if further forest loss and

modification occur. This information may help guide forest managers and other relevant

authorities to set a conservation target for the protection and restoration of forest in the

landscape to avoid any further loss and decline of species from the region.


My study illustrated potential drivers of forest disturbances and consequent effects on habitat

structure and avifaunal communities at different spatial scales in lowland Terai forests. On

the basis of my findings, I have three major management recommendations. These relate to:

1) the maintenance of complexity of forest stand structure, in particular, retention of large

trees, canopy cover and shrub density; 2) the protection and restoration of forest extent in the

landscape; and 3) off-reserve forest conservation through community forestry approaches.

At the local scale, large tree density, tree canopy cover and shrub density were the most

important habitat features that determined the species richness and abundance of all birds,

bark-gleaners and foliage-gleaning insectivores (Chapter 2, Chapter 3). These habitat features

were key for forest specialist species such as green-billed malkoha Phaenicophaeus tristis,

grey-headed woodpecker, Abbot’s babbler and the greater flameback. As I outlined in

Chapter 3, forest-use practices such as logging, grazing and lopping have contributed to a

reduction of the structural complexity of vegetation, but particularly in state-managed forests.

To optimize species richness at the site level, extraction of forest biomass such as standing

trees, snags, woody debris and large tress should be reduced and strictly regulated. Similarly,

retention of larger trees and tree canopy cover through habitat restoration activities should be

strategy focus for forest conservation.

100

At the landscape-scale, the extent of forest cover was an important predictor of species

richness across the landscape. The maintenance of forest in the landscape supports not only

species-rich assemblages, but also potentially reduces the impacts of disturbance on forest

bird assemblages (Chapter 3, Chapter 4). Furthermore, the protection of a minimum extent of

forest cover in the landscape reduces the chance of abrupt species declines (Andren 1994).

Native remnants of Sal -dominated forest represent the last remaining habitats for much of

biodiversity in lowland landscapes. Although this study was carried out in the lowland Terai

forest of Nepal, my findings are relevant to Sal-dominated lowland landscapes of south Asian

region, and extend our understanding of landscape-level species-area relationships.

In addition to the forest cover, the extent of water bodies (rivers, creeks, lakes) also

influenced species richness. Greater areas of water bodies resulted in higher species richness

of all birds and frugivores (Chapter 4). Thus protection of riparian habitat is likely to

contribute to maintaining bird species richness and abundance in the landscape. Nepal’s

lowland rivers face increased extraction of gravel, sand and boulders, which severely affects

water discharge from the systems (Dahal et al. 2012). These activities also pose a threat to the

vegetation surrounding extraction sites (BCN and DNPWC 2010) and associated bird

communities. It is therefore important that conservation of water resources receives greater

emphasis by implementing sustainable river bed extraction planning in the region.

While Nepal has designated about 23% of its land mass as protected areas, the majority of its

protected land is concentrated in the high Himalayas and throughout the less-productive

landscapes (HMG/MFSC 2002, Heinen and Shrestha 2006). For example, about 48% of high

Himalayas are protected, whereas only 0.8% of the Mild Hills and 5.5% of the Terai zone are

protected (Shrestha et al. 2010). Thus, only a small proportion of Nepal’s most productive

lowland forest is represented in the current protected area network. A large tract of lowland

Terai forest is located adjacent to protected areas. It is therefore critical to adopt landscape

planning and management strategies focused not only on protected areas, but also on off-

reserve forests, which are the last remaining natural habitats in lowland Nepal. Effective

management of these forests should not only improve the biodiversity in those forests, but

also in the adjacent protected areas (Kindlmann 2011).

Protected areas are subject to a relatively strict management regime, with habitat

management and monitoring done by the Department of National Parks and Wildlife

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Conservation (DNPWC). The DNPWC has its sector offices and range posts in the field, and

management is controlled from these field offices. Forest management activities include

regular burning, ploughing and uprooting of unwanted tree species for grassland

management, control of invasive species in particular management of Mikania micrantha,

regular patrolling to prevent illegal logging and hunting of wildlife species. The National

Park and Wildlife Conservation Act in 1992 made provision for buffer zones and provided

limited rights to local communities to manage forest adjacent to national park boundaries

(Paudel et al. 2007), however the buffer zone management committee has no role in

management of the protected areas themselves. Protected areas may be more effectively

managed if the community living in and using the buffer zone participates in conservation

planning and decision-making processes.

State-managed forests occupy a significant proportion of off- reserve forests in lowland

Nepal and are centrally managed. The centralized forest management approach limits

participation of surrounding communities in management and conservation of forest

resources. Unlike protected areas, state-managed forests are managed for production rather

than conservation by the Department of Forest (DoF). The harvesting of forest products, in

particular, collection of logs for timber and fuelwood, is the primary objective of

management of these forests. Excessive extraction of forest products for timber and fuelwood

have negative effects on forest condition (e.g. Kanel and Dahal 2008), so it is important to

develop and implement a sustainable forest management plan that reduces human pressures

on forests to maintain vegetation complexity both at the site- and landscape-scale.

Unlike state-managed and protected area forest tenure types, community managed forests are

based on a participatory approach where communities have the rights of access to resources

and their management. This approach focuses on the collective management of forest

resources to improve both human well-being and biodiversity conservation. Under this

management framework, Community Forest User Groups (CFUGs) are formed and

management authority is given to these user groups for protection, management and

utilization of forest products. Forest restoration and management activities, for example,

plantation of trees and silvicultural operations such as thinning and pruning, removal of

unwanted weeds and forests patrolling to prevent illegal logging and grazing (Nagendra et al.

2005, Kanel and Dahal 2008), are currently undertaken by the community forest user

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committees. This initiative does not only contribute to better forest condition, but also offers

significant benefits to the local community from forest products (e.g. Kanel and Dahal 2008).

In recent years, community-managed forests have become increasingly recognised for their

conservation values for biodiversity (Nagendra and Gokhale 2008, Birendra et al. 2014,

Chapter 2). They are important breeding habitats for many threatened bird species (Chapter 2,

Inskipp 1989). For instance, the last remaining population of blue–eared barbet (<50

individuals) and Abbott’s babbler (<250 individuals) are in the community-managed forests

in eastern Nepal (Inskipp et al. 2013). Moreover, the greater flameback that requires larger

trees to breed (Kumar et al. 2011) was recorded only in the community-managed forests

during this study (Chapter 2).

However, despite the importance of community forests in faunal conservation, the roll out of

community forestry programs in the lowland Terai are progressing slowly because these

forests are commercially valuable and also a major source of government revenue (Kanel and

Dahal 2008). Thus, only 10% of the Terai forests have transferred to community management

(Kanel and Dahal 2008), compared to 24% in the hill regions of Nepal (Bhattarai 2006).

Strengthening community forestry programs can not only ameliorate habitat loss and

degradation (Gautam et al. 2004; Kanel and Dahal 2008) but also generate livelihood

opportunities for surrounding communities and reduce pressure on protected areas (Straede et

al. 2002). I therefore recommend that the extension of community forestry programs should

be prioritized and given strong and urgent Government support in order to minimise further

habitat deterioration.

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Table 5.1 Summary of study themes and key findings of this thesis

Theme Key findings Chapter

Conservation

values of off-

reserve forests

1. Off-reserve forests, in particular, community forests, have complementary bird assemblages to

protected areas, supporting species not represented in formal conservation reserves.

2. Habitat features of sites in protected areas and community forests were relatively more similar, with no

differences in tree canopy cover, large mature tree density and shrub density.

3. Beta diversity (i.e. species turnover) among sites in off-reserve forest tenures was higher than among

sites in protected areas.

4. Local habitat characteristics such as tree density, shrub cover, tree canopy cover and mature tree

density had strong influence on bird communities.

Management implications

1. Strengthen community forestry programs across off-reserve forests in lowland landscapes.

2. Provide technical, managerial and organizational support for institutionalization of community forests

in particular those they are newly formed.

3. Develop and implement sustainable forest extraction guidelines across the off-reserve forests to prevent

further degradation of local habitat characteristics.

2

2

2

2,3

Effects of forest

use practices on

vegetation and

associated

1. Logging, grazing and lopping activities had deleterious effects on vegetation structure and associated

avifaunal communities in the lowland landscape.

2. The density and basal area of large trees and tree canopy cover were significantly lower in heavily-

logged sites than in lightly-logged sites. Lopping activities significantly affected shrub density and

2,3

3

104

avifaunal

communities

shrub cover, while grazing activities reduced herbaceous cover.

3. The overall species richness of birds in sites in heavily grazed and heavily lopped areas was

significantly lower than in lightly grazed and lightly lopped sites. Similarly, the average abundance of

birds was significantly lower in heavily logged sites than lightly logged sites.

4. The effects of forest disturbance on birds were not restricted only to the site level, as higher intensities

of disturbance also negatively affected the landscape-level species richness of birds.


1. Develop and implement a sustainable forest management plan that reduces human pressures on forests

to maintain vegetation complexity at the site-scale.

2. Excessive grazing, over extraction of fire wood and fodder, lopping and removal of tree canopy should

be control with a regulatory mechanism.

3. Retention of larger trees and tree canopy cover in the landscape through habitat restoration activities

should be a focus for forest conservation.

3

4

Relative effects of

site-and landscape-

level factors on

bird communities

1. The structural features of forest stands such as canopy structure, tree sizes, and shrub density had a

significant positive influence on avifaunal assemblages.

2. Species of birds such as bark-gleaning and foliage-gleaning insectivores that require large trees for

their feeding and nesting were strongly related to the density of large trees and tree canopy cover for

their survival.

3. The extent of forest cover in the landscape had a strong positive influence on species richness and

abundance for all bird response groups.

2,3

2,3

2,3,4

105


1. The extraction of forest biomass such as standing trees, snags, woody debris and large tress should be

reduced and strictly regulated to optimize species richness at the site level.

2. Development and effective implementation of policies to restore degraded forests in lowland Terai in

particular Parsa and eastern landscapes is urgent.

3. Maintain existing forest cover as it boosts landscape-level species richness in the region.

Effects of

landscape context

and interaction

effects on bird

communities

1. The extent of forest cover within a 500 m and a 2.5 km radius of each survey site had a strong positive

influence on species richness and abundance for all bird response groups.

2. The strongest effect of forest cover on all species richness occurred in landscapes with a greater extent

of disturbance, indicating a role of forest extent in moderating the effects of disturbance.

3. The species richness of foliage-gleaning insectivores across the landscape was also dependent on the

interaction between rainfall and forest cover, suggesting that the extent of forest cover has less effect on

foliage-gleaning insectivores in higher-rainfall landscapes.


1. Habitat restoration and maintenance of forest extent should be prioritized in areas of low forest cover

particularly in landscapes of eastern lowland.

4. Develop and implement a sustainable forest management plan that reduces human pressures on forests

to maintain vegetation complexity both at the site- and landscape-scale.

3,4

3,4

4

106

Relationship

between species

richness and forest

extent vary with

foraging guilds

1. The landscape-level species richness of forest birds was consistently positively related to the extent of

forest cover in the landscape.

2. However, the strength of the relationship varied substantially among foraging guilds, with effects

strongest for foliage-gleaning and frugivores.

3. The relationship between species richness and extent of forest cover was somewhat nonlinear, with

species richness decreasing more steeply below about 20-30% forest cover in the landscape.


1. Habitat restoration and maintenance of forest extent should be prioritized in areas of low forest cover

and high degree of deforestation to prevent further loss of avifaunal populations in the region.

2. Supporting community forestry program across the off-reserve forests of lowland region may prevent

deforestation and increase forest extent through active restoration and tree planting.

3,4

4

4

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5.7 Limitations of my research

There are number of limitations affecting this study. One limitation was the difficulty in

selection of accessible sites and landscapes proportionally across state forests, community

forests, and protected areas in lowland Terai forests. However, accessibility to all parts of

protected areas was impossible due to transportation challenges and also risks from wild

animals, particularly the greater one horned-rhinoceros Rhinoceros unicornis and the Asian

wild elephant Elephas maximus. This reduced the opportunity for study landscapes within the

protected areas. To address this limitation, some sites were selected from within buffer zone

forests. The buffer zone forests are forests at the immediate periphery of protected area that

have been managed by the protected area management authority and habitat conditions are

largely similar to those of protected areas.

A second limitation of this study was number of bird survey sites that I was able to visit

within each landscape. I established a total of four bird survey sites in each landscape. This

number may be inadequate for full characterisation of the avifauna of landscapes, particularly

in the context of large landscape size (5 x 5 km). However, survey effort was equal among

landscapes, with at least one site located in each quadrant of the landscape, and species

richness estimation was used to correct for differences in survey completeness among

landscapes. Another limitation was that I was only able to survey birds from November to

May, which covered only winter and early-summer bird assemblages. Thus, species that were

summer visitors were poorly captured by this study. These limitations should be considered

in future research.

5.8 Future research

This study revealed consistent relationships between the extent of forest cover and most

groups of birds. However, the observed pattern of consistent positive relationships may not

be due solely to independent effects of forest cover on the landscape. Such relationships may

be due to a correlation between forest extent and the diversity of resources available.

Previous studies have shown that larger areas of forest in the landscape support diverse

vegetation types and therefore provide an array of resources across a range of biodiversity

including avifauna (Miller et al. 1997, Williams et al. 2002). Although lowland forest is

dominated by the Shorea robusta species, it also contains other vegetation types such as

riverine vegetation, mixed hardwood forest, and different gradients of habitat conditions

(Joshi et al. 2003). Thus, the lowland landscape is a heterogeneous mosaic of different

108

vegetation types that in turn, may have strong influences on forest bird assemblages.

Therefore future studies on bird assemblages in the lowland forests should examine these

potential mechanisms behind the patterns I observed.

In this study, I found that the relationship between species richness and extent of forest cover

was non-linear, with species richness decreasing more steeply below about 20-30% forest

cover in the landscape (Chapter 4). However, recent empirical research by Maron et al.

(2012) has shown that the response of species to habitat cover may vary across different

landscape types. The study landscapes differed in terms of rainfall, soil types and topography.

Thus further research is required to investigate the relationships between landscape-scale

forest extent and forest bird assemblages across different landscape types/or landscape

productivity. Such studies would require greater replication of landscapes, but would further

inform managers in setting management targets for maintenance of forest cover to provide

effective protection of biodiversity.

I compared species richness and abundance and community composition of forest birds

among differently-managed forest such as protected areas, state forests and community

forests (Chapter 2). Although the main question here was to explore whether the off-reserve

forest support complementary bird assemblages in the lowland Terai forests, a secondary

interest was to investigate the role of community forestry in bird conservation. I found that

community forests support richer assemblages than state-managed forests (Chapter 2).

However, in this study, I did not examine the effects of different durations of community

forest management on forest bird assemblages. The protected areas and state-managed forests

often have a longer history of similar forest management than the community-based forests.

The forest tenure transfers to community have occurred at different intervals over time in

lowland Terai forests since 1990. It is therefore important to evaluate the effectiveness of

participatory forest management practices for bird conservation across different intervals of

time since the tenure rights transferred to the local communities. This would allow

documentation of the conservation outcomes of community managed forests across time and

space.

5.9 Conclusion

The remnant lowland Terai forests are subject to intense anthropogenic pressures due to

agricultural intensification and forestry practices. These human-induced disturbances

109

threatened ecological communities. However, the influence of landscape modifications on

habitat features and their constituent biota at different spatial scales is poorly understood in

the region. Biodiversity conservation in such landscapes depends on the effective

conservation of remnant forests and associated habitat attributes. Thus, identifying how

species respond to habitat characteristics at different spatial is crucial for conservation of

avifauna in the region.

This thesis makes an important contribution to our understanding of relative effects of habitat

characteristics on bird assemblages at multiple spatial scales in human-dominated multiple-

use landscapes. It provides a broader picture about the effects of forest disturbances on

species assemblages as result of subsistence forest-use practices, reveals a significant role of

landscape-level forest extent to reduce these effects on bird assemblages, and describes the

nature and shape of species and forest extent relationships for birds of different functional

groups in the lowland Terai forests. This study therefore provides important ecological

information for forest managers and other stakeholders seeking to achieve landscape-level


110

REFERENCES

Adams, E.M., Morrison, M.L., 1993. Effects of Forest Stand Structure and Composition on

Red-Breasted Nuthatches and Brown Creepers. The Journal of Wildlife Management,

57(3): 616-629.

Aleixo, A. 1999. Effects of selective logging on a bird community in the Brazilian Atlantic

forest. The Condor 101:537-548.

Alexander, J. D., J. L. Stephens, and N. E. Seavy. 2008. Livestock utilization and bird

community composition in mixed-conifer forest and oak woodland in southern

Oregon. Northwest Science 82:7-17.

Ali, S., Ripley, S.D., 1983. Handbook of the birds of the Indian subcontinent. Compact

edition. Oxford University Press, Delhi.

Ali, S., S. D. Ripley, and J. H. Dick. 1987. Compact handbook of the birds of India and

Pakistan.

Anand, M. O., J. Krishnaswamy, and A. Das. 2008. Proximity to forests drives bird

conservation value of coffee plantations: implications for certification. Ecological

Applications 18:1754-1763.

Anderson, M.J., 2001. A new method for non-parametric multivariate analysis of variance.

Austral Ecology 26: 32-46.

Anderson, M.J., Ellingsen, K.E.,McArdle, B.H., 2006. Multivariate dispersion as a measure

of beta diversity. Ecology letters, 9(6): 683-93.

Anderson, M.J., Gorley, R.N., Clarke, K.R., 2008. PERMANOVA for PRIMER: guide to

software and statistical methods.PRIMER–E Ltd., Plymouth, United Kingdom, 214

pp.

Andren, H. 1994. Effects of habitat fragmentation on birds and mammals in landscapes with

different proportions of suitable habitat: a review. Oikos:355-366.

Bailis, R., J. L. Chatellier, and A. Ghilardi. 2012. Ecological sustainability of woodfuel as an

energy source in rural communities. Pages 299-325 Integrating Ecology and Poverty

Reduction. Springer.

Bakermans, M. H., A. D. Rodewald, and A. C. Vitz. 2012. Influence of forest structure on

density and nest success of mature forest birds in managed landscapes. The Journal of

Wildlife Management 76:1225-1234.

Balmford, A., R. Green, and B. Phalan. 2012. What conservationists need to know about

farming. Proceedings of the Royal Society B: Biological Sciences:rspb20120515.

111

Baral, H. S., A. K. Ram, B. Chaudhary, S. Basnet, H. Chaudhary, T. R. Giri, and D.

Chaudhary. 2012. Conservation status of Bengal Florican Houbaropsis bengalensis

bengalensis (Gmelin, 1789)(Gruiformes: Otididae) in Koshi Tappu Wildlife Reserve

and adjoining areas, eastern Nepal. Journal of Threatened Taxa 4:2464-2469.

Baral, H., B. Sahgal, S. Mohsanin, K. Namgay, and A. Khan. 2014. Species and habitat

conservation through small locally recognised and community managed Special

Conservation Sites. Journal of Threatened Taxa 6:5677-5685.

Baral, H.S., Inskipp, C., 2005. Important Bird Areas in Nepal: key sites for conservation.

Bird Conservation Nepal and BirdLife International,Kathmandu and Cambridge.

Baral, H.S.,Inskipp, C., 2004. The State of Nepal’s Birds 2004. Department of National Parks

and Wildlife Conservation, Bird Conservation Nepal & IUCN Nepal, Kathmandu,

64pp.

Barr, L. M., R. L. Pressey, R. A. Fuller, D. B. Segan, E. McDonald-Madden, and H. P.

Possingham. 2011. A new way to measure the world's protected area coverage. PLoS

One 6:e24707.

Bartoń, K. 2012. MuMIn: multi-model inference. R package version 1.

Bastian, O., D. Haase, and K. Grunewald. 2012. Ecosystem properties, potentials and

services–the EPPS conceptual framework and an urban application example.

Ecological Indicators 21:7-16.

Bates, D., M. Maechler, B. Bolker, and S. Walker. 2014. lme4: Linear mixed-effects models

using Eigen and S4. R package version 1.1-7, http://CRAN.R-

project.org/package=lme4.

BCN and DNPWC. 2010. The State of Nepal’s Birds 2010. Bird Conservation Nepal and

Department of National Parks and Wildlife Conservation, Kathmandu.

BCN, DNPWC, 2012. Conserving biodiversity and delivering ecosystem services at

Important Bird Areas in Nepal. Kathmandu and Cambridge, UK: Bird Conservation

Nepal,Department of National Parks and Wildlife Conservation, and BirdLife

International.

Beck, N. and J. N. Katz. 1995. What to do (and not to do) with time-series cross-section data.

American political science review 89:634-647.

Becker, C. G., Fonseca, C. R., Haddad, C. F. B., Batista, R. F. and Prado, P. I. 2007. Habitat

split and the global decline of amphibians. Science, 318(5857), 1775-1777.

112

Bennett, A. F. and D. A. Saunders. 2010. Habitat fragmentation and landscape change.

Conservation Biology for All 93:1544-1550.

Bennett, A. F., D. G. Nimmo, and J. Q. Radford. 2014. Riparian vegetation has

disproportionate benefits for landscape‐scale conservation of woodland birds in highly

modified environments. Journal of Applied Ecology 51:514-523.

Bennett, A. F., J. Q. Radford, and A. Haslem. 2006. Properties of land mosaics: implications

for nature conservation in agricultural environments. Biological Conservation

133:250-264.

Berry, N. J., Phillips, O. L., Ong, R. C., and Hamer, K. C. 2008. Impacts of selective logging

on tree diversity across a rainforest landscape: the importance of spatial scale.

Landscape ecology, 23(8), 915-929.

Betts, M. G., G. J. Forbes, and A. W. Diamond. 2007. Thresholds in songbird occurrence in

relation to landscape structure. Conservation Biology 21:1046-1058.

Bhagwat, S.A., Kushalappa, C.G., Williams, P.H., Brown, N.D., 2005. The Role of Informal

Protected Areas in Maintaining Biodiversity in the Western Ghats of India. Ecology

and Society, 10(1): 1-42.

Bhagwat, S.A.,Rutte, C., 2006. Sacred Groves: Potential for Biodiversity Management.

Frontiers in Ecology and the Environment, 4(10): 519-524.

Bhattarai, B. 2006. Widening the gap between terai and hill farmers in Nepal: the

implications of the New Forest Policy 2000. In: Mahanty, S., Fox, J., Nurse,

M.,Stephen, P., McLees, L., (Eds.), Hanging in the Balance: Equity in Community-

Based Natural Resource Management in Asia, RECOFTC, Bangkok, Thailand, and

East-West Center, Honolulu, Hawaii, USA. Brawn, J.D., Robinson, S.K., Thompson,

F.R., 2001.

Bhusal, N. P. 2014. Buffer Zone Management System in Protected Areas of Nepal. The Third

Pole: Journal of Geography Education 11:34-44.

Bibby, C. J. 2000. Bird census techniques. Elsevier.

BirdLife International. 2004. Threatened birds of the world 2004 CD-ROM. Cambridge, UK:

BirdLife International.

Birendra, K., A. J. Mohammod, and M. Inoue. 2014. Community Forestry in Nepal’s Terai

Region: Local Resource Dependency and Perception on Institutional Attributes.

Environment and Natural Resources Research 4:p142.

113

Bjornstad, O. N. 2013. spatial nonparametric covariance functions. Package ‘ncf’. URL

http://onb.ent.psu.edu/onb1/R.

Blake, J. G. and B. A. Loiselle. 2001. Bird assemblages in second-growth and old-growth

forests, Costa Rica: perspectives from mist nets and point counts. The Auk 118:304-

326.

Bolker, B. M., M. E. Brooks, C. J. Clark, S. W. Geange, J. R. Poulsen, M. H. H. Stevens, and

J.-S. S. White. 2009. Generalized linear mixed models: a practical guide for ecology

and evolution. Trends in ecology & evolution 24:127-135.

Borah, N., Athokpam, F. D., Garkoti, S., Das, A. K., and Hore, D. 2014. Structural and

compositional variations in undisturbed and disturbed tropical forests of Bhuban hills

in south Assam, India. International Journal of Biodiversity Science, Ecosystem

Services & Management, 10(1), 9-19.

Brawn, J.D., Robinson, S.K., Thompson, F.R., 2001. The role of disturbance in the ecology

and conservation of birds. Annual Review of Ecology and Systematics, 32: 251-276.

Bried, J. T., K. E. Langwig, A. A. Dewan, and N. A. Gifford. 2011. Habitat associations and

survey effort for shrubland birds in an urban pine barrens preserve. Landscape and

Urban Planning 99:218-225.

Brooks, T. M., M. I. Bakarr, T. Boucher, G. A. Da Fonseca, C. Hilton-Taylor, J. M. Hoekstra,

T. Moritz, S. Olivieri, J. Parrish, and R. L. Pressey. 2004. Coverage provided by the

global protected-area system: is it enough? BioScience 54:1081-1091.

Bruton, M. J. 2014. Multi-scale patterns of habitat use by reptiles in regenerating dryland

landscapes. doi:10.14264/uql.2014.84.

Buckland, S. T. and C. Handel. 2006. Point-transect surveys for songbirds: robust

methodologies. The Auk 123:345-357.

Buckland, S. T., D. R. Anderson, K. P. Burnham, and J. L. Laake. 2005. Distance sampling.

Wiley Online Library.

Buffum, B., G. Gratzer, and Y. Tenzin. 2009. Forest grazing and natural regeneration in a late

successional broadleaved community forest in Bhutan. Mountain Research and

Development 29:30-35.

Burgman, M., D. Lindenmayer, and J. Elith. 2005. Managing landscapes for conservation

under uncertainty. Ecology 86:2007-2017.

Burnham, K. P. and D. R. Anderson. 2002. Model selection and multimodel inference: a

practical information-theoretic approach. Springer.

http://onb.ent.psu.edu/onb1/R

114

Butchart, S. H., J. P. Scharlemann, M. I. Evans, S. Quader, S. Arico, J. Arinaitwe, M.

Balman, L. A. Bennun, B. Bertzky, and C. Besancon. 2012. Protecting important sites

for biodiversity contributes to meeting global conservation targets. PLoS One

7:e32529.

Capinera, J. 2011. Insects and wildlife: arthropods and their relationships with wild vertebrate

animals. John Wiley & Sons.

Central Bureau of Statistics. 2011. Population census 2011.National report. Central Bureau of

Statistics, Government of Nepal, Kathmandu, Nepal.

Chandler, C. C., D. I. King, and R. B. Chandler. 2012. Do mature forest birds prefer early-

successional habitat during the post-fledging period? Forest Ecology and

Management 264:1-9.

Chao, S. 2012. FOREST PEOPLES: Numbers across the world. Forest Peoples Programme.

Chappell, M. J. and L. A. LaValle. 2011. Food security and biodiversity: can we have both?

An agroecological analysis. Agriculture and Human Values 28:3-26.

Charnley, S. and M. R. Poe. 2007. Community Forestry in Theory and Practice: Where Are

We Now?*. Annual Review of Anthropology 36:301.

Chettri, N., E. Sharma, D. Deb, and R. Sundriyal. 2002. Impact of firewood extraction on tree

structure, regeneration and woody biomass productivity in a trekking corridor of the

Sikkim Himalaya. Mountain Research and Development 22:150-158.

Chisholm, C., Z. Lindo, and A. Gonzalez. 2011. Metacommunity diversity depends on

connectivity and patch arrangement in heterogeneous habitat networks. Ecography

34:415-424.

Christensen, M., S. Rayamajhi, and H. Meilby. 2009. Balancing fuelwood and biodiversity

concerns in rural Nepal. Ecological Modelling 220:522-532.

Clark, N. E., E. H. Boakes, P. J. McGowan, G. M. Mace, and R. A. Fuller. 2013. Protected

Areas in South Asia Have Not Prevented Habitat Loss: A Study Using Historical

Models of Land-Use Change. PLoS One 8:e65298.

Clarke, K.R., 1993. Non-parametric multivariate analysis of changes in community structure.

Australian Journal of Ecology 18: 117-143.

Clarke, K.R., Warwick, R.M., 2001. Change in marine communities: an approach to

statistical analysis and interpretation. 2nd Edn. (PRIMER-E: Plymouth.).

Cleary, D. F., T. J. Boyle, T. Setyawati, C. D. Anggraeni, E. E. V. Loon, and S. B. Menken.

2007. Bird species and traits associated with logged and unlogged forest in Borneo.

Ecological Applications 17:1184-1197.

115

Cleveland, W. S. and S. J. Devlin. 1988. Locally weighted regression: an approach to

regression analysis by local fitting. Journal of the American Statistical Association

83:596-610.

Colwell, R. 2013. Estimates 9.1. 0 User's Guide. Storrs, CT: Department of Ecology and

Evolutionary Biology, University of Connecticut.

Condit, R., N. Pitman, E. G. Leigh, J. Chave, J. Terborgh, R. B. Foster, P. Núnez, S. Aguilar,

R. Valencia, and G. Villa. 2002. Beta-diversity in tropical forest trees. Science

295:666-669.

Cox, R.L., Underwood, E.C., 2011. The Importance of Conserving Biodiversity Outside of

Protected Areas in Mediterranean Ecosystems. PLoS ONE, 6(1): e14508.

Crawley, M. J. 2007. Mixed‐Effects Models. The R Book, Second Edition:681-714.

Cushman, S. A. and K. McGarigal. 2003. Landscape-level patterns of avian diversity in the

Oregon Coast Range. Ecological Monographs 73:259-281.

Dahal, B. R., C. A. McAlpine, and M. Maron. 2014. Bird conservation values of off-reserve

forests in lowland Nepal. Forest Ecology and Management 323:28-38.

Dahal, B. R., P. J. McGowan, S. J. Browne. 2009. An assessment of census techniques,

habitat use and threats to Swamp Francolin Francolinus gularis in Koshi Tappu

Wildlife Reserve, Nepal. Bird Conservation International 19:137-147.

Dahal, K. R., S. Sharma, and C. M. Sharma. 2012. A review of riverbed extraction and its

effects on aquatic environment with special reference to Tinau River, Nepal. Hydro

Nepal: Journal of Water, Energy and Environment 11:49-56.

Davies, R. B. 1987. Hypothesis testing when a nuisance parameter is present only under the

alternative. Biometrika 74:33-43.

De La Montana, E., Rey-Benayas, J.M.,Carrascal, L.M., 2006. Response of bird communities

to silvicultural thinning of Mediterranean maquis. Journal of Applied Ecology, 43(4):

651-659.

Deconchat, M., Brockerhoff, E., and Barbaro, L. 2009. Effects of surrounding landscape

composition on the conservation value of native and exotic habitats for native forest

birds. Forest Ecology and Management, 258, S196-S204.

Deconchat, M., E. Brockerhoff, and L. Barbaro. 2009. Effects of surrounding landscape

composition on the conservation value of native and exotic habitats for native forest

birds. Forest Ecology and Management 258:S196-S204.

116

DeFries, R. S., J. A. Foley, and G. P. Asner. 2004. Land-use choices: balancing human needs

and ecosystem function. Frontiers in Ecology and the Environment 2:249-257.

Dennis, P., J. Skartveit, D. I. McCracken, R. J. Pakeman, K. Beaton, A. Kunaver, and D. M.

Evans. 2008. The effects of livestock grazing on foliar arthropods associated with bird

diet in upland grasslands of Scotland. Journal of Applied Ecology 45:279-287.

Díaz, I. A., J. J. Armesto, S. Reid, K. E. Sieving, and M. F. Willson. 2005. Linking forest

structure and composition: avian diversity in successional forests of Chiloé Island,

Chile. Biological Conservation 123:91-101.

Döbert, T. F., B. L. Webber, A. D. Barnes, K. J. Dickinson, R. K. Didham, C. Kettle, and L.

Koh. 2014. 4 Forest Fragmentation and Biodiversity Conservation in Human-

dominated Landscapes. Global Forest Fragmentation:28.

DoF. 2009. Community forestry database.Ministry of Forest and Soil Conservation.

Kathmandu, Nepal.

Dudley, N. and A. Phillips. 2006. Forests and Protected Areas: Guidance on the use of the

IUCN protected area management categories. IUCN, Gland, Switzerland and

Cambridge, UK. x + 58pp.

Dunning, J. B., B. J. Danielson, and H. R. Pulliam. 1992. Ecological processes that affect

populations in complex landscapes. Oikos:169-175.

Ellis, E. A. and L. Porter-Bolland. 2008. Is community-based forest management more

effective than protected areas?: A comparison of land use/land cover change in two

neighboring study areas of the Central Yucatan Peninsula, Mexico. Forest Ecology

and Management 256:1971-1983.

Ellis, E. C. and N. Ramankutty. 2008. Putting people in the map: anthropogenic biomes of

the world. Frontiers in Ecology and the Environment 6:439-447.

Estavillo, C., R. Pardini, and P. L. B. da Rocha. 2013. Forest Loss and the Biodiversity

Threshold: An Evaluation Considering Species Habitat Requirements and the Use of

Matrix Habitats. PLoS One 8:e82369.

Ewers, R. M. and R. K. Didham. 2006. Confounding factors in the detection of species

responses to habitat fragmentation. Biological Reviews 81:117-142.

Eyre, T. J., M. Maron, M. T. Mathieson, and M. Haseler. 2009. Impacts of grazing, selective

logging and hyper‐aggressors on diurnal bird fauna in intact forest landscapes of the

Brigalow Belt, Queensland. Austral Ecology 34:705-716.

Fahrig, L. 2001. How much habitat is enough? Biological Conservation 100:65-74.

117

Fahrig, L. 2003. Effects of habitat fragmentation on biodiversity. Annual review of ecology,

evolution, and systematics:487-515.

FAO. 2005. Food and Agriculture Organization of the United Nations; Rome, Italy: 2005.

Global forest resource assessment.

Fischer, J. and D. B. Lindenmayer. 2007. Landscape modification and habitat fragmentation:

a synthesis. Global Ecology and Biogeography 16:265-280.

Flynn, D. F., M. Gogol‐Prokurat, T. Nogeire, N. Molinari, B. T. Richers, B. B. Lin, N.

Simpson, M. M. Mayfield, and F. DeClerck. 2009. Loss of functional diversity under

land use intensification across multiple taxa. Ecol Lett 12:22-33.

Foley, J. A., R. DeFries, G. P. Asner, C. Barford, G. Bonan, S. R. Carpenter, F. S. Chapin, M.

T. Coe, G. C. Daily, and H. K. Gibbs. 2005. Global consequences of land use. Science

309:570-574.

Forsman, J. T., P. Reunanen, J. Jokimäki, M. Mönkkönen. 2010. The effects of small-scale

disturbance on forest birds: a meta-analysis. Can. J. For. Res. , 40: 1833-1842

Fox, J., S. Weisberg, D. Bates, and M. J. Fox. 2012. Package ‘car’. R Foundation for

Statistical Computing, Vienna, Austria.

FRA/DFRS. 2014. Terai Forests of Nepal (2010 – 2012). Babarmahal, Kathmandu: Forest

Resource Assessment Nepal Project/Department of Forest Research and Survey.

Department of Forest Research and Survey, P.O. Box 3339, Babarmahal, Kathmandu,

Nepal.

Franklin, A. B., B. R. Noon, and T. L. George. 2002. What is habitat fragmentation? Studies

in Avian Biology 25:20-29.

Galitsky, C. and J. J. Lawler. 2015. Relative influence of local and landscape factors on bird

communities vary by species and functional group. Landscape Ecology:1-13.

Gardner, T. A., J. Barlow, R. Chazdon, R. M. Ewers, C. A. Harvey, C. A. Peres, and N. S.

Sodhi. 2009. Prospects for tropical forest biodiversity in a human‐modified world.

Ecol Lett 12:561-582.

Gauch, H.G., 1982. Multivariate Analysis in Community Ecology. Cambridge University

Press, Cambridge, England. 298 pages. Chinese edition 1989.

Gautam, A. P. 2007. Group size, heterogeneity and collective action outcomes: Evidence

from community forestry in Nepal. International Journal of Sustainable Development

& World Ecology 14:574-583.

118

Gautam, A. P., G. P. Shivakoti, and E. L. Webb. 2004. Forest cover change, physiography,

local economy, and institutions in a mountain watershed in Nepal. Environmental

Management 33:48-61.

Gautam, A., G. Shivakoti, and E. Webb. 2004. A review of forest policies, institutions, and

changes in the resource condition in Nepal. International Forestry Review 6:136-148.

Gelman, A., Y.-S. Su, M. Yajima, J. Hill, M. Grazia Pittau, J. Kerman, and T. Zheng. 2012.

Arm: data analysis using regression and multilevel/hierarchical models. R package v.

1.5–02.

Gibbs, H. K., A. Ruesch, F. Achard, M. Clayton, P. Holmgren, N. Ramankutty, and J. Foley.

2010. Tropical forests were the primary sources of new agricultural land in the 1980s

and 1990s. Proceedings of the National Academy of Sciences 107:16732-16737.

Gil-Tena, A., O. Torras, and S. Saura. 2008. Relationships between forest landscape structure

and avian species richness in NE Spain. Ardeola 55:27-40.

Graham, C. H. andBlake, J. G. 2001. Influence of patch-and landscape-level factors on bird

assemblages in a fragmented tropical landscape. Ecological Applications, 11(6),

1709-1721.

Gregory, R. D., D. W. Gibbons, and P. F. Donald. 2004. Bird census and survey techniques.

Bird ecology and conservation:17-56.

Greve, M., S. L. Chown, B. J. van Rensburg, M. Dallimer, and K. J. Gaston. 2011. The

ecological effectiveness of protected areas: a case study for South African birds.

Animal Conservation 14:295-305.

Grimmett, R., C. Inskipp, and T. Inskipp. 1999. Guide to the birds of India, Pakistan, Nepal,

Bangladesh, Bhutan, Sri Lanka, and the Maldives.

Grimmett, R., C. Inskipp, and T. Inskipp. 2000. Birds of Nepal. Christopher Helm.

Grimmett, R., Inskipp, C. and Inskipp, T., 2009. Birds of Nepal. Om Books International.

Guenette, J. S. and M. A. Villard. 2005. Thresholds in forest bird response to habitat

alteration as quantitative targets for conservation. Conservation Biology 19:1168-

1180.

Gurung, A., Bryesecon, I., Oh, S., 2011. Micro hydropower: a promising decentralized

renewable technology and its impact on rural livelihoods. Scientific Research and

Essays 6(6): 1240-1248.

Haddad, N. M., G. M. Crutsinger, K. Gross, J. Haarstad, J. M. Knops, and D. Tilman. 2009.

Plant species loss decreases arthropod diversity and shifts trophic structure. Ecol Lett

12:1029-1039.

119

Halley, J. M. and Y. Iwasa. 2011. Neutral theory as a predictor of avifaunal extinctions after

habitat loss. Proceedings of the National Academy of Sciences 108:2316-2321.

Hansen, A.J.,DeFries, R., 2007. Ecological mechanisms linking protected areas to

surrounding lands. Ecological Applications, 17(4): 974-988.

Harvey, C. A., A. Medina, D. M. Sánchez, S. Vílchez, B. Hernández, J. C. Saenz, J. M. Maes,

F. Casanoves, and F. L. Sinclair. 2006. Patterns of animal diversity in different forms

of tree cover in agricultural landscapes. Ecological Applications 16:1986-1999.

Haslem, A. and A. F. Bennett. 2008. Birds in agricultural mosaics: the influence of landscape

pattern and countryside heterogeneity. Ecological Applications 18:185-196.

Hayes, T. and E. Ostrom. 2005. Conserving the world's forests: Are protected areas the only

way. Ind. L. Rev. 38:595.

Hegde, R. and T. Enters. 2000. Forest products and household economy: a case study from

Mudumalai Wildlife Sanctuary, Southern India. Environmental Conservation 27:250-

259.

Heikkinen, R. K., M. Luoto, R. Virkkala, and K. Rainio. 2004. Effects of habitat cover,

landscape structure and spatial variables on the abundance of birds in an agricultural–

forest mosaic. Journal of applied ecology 41:824-835.

Heinen, J. T. and S. K. Shrestha. 2006. Evolving policies for conservation: an historical

profile of the protected area system of Nepal. Journal of Environmental Planning and


Hinsley, S. A. 2000. The costs of multiple patch use by birds. Landscape Ecology 15:765-

775.

HMG/MFSC. 2002. Nepal Biodiversity Strategy. Supported by the Global Environment

Facility (GEF) and UNDP. Kathmandu, Nepal: HMG.

Hoekstra, J. M., T. M. Boucher, T. H. Ricketts, and C. Roberts. 2005. Confronting a biome

crisis: global disparities of habitat loss and protection. Ecol Lett 8:23-29.

Hrabovszky, J. P. and K. Miyan. 1987. Population growth and land use in Nepal" the great

turnabout". Mountain Research and Development:264-270.

Hu, G., J. Wu, K. J. Feeley, G. Xu, and M. Yu. 2012. The effects of landscape variables on

the species-area relationship during late-stage habitat fragmentation. PLoS One

7:e43894.

Huggett, A. J. 2005. The concept and utility of ‘ecological thresholds’ in biodiversity

conservation. Biological Conservation 124:301-310.

120

Inskipp, C. 1989. Nepal’s Forest Birds: Their Status and Conservation. ICBP. Monograph No

4. International Council for Bird Preservation, Cambridge, UK.

Inskipp, C., Baral, H., Inskipp, T.,Stattersfield, A., 2013. The state of Nepal birds 2010.

Journal of Threatened Taxa, 5(1): 3473-3503.

IUCN. 2009. Red List of Threatened Species. International Union for Conservation of

Nature, Cambridge, UK.

Iwata, T., S. Nakano, and M. Murakami. 2003. Stream meanders increase insectivorous bird

abundance in riparian deciduous forests. Ecography 26:325-337.

Jackson, J. K., C. Stapleton, and J.-P. Jeanrenaud. 1994. Manual of afforestation in Nepal.

Jacoby, W. G. 2000. Loess:: a nonparametric, graphical tool for depicting relationships

between variables. Electoral Studies 19:577-613.

Järvinen, O. and R. Väisänen. 1975. Estimating relative densities of breeding birds by the line

transect method. Oikos:316-322.

Jenkins, C. N. and L. Joppa. 2009. Expansion of the global terrestrial protected area system.

Biological Conservation 142:2166-2174.

Jetz, W., C. H. Sekercioglu, and J. E. Watson. 2008. Ecological correlates and conservation

implications of overestimating species geographic ranges. Conservation Biology

22:110-119.

Johnson, M., Reich, P., Nally, R.M., 2007. Bird assemblages of a fragmented agricultural

landscape and the relative importance of vegetation structure and landscape pattern.

Wildlife Research, 34: 185-193.

Joppa, L. N. and A. Pfaff. 2011. Global protected area impacts. Proceedings of the Royal

Society B: Biological Sciences 278:1633-1638.

Joppa, L.N., Loarie, S.R.,Pimm, S.L., 2008. On the protection of “protected areas”.

Proceedings of the National Academy of Sciences, 105(18): 6673-6678.

Joshi, A. R., M. Shrestha, J. L. Smith, and S. Ahearn. 2003. Forest classification of Terai Arc

Landscape (TAL) based on Landsat7 satellite data. A Final Report submitted WWF-

US.

Kallimanis, A. S., A. D. Mazaris, J. Tzanopoulos, J. M. Halley, J. D. Pantis, and S. P.

Sgardelis. 2008. How does habitat diversity affect the species–area relationship?

Global Ecology and Biogeography 17:532-538.

Kallimanis, A. S., A. D. Mazaris, J. Tzanopoulos, J. M. Halley, J. D. Pantis, and S. P.

Sgardelis. 2008. How does habitat diversity affect the species–area relationship?

Global Ecology and Biogeography 17:532-538.

121

Kanel, K. R. and G. R. Dahal. 2008. Community forestry policy and its economic

implications: an experience from Nepal. International Journal of Social Forestry 1:50-

60.

Kessler, M., S. Abrahamczyk, M. Bos, D. Buchori, D. D. Putra, S. R. Gradstein, P. Höhn, J.

Kluge, F. Orend, and R. Pitopang. 2009. Alpha and beta diversity of plants and

animals along a tropical land-use gradient. Ecological Applications 19:2142-2156.

Kessler, W. B. and K. A. Milne. 1982. Morning versus evening detectability of southeast

Alaskan birds. Condor 84:447-448.

Kettle, C. J. and L. P. Koh. 2014. Global Forest Fragmentation. CABI.

Khanaposhtani, M. G., M. Kaboli, M. Karami, and V. Etemad. 2012. Effect of habitat

complexity on richness, abundance and distributional pattern of forest birds. Environ

Manage 50:296-303.

Kindlmann, P. 2011. Himalayan biodiversity in the changing world. Springer Science &

Business Media.

Kissling, W. D., C. Rahbek, and K. Böhning-Gaese. 2007. Food plant diversity as broad-scale

determinant of avian frugivore richness. Proceedings of the Royal Society B:

Biological Sciences 274:799-808.

Kitamura, K. and R. A. Clapp. 2013. Common property protected areas: Community control

in forest conservation. Land Use Policy 34:204-212.

Klooster, D. and O. Masera. 2000. Community forest management in Mexico: carbon

mitigation and biodiversity conservation through rural development. Global

Environmental Change 10:259-272.

Knopf, F. L. and F. B. Samson. 1994. Scale perspectives on avian diversity in western

riparian ecosystems. Conservation Biology 8:669-676.

Koleff, P., K. J. Gaston, and J. J. Lennon. 2003. Measuring beta diversity for presence–

absence data. Journal of Animal Ecology 72:367-382.

Kumar, R. and G. Shahabuddin. 2005. Effects of biomass extraction on vegetation structure,

diversity and composition of forests in Sariska Tiger Reserve, India. Environmental

Conservation 32:248-259.

Kumar, R., G. Shahabuddin, and A. Kumar. 2011. How good are managed forests at

conserving native woodpecker communities? A study in sub-Himalayan dipterocarp

forests of northwest India. Biological Conservation 144:1876-1884.

122

Kumar, R., G. Shahabuddin, and A. Kumar. 2014. Habitat determinants of woodpecker

abundance and species richness in sub-Himalayan dipterocarp forests of north-west

India. Acta Ornithologica 49:243-256.

Laiolo, P., E. Caprio, and A. Rolando. 2004. Can forest management have season-dependent

effects on bird diversity? Biodiversity & Conservation 13:1925-1941.

Lambrick, F. H., N. D. Brown, A. Lawrence, and D. P. Bebber. 2014. Effectiveness of

Community Forestry in Prey Long Forest, Cambodia. Conservation Biology 28:372-

381.

Laurance, W. F. 2010. Habitat destruction: death by a thousand cuts. Conservation Biology

for All 1:73-88.

Laurance, W.F. et al., 2012. Averting biodiversity collapse in tropical forest protected areas.

Nature, 489(7415): 290-4.

Law, B. and C. Dickman. 1998. The use of habitat mosaics by terrestrial vertebrate fauna:

implications for conservation and management. Biodiversity & Conservation 7:323-

333.

Lawrence, A., K. Paudel, R. Barnes, and Y. Malla. 2006. Adaptive value of participatory

biodiversity monitoring in community forestry. Environmental Conservation 33:325-

334.

Leal, A. I., R. A. Correia, J. M. Palmeirim, and J. P. Granadeiro. 2013. Does canopy pruning

affect foliage-gleaning birds in managed cork oak woodlands? Agroforestry systems

87:355-363.

Lee, M.-B. and J. P. Carroll. 2014. Relative importance of local and landscape variables on

site occupancy by avian species in a pine forest, urban, and agriculture matrix. Forest

Ecology and Management 320:161-170.

Lee, T. M., N. S. Sodhi, and D. M. Prawiradilaga. 2007. The importance of protected areas

for the forest and endemic avifauna of Sulawesi (Indonesia). Ecological Applications

17:1727-1741.

Lindenmayer, D. and G. Luck. 2005. Synthesis: thresholds in conservation and management.


Lindenmayer, D. and M. Burgman. 2005. Practical conservation biology. Csiro Publishing.

Lindenmayer, D. B. 2009. Forest wildlife management and conservation. Annals of the New

York Academy of Sciences 1162:284-310.

Lindenmayer, D. B. and J. Fischer. 2006. Habitat fragmentation and landscape change: an

ecological and conservation synthesis. Island Press.

123

Lindenmayer, D. B. and S. A. Cunningham. 2013. Six principles for managing forests as

ecologically sustainable ecosystems. Landscape Ecology 28:1099-1110.

Lindenmayer, D. B., W. F. Laurance, and J. F. Franklin. 2012. Global decline in large old

trees. Science 338:1305-1306.

Ludwig, J. A., R. W. Eager, A. C. Liedloff, J. C. McCosker, D. Hannah, N. Y. Thurgate, J. C.

Woinarski, and C. P. Catterall. 2000. Clearing and grazing impacts on vegetation

patch structures and fauna counts in eucalypt woodland, Central Queensland. Pacific

Conservation Biology 6:254.

Magurran, A. E. 2004. Measuring biological diversity.

Manuwal, D. A. and A. B. Carey. 1991. Methods for measuring populations of small, diurnal

forest birds.

Maron, M. and A. Lill. 2005. The influence of livestock grazing and weed invasion on habitat

use by birds in grassy woodland remnants. Biological Conservation 124:439-450.

Maron, M. and S. Kennedy. 2007. Roads, fire and aggressive competitors: determinants of

bird distribution in subtropical production forests. Forest Ecology and Management

240:24-31.

Maron, M., M. Bowen, R. A. Fuller, G. C. Smith, T. J. Eyre, M. Mathieson, J. E. Watson, and

C. A. McAlpine. 2012. Spurious thresholds in the relationship between species

richness and vegetation cover. Global Ecology and Biogeography 21:682-692.

Martensen, A. C., M. C. Ribeiro, C. Banks‐Leite, P. I. Prado, and J. P. Metzger. 2012.

Associations of forest cover, fragment area, and connectivity with Neotropical

understory bird species richness and abundance. Conservation Biology 26:1100-1111.

Martin, T. G. and S. McIntyre. 2007. Impacts of livestock grazing and tree clearing on birds

of woodland and riparian habitats. Conservation Biology 21:504-514.

Marzluff, J. M., M. G. Raphael, and R. Sallabanks. 2000. Understanding the effects of forest

management on avian species. Wildlife Society Bulletin:1132-1143.

Mathur, P. and P. Sinha. 2008. Looking beyond protected area networks: a paradigm shift in

approach for biodiversity conservation. International Forestry Review 10:305-314.

Mayer, A., V. Stöckli, W. Konold, and M. Kreuzer. 2006. Influence of cattle stocking rate on

browsing of Norway spruce in subalpine wood pastures. Agroforestry systems

66:143-149.

McAlpine, C. A., J. R. Rhodes, J. G. Callaghan, M. E. Bowen, D. Lunney, D. L. Mitchell, D.

V. Pullar, and H. P. Possingham. 2006. The importance of forest area and

124

configuration relative to local habitat factors for conserving forest mammals: a case

study of koalas in Queensland, Australia. Biological Conservation 132:153-165.

McAlpine, C., M. Bowen, G. Smith, G. Gramotnev, A. Smith, A. L. Cascio, W. Goulding,

and M. Maron. 2015. Reptile abundance, but not species richness, increases with

regrowth age and spatial extent in fragmented agricultural landscapes of eastern

Australia. Biological Conservation 184:174-181.

McGarigal, K. and S. A. Cushman. 2002. Comparative evaluation of experimental

approaches to the study of habitat fragmentation effects. Ecological Applications

12:335-345.

McGarigal, K. and W. C. McComb. 1995. Relationships between landscape structure and

breeding birds in the Oregon Coast Range. Ecological Monographs 65:235-260.

Millennium Ecosystem Assessment, 2005. Ecosystems and human well-being. Island Press

Washington, DC.

Miller, J. N., R. P. Brooks, and M. J. Croonquist. 1997. Effects of landscape patterns on

biotic communities. Landscape Ecology 12:137-153.

Mishra, B., O. Tripathi, R. Tripathi, and H. Pandey. 2004. Effects of anthropogenic

disturbance on plant diversity and community structure of a sacred grove in

Meghalaya, northeast India. Biodiversity & Conservation 13:421-436.

MoAC. 2004. Statistical Information on Nepalese Agriculture. His Majesty's Government,

Ministry of Agricultuure and Cooperative. Agribusiness Promotion and Statistics

Division, Singh Durbar, Kathmandu, Nepal.

Moktan, M. R., G. Gratzer, W. H. Richards, T. B. Rai, and D. Dukpa. 2009. Regeneration

and structure of mixed conifer forests under single-tree harvest management in the

western Bhutan Himalayas. Forest Ecology and Management 258:243-255.

Molnar, A., M. France, L. Purdy, and J. Karver. 2011. Community-Based Forest

Management. The Extent and Potential Scope of Community and Smallholder Forest

Management and Enterprise.

Morrison, M. L., B. Marcot, and W. Mannan. 2006. Wildlife-habitat relationships: concepts

and applications. Island Press.

Mortelliti, A., S. Fagiani, C. Battisti, D. Capizzi, and L. Boitani. 2010. Independent effects of

habitat loss, habitat fragmentation and structural connectivity on forest‐dependent

birds. Diversity and Distributions 16:941-951.

125

Moura, N. G., A. C. Lees, C. B. Andretti, B. J. Davis, R. R. Solar, A. Aleixo, J. Barlow, J.

Ferreira, and T. A. Gardner. 2013. Avian biodiversity in multiple-use landscapes of

the Brazilian Amazon. Biological Conservation 167:339-348.

Muggeo, V. 2012. segmented: segmented relationships in regression models with

breakpoints/changepoints estimation: R package. See: http://cran. r-project.

org/web/packages/segmented/index. html.

Muggeo, V. M. 2008. Segmented: an R package to fit regression models with broken-line

relationships. R news 8:20-25.

Nagendra, H. 2007. Drivers of reforestation in human-dominated forests. Proceedings of the

National Academy of Sciences 104:15218-15223.

Nagendra, H. 2008. Do parks work? Impact of protected areas on land cover clearing.

AMBIO: A Journal of the Human Environment 37:330-337.

Nagendra, H. and Y. Gokhale. 2008. Management regimes, property rights, and forest

biodiversity in Nepal and India. Environ Manage 41:719-733.

Nagendra, H., 2002. Tenure and forest conditions: community forestry in the Nepal Terai.

Environmental Conservation, 29(04).

Nagendra, H., M. Karmacharya, and B. Karna. 2005. Evaluating forest management in Nepal:

views across space and time. Ecology and Society 10:24.

Nagendra, H., S. Pareeth, B. Sharma, C. M. Schweik, and K. R. Adhikari. 2008. Forest

fragmentation and regrowth in an institutional mosaic of community, government and

private ownership in Nepal. Landscape Ecology 23:41-54.

Nakagawa, S. and H. Schielzeth. 2013. A general and simple method for obtaining R2 from

generalized linear mixed‐effects models. Methods in Ecology and Evolution 4:133-

142.

Ochoa‐Quintero, J. M., T. A. Gardner, I. Rosa, S. F. Barros Ferraz, and W. J. Sutherland.

2015. Thresholds of species loss in Amazonian deforestation frontier landscapes.

Conservation Biology.

Ojha, H., Persha, L., Chhatre, A., 2009. Community Forestry in Nepal: A Policy Innovation

for Local Livelihoods and Food Security. Working Paper No. W09I-02. International

Food Policy Research Institute, Ann Arbor, Michigan, 34pp.

Oksanen, J. et al., 2011. vegan: Community Ecology Package. R package version 1. 17–6.

Palmer, G. C. and A. F. Bennett. 2006. Riparian zones provide for distinct bird assemblages

in forest mosaics of south-east Australia. Biological Conservation 130:447-457.

126

Pardini, R., Bueno, A. d. A., Gardner, T. A., Prado, P. I. and Metzger, J. P. 2010. Beyond the

fragmentation threshold hypothesis: regime shifts in biodiversity across fragmented

landscapes. PLoS ONE, 5(10), e13666.

Pattison, R. R., C. M. D'Antonio, T. L. Dudley, K. K. Allander, and B. Rice. 2011. Early

impacts of biological control on canopy cover and water use of the invasive saltcedar

tree (Tamarix spp.) in western Nevada, USA. Oecologia 165:605-616.

Paudel, N. S., P. Budhathoki, and U. R. Sharma. 2007. Buffer zones: New frontiers for

participatory conservation. Journal of Forest and Livelihood 6:44-53.

Peh, K. S. H., J. d. Jong, N. S. Sodhi, S. L. H. Lim, and C. A. M. Yap. 2005. Lowland

rainforest avifauna and human disturbance: persistence of primary forest birds in

selectively logged forests and mixed-rural habitats of southern Peninsular Malaysia.


Persha, L., H. Fischer, A. Chhatre, A. Agrawal, and C. Benson. 2010. Biodiversity

conservation and livelihoods in human-dominated landscapes: forest commons in

South Asia. Biological Conservation 143:2918-2925.

Piana, R. P. and S. J. Marsden. 2014. Impacts of cattle grazing on forest structure and raptor

distribution within a neotropical protected area. Biodiversity and Conservation:1-14.

Pinheiro, J. C. and D. M. Bates. 2000. Mixed effects models in S and S-PLUS. Springer.

Pizo, M.A., Simao, I., Galetti, M., 1997. Daily variation in activity and flock size of two

parakeet species from southeastern Brazil. The Wilson Bulletin, 109(2): 343-348.

Pokharel, B. K., P. Branney, M. Nurse, and Y. B. Malla. 2007. Community forestry:

Conserving forests, sustaining livelihoods and strengthening democracy. Journal of

Forest and Livelihood 6:8-19.

Politi, N., M. Hunter Jr, and L. Rivera. 2012. Assessing the effects of selective logging on

birds in Neotropical piedmont and cloud montane forests. Biodiversity and


Porter-Bolland, L., E. A. Ellis, M. R. Guariguata, I. Ruiz-Mallén, S. Negrete-Yankelevich,

and V. Reyes-García. 2012. Community managed forests and forest protected areas:

An assessment of their conservation effectiveness across the tropics. Forest Ecology

and Management 268:6-17.

Poudyal, L. P., P. B. Singh, and S. Maharajan. 2008. Bengal Florican Houbaropsis

bengalensis in Nepal: an update. Birding Asia 10: 43 47.

Powell, L. A., J. D. Lang, D. G. Krementz, and M. J. Conroy. 2005. Use of radio-telemetry to

reduce bias in nest searching. Journal of Field Ornithology 76:274-278.

127

Powlesland, R., J. Knegtmans, and A. Styche. 2000. Mortality of North Island tomtits

(Petroica macrocephala toitoi) caused by aerial 1080 possum control operations,

1997-98, Pureora Forest Park. New Zealand Journal of Ecology:161-168.

Pulliam, H. R. 1988. Sources, sinks, and population regulation. American naturalist:652-661.

R CoreTeam, R. 2014. A language and environment for statistical computing. R Foundation

for Statistical Computing, Vienna, Austria. <http:// www.R-project.org/>. ISBN 3-

900051-07-0.

Radford, J. Q. and A. F. Bennett. 2007. The relative importance of landscape properties for

woodland birds in agricultural environments. Journal of Applied Ecology 44:737-747.

Radford, J. Q., A. F. Bennett, and G. J. Cheers. 2005. Landscape-level thresholds of habitat

cover for woodland-dependent birds. Biological Conservation 124:317-337.

Raman, T. 2001. Effect of Slash‐and‐Burn Shifting Cultivation on Rainforest Birds in

Mizoram, Northeast India. Conservation Biology 15:685-698.

Rodewald, A. D., L. J. Kearns, and D. P. Shustack. 2013. Consequences of urbanizing

landscapes to reproductive performance of birds in remnant forests. Biological


Rodrigues, A. S., S. J. Andelman, M. I. Bakarr, L. Boitani, T. M. Brooks, R. M. Cowling, L.

D. Fishpool, G. A. da Fonseca, K. J. Gaston, and M. Hoffmann. 2004. Effectiveness

of the global protected area network in representing species diversity. Nature

428:640-643.

Sagar, R. and J. S. Singh. 2004. Local plant species depletion in a tropical dry deciduous

forest of northern India. Environmental Conservation 31:55-62.

Sah, S.P., Sharma, C.K., Sehested, F., 2002. Possible role of the soil in the Sissoo forest

(Dalbergia sissoo, Roxb.) decline in the Nepal terai. In: W.J. Horst et al. (Editors),

Plant Nutrition. Developments in Plant and Soil Sciences. Springer Netherlands, pp.

930-931.

Sapkota, I. P., M. Tigabu, and P. C. Odén. 2009. Spatial distribution, advanced regeneration

and stand structure of Nepalese Sal (Shorea robusta) forests subject to disturbances of

different intensities. Forest Ecology and Management 257:1966-1975.

Sapkota, I. P., M. Tigabu, and P. C. Odén. 2010. Changes in tree species diversity and

dominance across a disturbance gradient in Nepalese Sal (Shorea robusta Gaertn. f.)

forests. Journal of Forestry Research 21:25-32.

128

Saunders, D. 1990. Problems of survival in an extensively cultivated landscape: the case of

Carnaby's cockatoo Calyptorhynchus funereus latirostris. Biological


Saunders, D. A., R. J. Hobbs, and C. R. Margules. 1991. Biological consequences of

ecosystem fragmentation: a review. Conservation Biology 5:18-32.

Schmiegelow, F. K. and M. Mönkkönen. 2002. Habitat loss and fragmentation in dynamic

landscapes: avian perspectives from the Boreal forest. Ecological Applications

12:375-389.

Schneider, N. A. and M. Griesser. 2009. Influence and value of different water regimes on

avian species richness in arid inland Australia. Biodiversity and Conservation 18:457-

471.

Sekercioglu, C. H. 2002. Effects of forestry practices on vegetation structure and bird

community of Kibale National Park, Uganda. Biological Conservation 107:229-240.

Sekercioglu, C. H., P. R. Ehrlich, G. C. Daily, D. Aygen, D. Goehring, and R. F. Sandi. 2002.

Disappearance of insectivorous birds from tropical forest fragments. Proc Natl Acad

Sci U S A 99:263-267.

Shahabuddin, G. and R. Kumar. 2006. Influence of anthropogenic disturbance on bird

communities in a tropical dry forest: role of vegetation structure. Animal


Shahabuddin, G. and R. Kumar. 2007. Effects of extractive disturbance on bird assemblages,

vegetation structure and floristics in tropical scrub forest, Sariska Tiger Reserve,

India. Forest Ecology and Management 246:175-185.

Sharma, U. R. 1990. An overview of park-people interactions in Royal Chitwan National

Park, Nepal. Landscape and Urban Planning 19:133-144.

Shrestha, U. B., S. Shrestha, P. Chaudhary, and R. P. Chaudhary. 2010. How representative is

the protected areas system of Nepal? A gap analysis based on geophysical and

biological features. Mountain Research and Development 30:282-294.

Shyamsundar, P., Ghate, R., 2011. Rights, Responsibilities and Resources: Examining

Community Forestry in South Asia. Working Paper, SANDEE, 59-11: 1-24.

Singh, B. and D. Chapagain. 2006. Trends in forest ownership, forest resources tenure and

institutional arrangements: are they contributing to better forest management and

poverty reduction

Sinha, D.R., 2011. Betrayal or'Business as Usual'? Access to Forest Resourecs in the Nepal

Terai. Environment and History, 17: 433-460.

129

Smith, A. C., L. Fahrig, and C. M. Francis. 2011. Landscape size affects the relative

importance of habitat amount, habitat fragmentation, and matrix quality on forest

birds. Ecography 34:103-113.

Sodhi, N. S. and P. R. Ehrlich. 2010. Conservation biology for all. Oxford University Press

Oxford, United Kingdom.

Sofaer, H. R., T. S. Sillett, S. I. Peluc, S. A. Morrison, and C. K. Ghalambor. 2012.

Differential effects of food availability and nest predation risk on avian reproductive

strategies. Behavioral Ecology:ars212.

Springate-Baginski, O., Dev, O.P., Yadav, N.P., Soussan, J., 2003. Community Forest

Management in the Middle Hills of Nepal: The Changing Context. Journal of Forest

and Livelihood, 3(1): 5-20.

Straede, S., G. Nebel, and A. Rijal. 2002. Structure and floristic composition of community

forests and their compatibility with villagers' traditional needs for forest products.

Biodiversity & Conservation 11:487-508.

Suarez-Rubio, M., S. Wilson, P. Leimgruber, and T. Lookingbill. 2013. Threshold responses

of forest birds to landscape changes around exurban development. PLoS One

8:e67593.

Sundriyal, R. and E. Sharma. 1996. Anthropogenic pressure on tree structure and biomass in

the temperate forest of Mamlay watershed in Sikkim. Forest Ecology and


Symonds, M. R. and A. Moussalli. 2011. A brief guide to model selection, multimodel

inference and model averaging in behavioural ecology using Akaike’s information

criterion. Behavioral Ecology and Sociobiology 65:13-21.

Tantipisanuh, N., Gale, G.A., 2013. Representaion of thretened vertebrates by a protected

area system in southeast Asia: The importance of non-forest habitats. The raffles

bulletin of zoology: National University of Singapore, 1: 359–395.

Tasker, E. M. and R. A. Bradstock. 2006. Influence of cattle grazing practices on forest

understorey structure in north‐eastern New South Wales. Austral ecology 31:490-502.

Taylor, R. S., S. J. Watson, D. G. Nimmo, L. T. Kelly, A. F. Bennett, and M. F. Clarke. 2012.

Landscape‐scale effects of fire on bird assemblages: does pyrodiversity beget

biodiversity? Diversity and Distributions 18:519-529.

Tewksbury, J. J., D. J. Levey, N. M. Haddad, S. Sargent, J. L. Orrock, A. Weldon, B. J.

Danielson, J. Brinkerhoff, E. I. Damschen, and P. Townsend. 2002. Corridors affect

130

plants, animals, and their interactions in fragmented landscapes. Proceedings of the

National Academy of Sciences 99:12923-12926.

Tews, J., U. Brose, V. Grimm, K. Tielbörger, M. Wichmann, M. Schwager, and F. Jeltsch.

2004. Animal species diversity driven by habitat heterogeneity/diversity: the

importance of keystone structures. Journal of Biogeography 31:79-92.

Thapa, S. and D. S. Chapman. 2010. Impacts of resource extraction on forest structure and

diversity in Bardia National Park, Nepal. Forest Ecology and Management 259:641-

649.

Thinh, V. T., P. F. Doherty, and K. P. Huyvaert. 2012. Effects of different logging schemes

on bird communities in tropical forests: a simulation study. Ecological Modelling

243:95-100.

Timilsina, N., Ross, M.S., Heinen, J.T., 2007. A community analysis of sal (Shorea robusta)

forests in the western Terai of Nepal. Forest Ecology and Management, 241(1-3):

223-234.

Touihri, M., M.-A. Villard, and F. Charfi. 2014. Cavity-nesting birds show threshold

responses to stand structure in native oak forests of northwestern Tunisia. Forest

Ecology and Management 325:1-7.

Trzcinski, M. K., L. Fahrig, and G. Merriam. 1999. Independent effects of forest cover and

fragmentation on the distribution of forest breeding birds. Ecological Applications

9:586-593.

Tscharntke, T., A. M. Klein, A. Kruess, I. Steffan‐Dewenter, and C. Thies. 2005. Landscape

perspectives on agricultural intensification and biodiversity–ecosystem service

management. Ecol Lett 8:857-874.

van den Berg, L. J., J. M. Bullock, R. T. Clarke, R. H. Langston, and R. J. Rose. 2001.

Territory selection by the Dartford warbler ( Sylvia undata) in Dorset,

England: the role of vegetation type, habitat fragmentation and population size.


Vergara, P. M. and J. J. Armesto. 2009. Responses of Chilean forest birds to anthropogenic

habitat fragmentation across spatial scales. Landscape Ecology 24:25-38.

Vergara, P. M. and P. A. Marquet. 2007. On the seasonal effect of landscape structure on a

bird species: the thorn-tailed rayadito in a relict forest in northern Chile. Landscape

Ecology 22:1059-1071.

131

Villard, M. A., M. K. Trzcinski, and G. Merriam. 1999. Fragmentation effects on forest birds:

relative influence of woodland cover and configuration on landscape occupancy.

Conservation Biology 13:774-783.

Villela, D. M., M. T. Nascimento, L. E. O. Aragão, and D. M. Da Gama. 2006. Effect of

selective logging on forest structure and nutrient cycling in a seasonally dry Brazilian

Atlantic forest. Journal of Biogeography 33:506-516.

Webb, E. L. and R. N. Sah. 2003. Structure and diversity of natural and managed sal (

Shorea robusta Gaertn. f.) forest in the Terai of Nepal. Forest Ecology and


White, A., Martin, A., 2002. Who Owns the World's Forests? Forest Tenure and Public

Forests in Transition, Forest Trend, Washington, D.C.: 2-29.

Whitehorne, I., M. Harrison, N. A. Mahony, P. Robinson, A. Newbury, and D. J. Green.

2011. Effects of Cattle Grazing on Birds in Interior Douglas-Fir (Pseudotsuga

Whittaker, R. H. 1972. Evolution and measurement of species diversity. Taxon:213-251.

Whittaker, R. H. 1977. Evolution of species diversity in land communities. In Evolutionary

Biology. Vol. 10 pp. 1–67. Edited by M.K. Hecht and B.W.N.C. Steere. Plenum

Press, New York.

Whittaker, R. J. and J. M. Fernández-Palacios. 2007. Island biogeography: ecology,

evolution, and conservation. Oxford University Press.

Wiens, J. A. 1992. The ecology of bird communities. Cambridge University Press.

Wiersma, Y. F. and D. L. Urban. 2005. Beta diversity and nature reserve system design in the

Yukon, Canada. Conservation Biology 19:1262-1272.

Wikramanayake, E. et al., 2004. Designing a Conservation Landscape for Tigers in Human-

Dominated Environments. Conservation Biology 18:839-844.

Williams, S. E. and J. Middleton. 2008. Climatic seasonality, resource bottlenecks, and

abundance of rainforest birds: implications for global climate change. Diversity and

Distributions 14:69-77.

Williams, S. E., H. Marsh, and J. Winter. 2002. Spatial scale, species diversity, and habitat

structure: small mammals in Australian tropical rain forest. Ecology 83:1317-1329.

World Bank. 2012. World Development Indicators 2012. World Bank Publications.

World Commission on Forests and Sustainable Development.1999. Our forests our future.

Report of the World Commission on Forests and Sustainable Development.

Cambridge University Press, Cambridge, England.

132

World Resource Institute. 1997.The Last Frontier Forests: Ecosystems and Economies on the

Edge. Washington, DC.pp54.

World Resource Institute. 2005. World resources 2005: the wealth of the poor – managing

ecosystems to fight poverty. World Resources Institute, Washington, DC, USA.

WWF. 2014. Living planet report. Species,spaces,people and places. WWF International

Gland,Switzerland.

Zuckerberg, B. and W. F. Porter. 2010. Thresholds in the long-term responses of breeding

birds to forest cover and fragmentation. Biological Conservation 143:952-962.

Zurita, G. A. and M. I. Bellocq. 2012. Bird assemblages in anthropogenic habitats:

identifying a suitability gradient for native species in the Atlantic forest. Biotropica

44:412-419.

Zuur, A. F., E. N. Ieno, and C. S. Elphick. 2010. A protocol for data exploration to avoid

common statistical problems. Methods in Ecology and Evolution 1:3-14.

133

APPENDICES

Appendix A

Table A.1 Total number of individuals and their respective food guilds and habitat groups,

and relative percentage of species observation across sites in lowland tropical landscapes: CF

= Community forest, SF = State forest, PA = Protected area.

Species CF SF PA Food guild Habita

t group

Sightin

g%

Abbot's Babbler Malacocincla abbotti 5

Ins Fs 2

Alexandrine Parakeet Psittacula eupatria 26 34 29 fru-gra Fg 25

Ashy Drongo Dicrurus leucophaeus 1

Ai Fg 1

Ashy Minivet Pericrocotus divaricatus

2 Fgi Fg 1

Ashy Woodswallow Artamus fuscus 16

Ai Fg 2

Asian Koel Eudynamys scolopacea

1 1 Ins Fg 1

Asian Palm Swift Cypsiurus balasiensis 9

Ai Fg 4

Asian Pied Starling Sturnus contra

2

ins-fru-gra Fg 1

Bar-winged Flycatcher-shrike Hemipus picatus

2 8 Fgi Fes 4

Black Bulbul Hypsipetes leucocephalus

1 fru-ins-nec Fs 1

Black Drongo Dicrurus macrocercus 56 61 11 Si Fg 50

Black Eagle Ictinaetus malayensis 1

Ca Oc 1

Black-crested Bulbul Pycnonotus melanicterus 21 14 32 fru-ins Fs 24

Black-hooded Oriole Oriolus xanthornus 142 132 153 fru-nec-ins Fg 98

Black-naped Monarch Hypothymis azurea 6 2 4 Si Fes 4

Black-rumped Flameback Dinopium benghalense 33 47 12 Bgi Fg 46

Black-winged Cuckooshrike Coracina melaschistos 1 6 2 Fgi Fes 4

Blue-bearded Bee-eater Nyctyornis athertoni 3 2

Si Fes 4

Blue-eared Barbet Megalaima australis 2

Fru Fs 1

Blue-throated Barbet Megalaima asiatica 9 1

Fru Fg 6

Blyth's Leaf Warbler Phylloscopus reguloides 2

Fgi Fes 1

Bronzed Drongo Dicrurus aeneus 28 9 12 Si Fes 20

Brown Shrike Lanius cristatus

1

ins-ca Fes 1

Chestnut-bellied Nuthatch Sitta castanea 176 126 171 Bgi Fg 67

Chestnut-headed Bee-eater Merops leschenaulti 2

Si Fg 1

Chestnut-tailed Starling Sturnus malabaricus 100 44 5 fru-ins-nec Fg 13

Collared Falconet Microhierax caerulescens 3 2

Si Oc 4

Common Hawk Cuckoo Hierococcyx varius 4 1 6 Fgi Fg 9

Common Hoopoe Upupa epops

1 Ins Fg 1

134

Common Iora Aegithina tiphia 2 1 6 Fgi Fes 4

Common Myna Acridotheres tristis 12 30

Omn Fg 8

Common Tailorbird Orthotomus sutorius 30 43 31 Fgi Fg 52

Common Woodshrike Tephrodornis pondicerianus 9 5 29 Fgi Fg 9

Crested Serpent Eagle Spilornis cheela 6 3 6 Ca Fg 13

Crow-billed Drongo Dicrurus annectans 1 1

Si Fs 2

Drongo Cuckoo Surniculus lugubris 3 1 2 Fgi Fs 4

Dusky Warbler Phylloscopus fuscatus 8

Ins Fg 5

Emerald Dove Chalcophaps indica

4 gra-fru Fs 4

Eurasian Collared Dove Streptopelia decaocto 19 6 8 Gra Fg 14

Eurasian Wryneck Jynx torquilla 1

Ins Fg 1

Fulvous-breasted Woodpecker Dendrocopos macei 21 25 18 Bgi Fes 33

Golden-fronted Leafbird Chloropsis aurifrons 110 47 23 nec-ins-fru Fg 37

Golden-spectacled Warbler Seicercus burkii 1

Fgi Fg 1

Great Tit Parus major 62 72 134 Fgi Fg 57

Greater Coucal Centropus sinensis 3 2

Omn Fg 4

Greater Flameback Chrysocolaptes lucidus 2

Bgi Fg 1

Greater Racket-tailed Drongo Dicrurus paradiseus 92 57 81 Si Fes 78

Greater Yellownape Picus flavinucha 2

2 Bgi Fes 4

Green Bee-eater Merops orientalis 12 3 7 Si Fg 4

Green-billed Malkoha Phaenicophaeus tristis 1 2 5 Fgi Fs 4

Greenish Warbler Phylloscopus trochiloides 26 28 16 Fgi Fg 36

Grey-backed Shrike Lanius tephronotus

1

ins-ca Fg 1

Grey-bellied Tesia Tesia cyaniventer

1

Ins Fg 1

Grey-capped Pygmy Woodpecker Dendrocopos

canicapillus 9 2 17 Bgi Fg 1

Grey-headed Canary Flycatcher Culicicapa

ceylonensis 104 64 53 Si Fg 80

Grey-headed Woodpecker Picus canus 7 2 40 Bgi Fs 21

Grey-sided Bush Warbler Cettia brunnifrons

1

Fgi Fg 1

Himalayan Bulbul Pycnonotus leucogenys 14 7 1 fru-ins-nec Fg 6

Himalayan Flameback Dinopium shorii 27 14 48 Bgi Fs 38

Indian Roller Coracias benghalensis

1

Ins-ca Fg 1

Hume's Warbler Phylloscopus humei

3

Fgi Fes 1

Indian Cuckoo Cuculus micropterus 3 1

Fgi Fg 4

Indian Grey Hornbill Ocyceros birostris 1

Fru Fg 1

Indian Peafowl Pavo cristatus 3

15 Omn Fg 6

Jungle Babbler Turdoides striatus 264 215 236 Ins Fg 79

Jungle Myna Acridotheres fuscus

2 fru-gra-nec Fg 1

Jungle Owlet Glaucidium radiatum 6 3 13 ins-ca Fg 15

135

Large Cuckooshrike Coracina macei 61 40 30 ins-fru Fes 48

Large Woodshrike Tephrododornis gularis 35 18 18 Fgi Fes 8

Large-billed Crow Corvus macrorhynchos 8 3 20 Omn Fg 17

Lesser Racket-tailed Drongo Dicrurus remifer 2 1 1 Si Fs 4

Lesser Yellownape Picus chlorolophus 49 44 66 Bgi Fes 54

Lineated Barbet Megalaima lineata 20 8 3 Fru Fes 21

Little Pied Flycatcher Ficedula westermanni

1 Fgi Fs 1

Long-tailed Shrike Lanius schach 4 5 1 ins-ca Fg 7

Olive-back Pipit Anthus hodgsoni 3

1 Ins Fg 2

Orange-breasted Green pigeon Treron bicincta 6

9 Fru Fs 3

Orange-headed Thrush Zoothera citrina 5 2 22 ins-fru Fes 11

Oriental Magpie Robin Copsychus saularis 1 1

ins-nec Fg 2

Oriental Pied Hornbill Anthracoceros albirostris 23 5

Fru Fs 7

Oriental Turtle Dove Streptopelia orientalis 2

Gra Fg 1

Oriental White-eye Zosterops palpebrosus 22 5

nec-ins Fg 4

Osprey Pandion haliaetus

2

Pis Oc 1

Pale-chinned Flycatcher Cyornis poliogenys 14 10 1 Ai Fg 18

Pallas's Fish Eagle Haliaeetus leucoryphus

1 Pis Oc 1

Plain Prinia Prinia inornata

30 Fgi Fg 6

Plum-headed Parakeet Psittacula cyanocephala 150 84 52 fru-gra Fes 45

Puff-throated Babbler Pellorneumru ficeps 1 10

Ins Fg 4

Purple Sunbird Nectarinia asiatica

2 nec-ins Fg 1

Red Collared Dove Streptopelia tranquebarica

1

Gra Fg 1

Red Junglefowl Gallus gallus 32 11 24 Gra Fes 36

Red-billed Blue Magpie Urocissa erythrorhyncha 4 12 20 Omn Fes 10

Red-throated Flycatcher Ficedula parva 104 74 73 Si Fg 78

Red-vented Bulbul Pycnonotus cafer 68 41 33 fru-ins-nec Fg 22

Red-whiskered Bulbul Pycnonotus jocosus 4 8

fru-ins-nec Fg 4

Rose-ringed Parakeet Psittacula krameri 307 350 206 fru-gra Fg 17

Rosi Minivet Pericrocotus roseus

6 Fgi Fg 3

Rufous Treepie Dendrocitta vagabunda 113 83 93 Omn Fg 17

Scaly Thrush Zoothera dauma 1 1 3 ins-fru Fs 4

Scarlet Minivet Pericrocotus flammeus 184 175 105 Fgi Fes 59

Short-toed Snake Eagle Ciraetus gallicus

1

Ca Oc 1

Sirkeer Malkoha Phaenicophaeus leschenaultii

2 Ins Fg 1

Slaty-blue Flycatcher Ficedula tricolor 1 1

Ins Fg 2

Slaty-headed Parakeet Psittacula himalayana

2 fru-gra Fg 1

Slender-billed Oriole Oriolus tenuirostris 1 1 1 fru-nec-ins Fg 3

Small Minivet Pericrocotus cinnamomeus 26 65 48 Fgi Fg 17

Spangled Drongo Dicrurus hottentottus 178 126 135 nec-si Fes 79

136

Spotted Dove Streptopelia chinensis 95 42 13 Gra Fg 52

Spotted Owlet Athene brama

1

Si Fg 1

Streak-throated Woodpecker Picus xanthopygaeus

9 Ins Fg 1

Striped Tit Babbler Macronous gularis 2 1

Fgi Fes 3

Thick-billed Warbler Acrocephaluss aedon 4 1 1 Fgi Fes 4

Tickell's Leaf Warbler Phylloscopus affinis 2

Fgi Fes 1

Tree Pipit Anthus trivialis

4

ins-gra Fes 1

Velvet-fronted Nuthatch Sitta frontalis 54 36 20 Bgi Fes 24

Verditer Flycatcher Eumyias thalassina 3

3 Si Fg 3

White-bellied Drongo Dicrurus caerulescens 48 23 27 Si Fes 46

White-rumped Shama Copsychus malabaricus 25 25 13 Ins Fes 29

White-throated Fantail Rhipidura albicollis

4 Ins Fg 1

White-throated Kingfisher Halcyon smyrnensis 2

ins-ca Fg 2

Yellow-bellied Warbler Abroscopus superciliaris

1

Ins Fg 1

Yellow-crowned Woodpecker Dendrocopos

mahrattensis

3 Bgi Fg 2

Yellow-footed Green Pigeon Treron

phoenicopterus

1 6 Fru Fg 2

Zitting Cisticola Cisticola juncidis 1 Ins Fg 1

Food guilds: ins = insectivore, si = sallying insectivore, ai = aerial insectivore, fgi = foliage gleaning insectivore,

bgi = bark gleaning insectivore, fru = frugivore, gra = granivore, nec = nectarivore, , ca = carnivore, omn =

omnivore, pis = piscivore. Habitat groups: fg = forest specialist, fes = forest edge specialist, fs = forest

specialist, oc = open country species

137

Table A.2 Top ten contributing species to dissimilarities between management tenures using

SIMPER analysis based on Bray-Curtis dissimilarity.

Species

Mean abundance Contribution to

dissimilarity % Community Forest Protected Area

Rose-ringed Parakeet Psittacula krameri 7.21 5.58 4.54

Jungle Babbler Turdoides striatus 6.20 6.69 3.97

Plum-headed Parakeet Psittacula cyanocephala 3.08 2.88 3.74

Chestnut-bellied Nuthatch Sitta castanea 4.07 5.04 3.73

Great Tit Parus major 2.25 5.22 3.50

Scarlet Minivet Pericrocotus flammeus 4.27 3.24 3.44

RufousTreepie Dendrocitta vagabunda 4.53 4.27 2.75

Spotted Dove Streptopelia chinensis 3.59 1.01 2.67

Spangled Drongo Dicrurus hottentottus 5.64 5.05 2.56

Golden-fronted Leafbird Chloropsisaurifrons 2.79 1.00 2.47

Species Community Forest State Forest Contribution to

dissimilarity %






Black Drongo Dicrurus macrocercus 2.36 4.52 3.36



Spotted Dove Streptopelia chinensis 3.59 2.36 2.98

Black-hooded Oriole Oriolus xanthornus 5.20 6.99 2.85

Species Protected Area State Forest Contribution to

dissimilarity %





Black Drongo Dicrurus macrocercus 0.71 4.52 3.45

Great Tit Parus major 5.22 2.95 3.39




Black-hooded Oriole Oriolus xanthornus 6.09 6.99 2.48

138

Appendix B

Table B.1 Summary statistics of habitat characteristics (± SE) across sites in grazing, logging

and lopping disturbances in lowland tropical landscapes

Variables Grazed Light

grazed

Logged Light

logged

Lopped Light lopped

Total tree density

(ha-1)

623.1

(24.67)

643.9

(28.73)

639.9

(34.14)

627.6

(19.25)

645.1

(33.47)

623.5

(20.49)

Large tree density

(ha-1)

65.7

(5.06)

109.2

(7.59)

51.4

6(4.07)

117.9

(6.26)

67.8

(7.45)

103.2

(5.95)

Total basal area

(m2ha-1)

87.5

(5.68)

110.0

(5.59)

85.1

(5.86)

110.2

(5.36)

79.7

(5.55)

114.3

(5.19)

Large tree basal area

(m2ha-1)

38.5

(3.93)

72.6

(6.20)

29.4

(3.15)

77.7

(5.44)

37.4

(4.41)

70.2

(5.63)

Tree canopy cover

(%)

48.4

(1.49)

56.7

(1.40)

46.5

(1.47)

57.7

(1.26)

47.5

(1.55)

56.7

(1.32)

Shrub density

(ha-1)

2714.3

(218.95)

3101.4

(191.81)

2609.7

(240.23)

3159.8

(172.23)

2492.2

(213.19)

3249.0

(191.83)

Shrub cover

(%)

28.9

(1.79)

33.5

(1.71)

28.7

(2.02)

33.2

(1.52)

26.7

(1.73)

34.8

(1.66)

Herbaceous cover

(%)

28.6

(1.65)

34.2

(1.80)

31.5

(1.56)

31.2

(1.89)

30.6

(1.65)

31.9

(1.83)

139

Table B.2 Results of analysis of variance (ANOVA) for the vegetation characteristics across

the different disturbance types.

Disturbance types Variables F value P value

Grazing

Tree density 0.80 0.37

Large tree density 2.71 0.10

Basal area 0.84 0.36

Large basal area 2.25 0.14

Tree canopy cover 2.13 0.15

Shrub density 0.01 0.91

Shrub cover 0.57 0.45

Herbaceous cover 8.55 0.00

Logging









Lopping









Grazing*logging








140


Grazing*Lopping Tree density 0.01 0.93





141

Table B.2(continued)

Disturbance types Variables F value P value




Logging*Lopping









Grazing*Logging*Lopping









142

Table B.3 Results of analysis of variance (ANOVA) for overall species richness, total


insectivore

Disturbance types variable F-value P-value

Grazing Overall species richness 10.29 0.00

Average abundance 3.05 0.08

Bark gleaner species richness 0.89 0.35

Bark gleaner abundance 3.28 0.07

Foliage gleaner species richness 10.80 0.00

Foliage gleaner abundance 4.91 0.03

Logging Overall species richness 2.74 0.07






Lopping Overall species richness 27.60 0.00






Grazing * logging Overall species richness 0.66 0.42






Grazing * lopping Overall species richness 0.43 0.51






Logging * lopping Overall species richness 0.63 0.53

143






Grazing * logging * lopping Overall species richness 0.64 0.42






144

Table B.4 Model averaged coefficient and sum of Akaike weights (∑ωі) (fixed effects) for

each variable based on AIC. Significant estimates are in bold.

Response

group

Variable Coefficient SE Lower

CI

Upper

CI

∑ ωі

Overall bird

species

richness

Large trees 0.06 0.02 0.03 0.12 0.84

Forest extent (2.5 km radius) 0.14 0.03 0.05 0.17 1.00

Tree canopy cover 0.08 0.03 0.04 0.14 0.95

Shrub density 0.06 0.03 0.00 0.10 0.59


Road density -0.02 0.03 -0.06 0.04 0.25

Average

abundance

Large trees 0.06 0.02 0.02 0.09 0.99



Shrub density 0.02 0.01 -0.02 0.05 0.32

Forest extent (0.5 km radius) 0.02 0.01 -0.04 0.04 0.30

Road density 0.03 0.04 -0.06 0.09 0.32

Bark-gleaning

insectivores

richness

Large trees 0.16 0.06 0.02 0.27 0.82



Shrub density 0.12 0.06 -0.02 0.22 0.66


Road density -0.02 0.07 -0.16 0.11 0.25

Bark-gleaning

insectivores

abundance

Large trees 0.21 0.05 0.10 0.31 1.00


Tree canopy cover 0.14 0.04 -0.07 0.13 0.97

Shrub density 0.03 0.03 0.05 0.22 0.26


Road density 0.02 0.11 -0.20 0.23 0.26

Foliage-

gleaning

insectivores

richness

Large trees 0.10 0.06 0.00 0.24 0.58



Shrub density 0.13 0.06 -0.02 0.25 0.57


Road density -0.07 0.07 -0.21 0.08 0.27

Foliage-

gleaning

Large trees 0.09 0.05 -0.01 0.17 0.63


145

insectivores

abundance


Shrub density 0.10 0.04 0.03 0.20 0.79


Road density -0.08 0.10 -0.27 0.11 0.24

146

Table B.5 AIC, ∆value and Akaike weights (ωі) for models of overall species richness, total


insectivore

Response group Model AIC ∆ AIC ωі

Overall species

richness

Large trees density + forest extent (2.5 km radius) +

shrub density + tree canopy cover + forest extent (0.5

km radius)

621.10 0.00 0.27


tree canopy cover + forest extent (0.5 km radius)

621.45 0.34 0.23


shrub density + tree canopy cover + forest extent (0.5

km radius) + road density

623.42 2.32 0.08


tree canopy cover + forest extent (0.5 km radius) +

road density

623.49 2.39 0.08


shrub density + tree canopy cover

623.74 2.64 0.07

Forest extent (2.5 km radius) + tree canopy cover +

forest extent (0.5 km radius) + shrub density

624.19 3.09 0.06

147

Table B.5 (continued)


Overall

abundance

Large trees density + Forest extent (2.5 km radius) +

Tree canopy cover

969.5

1

0.00 0.39


Tree canopy cover + road density

971.0

4

1.53 0.18


Tree canopy cover + shrub density

971.7

7

2.25 0.13


Tree canopy cover + Forest extent (0.5 km radius)

971.7

9

2.28 0.12


shrub density + Tree canopy cover + road density

973.3

1

3.8 0.06


road density + Tree canopy cover + Forest extent (0.5

km radius)

973.3

6

3.85 0.06


shrub density + Tree canopy cover + Forest extent

(0.5 km radius)

974.0

9

4.57 0.04

148



Bark gleaning

insectivores

richness


Forest extent (0.5 km radius) + Tree canopy cover

405.49 0.00 0.18

Forest extent (2.5 km radius) + Forest extent (0.5 km

radius) + Tree canopy cover + Shrub density

406.36 0.87 0.13

Forest extent (2.5 km radius) + Large trees density +

Forest extent (0.5 km radius) + Shrub density

406.89 1.40 0.09



406.91 1.43 0.09


radius) + road density + Tree canopy cover+ Shrub

density

407.68 2.19 0.06

Forest extent (2.5 km radius) + Tree canopy cover +

Shrub density

408.03 2.54 0.05


Tree canopy cover + Shrub density

408.19 2.71 0.05



Shrub density + road density

408.60 3.11 0.04

149



Bark gleaning

insectivores

abundance



726.00 0.00 0.26


Tree canopy cover

726.53 0.53 0.20


Tree canopy cover+ Shrub density

727.63 1.63 0.12


Tree canopy cover + road density

727.28 2.28 0.08



Shrub density

728.50 2.49 0.08



road density

728.85 2.85 0.06


Tree canopy cover + Shrub density + road density

729.95 3.94 0.04

150



Foliage gleaning

insectivores

richness



395.19 0.00 0.17


Shrub density

395.46 0.27 0.15


Tree canopy cover + Shrub density + road density

396.62 1.43 0.08


Shrub density

396.72 1.53 0.08



397.17 1.98 0.06



Shrub density

397.38 2.19 0.06



397.55 2.36 0.05



398.03 2.84 0.04

151



Foliage gleaning

insectivores

abundance


Tree canopy cover + Shrub density + Forest extent (0.5

km radius)

859.47 0.00 0.24


Shrub density + Forest extent (0.5 km radius)

860.23 0.76 0.16



860.66 1.19 0.13


Shrub density

860.99 1.53 0.11



861.11 1.64 0.11

Forest extent (2.5 km radius) + Large trees density +

Tree canopy cover

861.94 2.47 0.07


radius) + Tree canopy cover

862.37 2.90 0.06


radius) + road density + Tree canopy cover + Shrub

density + Large trees density

862.76 3.29 0.05


radius) + road density + Tree canopy cover+ Shrub

density

863.21 3.74 0.04

152

Table B.6 AIC, ∆value and Akaike weights (Ѡі) for interaction models of overall species

richness, total average abundance, species richness and abundance of bark-gleaning and

foliage-gleaning insectivore


Overall species

richness

Forest extent (2.5 km radius) + disturbance +

disturbance * Forest extent (2.5 km radius)

623.7 0.00 0.61

Forest extent (2.5 km radius) + Forest extent (0.5

km radius) + disturbance + disturbance * Forest

extent (2.5 km radius) + disturbance * Forest

extent (0.5 km radius)

626 2.22 0.2


Forest extent (2.5 km radius)

626.76 2.99 0.14


disturbance * Forest extent (0.5 km radius)

628.87 5.09 0.05

Overall

abundance



extent (0.5 km radius) + disturbance * Forest


1336.79 0.00 0.56




1337.26 0.47 0.44

Bark- gleaning

insectivores

richness


km radius)

405.72 0.00 0.45




407.06 1.34 0.23


km radius) + disturbance

407.66 1.94 0.17



extent (2.5 km radius) + disturbance *Forest


409.29 3.57 0.08




409.84 4.12 0.06

153



Bark- gleaning

insectivores

abundance

Forest extent (2.5 km radius) + Forest extent

(0.5 km radius) + disturbance + disturbance *


898.63 0.00 0.69



Forest extent (0.5 km radius) + disturbance *


900.22 1.6 0.31

Foliage- gleaning

insectivores

richness


(0.5 km radius) + disturbance + Forest extent

(0.5 km radius) * disturbance

384.43 0.00 0.24


(0.5 km radius) + disturbance

384.55 0.11 0.23




384.58 0.15 0.22



(2.5 km radius) * disturbance + Forest extent


385.02 0.58 0.18


(0.5 km radius)

386.23 1.79 0.10


Forest extent (2.5 km radius) * disturbance

391.27 6.84 0.10

Foliage- gleaning

insectivores

abundance



Forest extent (0.5 km radius) + disturbance *


1044.98 0.00 0.96




1051.45 6.46 0.04

154

Table B. 7 Mean values of different disturbance types between sites classified as lightly and

heavily disturbed in lowland Terai forests.

Summary statistics

Lightly disturbed sites Heavily disturbed sites

Number of

lopped tree

branches

Number of

dung piles

Number of

cut stumps

Number of

lopped tree

branches

Number

of dung

piles

Number of

cut stumps

Mean 1.7 2.0 0.5 46.5 12.5 11.1

Standard Error 0.4 0.4 0.1 3.4 0.9 0.9

Standard Deviation 2.8 3.3 0.9 24.8 6.5 6.4

Range 9.6 11.0 5.0 79.8 28.0 25.5

Count 60.0 60.0 60.0 52.0 52.0 52.0

155

Table B.8 Correlation matrix of explanatory variables measured. Coefficients in bold shows

pairs highly correlated variables.

Explanatory variables 1 2 3 4 5 6 7 8 9

1. Forest extent (2.5 km radius) 1

2.Road density 0.12 1

3. Large tree density 0.33 0.07 1

4. Total basal area 0.36 0.13 0.88 1

5.Large trees basal area 0.35 0.08 0.89 0.93 1

6. Tree canopy cover 0.43 0.14 0.30 0.30 0.21 1

7.Shrub density 0.12 -0.15 0.22 0.08 0.15 0.17 1

8.Shrub cover 0.13 -0.15 0.35 0.19 0.21 0.15 0.57 1

9.Forest extent (0.5 km radius) 0.14 0.02 0.24 0.23 0.12 0.27 0.30 0.26 1

156

Appendix C

Table C.1 AIC, Δ value and Akaike weights (ωі) for models of overall estimated species,

frugivore, foliage-gleaning insectivore and sallying insectivore

Response group Model AIC ΔAIC ωі

All birds Forest extent +water body 252.85 0.00 0.19

Forest extent +rainfall +water body +forest extent

*rainfall 252.97 0.11 0.18

Forest extent +rainfall +water body 253.85 1.00 0.12

Forest extent +water body +rainfall +disturbance

+disturbance*forest extent +forest extent*rainfall

253.86 1.00 0.12

Disturbance +forest extent +water body+

disturbance*forest extent

254.62 1.76 0.08

Forest extent +rainfall 255.02 2.16 0.07

Disturbance +forest extent +water body 255.49 2.64 0.05

Disturbance +forest extent +rainfall +water body

+forest extent*rainfall

255.98 3.13 0.04

Forest extent +rainfall +forest extent*rainfall 256.49 3.63 0.03

Disturbance +forest extent +rainfall +water body 256.81 3.96 0.03

Disturbance +forest extent +water body +rainfall

+disturbance +disturbance*forest extent

256.97 4.11 0.02

Disturbance +forest extent+ rainfall 257.72 4.87 0.02

157

Table C.1 (continued)

Response groups Model AIC ΔAIC ωі

Frugivores Forest extent + water body 152.59 0.00 0.54

Forest extent +rainfall +water body 155.33 2.74 0.14

Disturbance + forest extent +water body 155.33 2.74 0.14

Disturbance + forest extent +water body

+disturbance*forest extent

157.67 5.07 0.04

Forest extent +rainfall+ water body +forest extent

*rainfall

157.86 5.26 0.04

Disturbance + forest extent+ rainfall + water body 158.32 5.73 0.03

Water body 159.37 6.77 0.02

Disturbance + water body 160.6 8.01 0.01

Rainfall + water body 160.78 8.18 0.01

158



Foliage-gleaning

insectivores

Forest extent +rainfall + forest extent*rainfall 146.56 0.00 0.39

Disturbance + forest extent +rainfall +forest

extent * rainfall

147.12 0.56 0.3

Forest extent +rainfall+ water body + forest

extent * rainfall

148.89 2.33 0.12

Disturbance +forest extent +rainfall +disturbance

* forest extent+ forest extent * rainfall

149.93 3.37 0.07

Disturbance +forest extent +rainfall+ water body

+ forest extent * rainfall

149.93 3.37 0.07

Disturbance +forest extent +rainfall+ water body

+ disturbance * forest extent + forest extent *

rainfall

152.88 6.31 0.02

Forest extent + rainfall + water body 154.49 7.93 0.01

159



Sallying

insectivores

Null 148.75 0.00 0.28

Forest 150.84 2.09 0.1

Water body 150.85 2.10 0.1

Rain 150.92 2.17 0.1

Disturbance 150.97 2.23 0.09

Forest +water body 152.96 4.21 0.03

Rain +water body 152.96 4.21 0.03

Forest + rain 153.12 4.37 0.03

Disturbance + forest + disturbance * forest 153.23 4.48 0.03

Disturbance + water body 153.25 4.51 0.03

Disturbance + rain 153.27 4.53 0.03

Disturbance + forest 153.34 4.59 0.03

Forest + rain + forest * rain 153.71 4.96 0.02

Forest + rain + water + forest * rain 154.38 5.63 0.02

Forest + rain + water 154.86 6.11 0.01

Disturbance + rain + water body 155.44 6.69 0.01

Disturbance + forest + water body 155.69 6.94 0.01

160

Table C.2 Model averaged coefficients across the 95% confidence set of models for all

explanatory variables

Response variables Explanatory variables Estimate Std error z-value p-value

All birds

Forest extent 0.18 0.03 4.90 <0.001

Water body 0.09 0.37 2.36 0.018

Rainfall 0.03 0.04 0.62 0.536

Disturbance -0.02 0.03 0.63 0.529

Forest extent*Rainfall 0.52 0.03 1.80 0.071

Disturbance *Forest extent 0.07 0.03 1.93 0.040

Frugivores

Forest extent 0.20 0.07 2.71 0.007

Water body 0.21 0.06 3.28 0.001

Rainfall 0.01 0.08 0.16 0.877

Disturbance 0.00 0.07 0.07 0.948

Forest extent *Rainfall -0.05 0.06 0.72 0.474


Foliage gleaners

Forest extent 0.55 0.10 5.02 <0.001

Water body 0.08 0.09 0.78 0.437

Rainfall 0.44 0.11 3.66 0.01

Disturbance 0.11 0.08 1.38 0.168

Forest extent*Rainfall -0.30 0.10 2.94 0.01


Sallying insectivores

Forest extent -0.05 0.08 0.53 0.597

Water body -0.06 0.09 0.61 0.543

Rainfall 0.05 0.08 0.53 0.600

Disturbance 0.02 0.08 0.25 0.802

Forest extent *Rainfall -0.11 0.07 1.45 0.148


161

Figure C.1 Species accumulation curves of all 28 studied landscapes based on Chao2/ICE

estimated richness.

Date post:	26-Jul-2020
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EFFECTS OF HABITAT EXTENT AND FOREST DISTURBANCE ON …372720/s... · site-level vegetation...

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