A Case Study on the Genetic Diversity of Small Mammal Communities in the Lower Kinabatangan Floodplain on Borneo, Sabah, Malaysia

UNIVERSITY OF VETERINARY MEDICINE HANNOVER Institute of Zoology

Effects of habitat fragmentation in a tropical rain forest ecosystem: a case study on the genetic diversity of small mammal communities in the Lower Kinabatangan floodplain on Borneo, Sabah, Malaysia

THESIS Submitted in partial fulfilment of the requirements of the degree of Doctor of Natural Sciences Doctor rerum naturalium (Dr. rer. nat.)

by Jennifer Brunke Gronau/Leine

Hannover, Germany 2020 Scientific Supervision: Prof. Dr. Ute Radespiel, University of Veterinary Medicine Hannover, Germany

1st Evaluation: Prof. Dr. Ute Radespiel Institute of Zoology University of Veterinary Medicine Hannover, Germany

2nd Evaluation: PD Dr. Christian Roos Primate Genetics Laboratory German Primate Centre Göttingen, Germany

Day of oral exam: 02.11.2020

Sponsorship: German Academic Exchange Service (DAAD) Calenberg Grubenhagensche Landschaft

Table of contents

Table of contents i Previously published excerpts of this thesis iii Summary iv Zusammenfassung vii

Chapter 1 General introduction 1 1.1 Southeast Asia, a leading hotspot of biodiversity 1 1.2 The Bornean rain forest, a vanishing ecosystem under pressure 2 1.3 Impacts of deforestation and forest fragmentation 3 1.4 Molecular markers in conservation genetics 6 1.5 Lower Kinabatangan Wildlife Sanctuary 7 1.6 Bornean small mammal community 9 1.7 Aims and Hypotheses 11 1.8 References 12

Chapter 2 Messing about on the river: The role of geographic barriers in shaping the genetic structure of Bornean small mammals in a fragmented landscape 19 2.1 Introduction 20 2.2 Material and Methods 22 2.3 Results 26 2.4 Discussion 30 2.5 References 36 2.6 Appendix 40

Chapter 3 Dispersal and genetic structure in a tropical small mammal, the Bornean tree shrew (Tupaia longipes), in a fragmented landscape along the Kinabatangan River, Sabah, Malaysia 51 3.1 Introduction 52 3.2 Materials and Methods 54 3.3 Results 58 3.4 Discussion 64 3.5 References 69 3.6 Appendix 72

i Chapter 4 General discussion 76 3.1 Composition of non-volant small mammal assemblages at the Lower Kinabatangan River 76 4.1.1 Impact of forest fragmentation on community composition 77 4.1.2 Presence and impact of invasive species 82 4.2 Impacts of landscape structures along the Kinabatangan River on species distribution, gene flow and population structure 83 4.2.1 Impacts of anthropogenic landscape modifications 83 4.2.2 The effectiveness of corridors in fragmented landscapes 87 4.2.3 Impacts of the Kinabatangan River and riverine structures 88 4.2.4 Genetic inference of historical landscape changes 92 4.2.5 Impacts of sex-specific dispersal patterns 93 4.3 Conservation implications 94 4.4 References 96 4.5 Appendix 103

Acknowledgements 133

ii Previously published excerpts of this thesis

Chapter 2 published as Jennifer Brunke, Ute Radespiel, Isa-Rita Russo, Michael W. Bruford, Benoit Goossens (2019) Messing about on the river: the role of geographic barriers in shaping the genetic structure of Bornean small mammals in a fragmented landscape. Conservation Genetics 20: 691 – 704. It was first published online 20 March 2019. This article is available at https://doi.org/10.1007/s10592-019-01159-3 with open access. The copyright for this article remains with the authors. It is distributed under the terms of the Creative Commons Attribution 4.0 International License (http://creatIvecommons.org/ licenses/by/4.0/). The study presented in Chapter 2 was designed by JB, UR, MWB, and BG. The data were acquired and analysed by JB, and interpreted by JB and UR. The article was drafted by JB and critically revised by all authors. All authors approved the publication of the final version. The peer-reviewed journal Conservation Genetics is the original place of publication. The format of the chapter has been adapted to the style of this thesis. The numbering of tables and figures differ slightly from the original article. Associated Supplementary Material is available at the DOI stated above. It was included in the chapter as appendix.

Chapter 3 published as Jennifer Brunke, Isa-Rita M. Russo, Pablo Orozco-terWengel, Elke Zimmermann, Michael W. Bruford, Benoit Goossens, Ute Radespiel (2020) Dispersal and genetic structure of a tropical small mammal, the Bornean tree shrew (Tupaia longipes), in a fragmented landscape along the Kinabatangan River, Sabah, Malaysia. BMC Genetics 21: 43. It was first published online 17 April 2020. This article is available at https://doi.org/10.1186/s12863-020-00849-z with open access. The copyright for this article remains with the authors. It is distributed under the terms of the Creative Commons Attribution 4.0 International License (http://creatIvecommons.org/ licenses/by/4.0/). The study presented in Chapter 3 has been designed by JB, UR, EZ, MWB, and BG. The data were acquired by JB and partly by IRR and analysed by JB and partly by PO. The interpretation of the data was done by JB and UR. The article was drafted by JB and critically revised by all authors. All authors approved the publication of the final version. The peer-reviewed journal BMC Genetics is the original place of publication. The format of the chapter has been adapted to the style of this thesis, and the numbering of tables and figures differ slightly from the original article. Associated Supplementary Material is available at the DOI stated above. It was included in the chapter as appendix.

iii Summary

Jennifer Brunke

Effects of habitat fragmentation in a tropical rain forest ecosystem: a case study on the genetic diversity of small mammal communities in the Lower Kinabatangan floodplainon Borneo, Sabah, Malaysia

An exceptionally high degree of biodiversity and endemic species is concentrated in the tropical forests of Southeast Asia. Large-scale anthropogenic deforestation in this region resulted in habitat loss and fragmentation, and causes severe problems for the maintenance of the biodiversity. Effects of habitat fragmentation have been studied comprehensively, but negative consequences are still controversially discussed and cannot be generalised. One reason is that the impacts of habitat fragmentation depend on several species-specific traits, and individuals of different species can be affected in a variety of ways. Knowledge of the ecological plasticity and sensitivity of a wide range of species is therefore essential to evaluate generalised fragmentation effects and to formulate effective conservation and management plans. Tropical small mammals, which occur in highly diverse and species-rich communities, are a suitable model to assess fragmentation effects on species with varying ecological requirements. This thesis investigated population genetic structures in non-volant small mammal communities inhabiting forest fragments in the Lower Kinabatangan floodplain on Borneo. The landscape along the Lower Kinabatangan River is highly degraded by anthropogenic modifications and dominated by oil palm plantations but still retains a high floral and faunal diversity. The impact of landscape modifications has so far been studied for larger animals but not for small mammals of this region. Equally scarce are population genetic studies assessing the influence of the fragmented landscape on population genetic structures to evaluate long-term viability and population persistence. Until now, this has only been conducted for the Bornean elephant and some primate species. Furthermore, in fragmented habitats, biogeographical barriers, such as rivers, might represent an important additive factor restricting migratory capabilities even further, leading to a reduction or interruption of gene flow and an increased genetic differentiation among populations. For example, the Kinabatangan itself represent an important dispersal barrier in orangutans. How the Kinabatangan River and other geographic characteristics (e.g. tributaries, roads) affect the dispersal of non-volant small mammals and what effect the forest fragmentation has on population structures are the questions which hasto be answered in this thesis. In the first part of this thesis the effects of forest fragmentation and geographical barriers, especially the Kinabatangan River, was assessed for the whole small mammal community. First, taxonomic

iv classifications based on mitochondrial cytochrome b (cyt b) and 16S rRNA sequences were conducted to classify all captured specimen correctly. Subsequent investigations on intraspecific molecular variations in cyt b sequences revealed heterogeneous barrier effects of the different landscape features for different taxa. Dispersal and gene flow were shown to be impeded by the Kinabatangan River in squirrels and tree shrews, but not in rats and the moonrat Echinosorex gymnura. Signals of genetic differentiation between populations indicated additional limitations in terrestrial vagility along riverbanks in a subset of species. Finally, to infer the dispersal regime for the study species, the sex-specific representation of cyt b haplotype diversity was evaluated for all species. A clear signal of sex-biased dispersal was only found in two murid species. The influence of habitat fragmentation on population genetic signatures was examined in more detail for one tree shrew species, the Bornean tree shrew (Tupaia longipes), in the second part of this thesis. Mitochondrial DNA sequences (cyt b) were used in combination with nuclear DNA marker (microsatellites) to assess the genetic connectivity and discontinuities in fragmented subpopulations inhabiting forest remnants along both sides of the Kinabatangan River. While gene flow was undisturbed between connected forest fragments, pronounced genetic differentiation and irregular haplotype distribution indicated genetic fragmentation in isolated forests. In particular, largeroil palm plantations seem to act as migration barriers, reducing genetic connectivity between fragmented subpopulations of T. longipes. Signals of reduced dispersal were distinct in males and females alike. In conclusion, the studies confirm a diverse small mammal community within the remaining forest fragments along the Lower Kinabatangan River. However, low abundancies of rare habitat specialists, the overrepresentation of common generalists, the unexpected absence of otherwhise common native species and the presence of invasive commensals indicate ongoing shifts within the Kinabatangan small mammal assemblages. The studies emphasise the importance of continuous habitats to maintain gene flow and preserve genetic diversity in small mammal species inhabiting anthropogenically disturbed landscapes. Certain geographic features were identified as important barriers to dispersal and gene flow in the highly modified landscape along the Lower Kinabatangan River. The Kinabatangan River itself was shown to represent an important factor, which both negatively (in squirrels and in tree shrews) and positively (in murids and E. gymnura) affected migration and gene flow and shaped genetic structures in small mammals. While based on both mitochondrial and nuclear markers, oil palm plantations were proposed as another major factor limiting spatial distribution of individuals in T. longipes and (based on mitochondrial marker) in some other small mammal species, the Kinabatangan road or tributaries seem not to impede dispersal. Barrier width and age were attributed to define the effectiveness of dispersal barriers in small mammals. Finally, habitat fragmentation may have induced shifts in the dispersal patterns of males

v and females. This would explain why this study did not find evidence for sex-biased dispersal neither based on maternally inherited mitochondrial marker (most small mammals) nor with a set of bi- parentally inherited microsatellites (T. longipes). However, sample sizes were often too small to reliably assess effects of plantations and/or other potential barriers (e.g. tributaries or roads) on genetic structures in the majority of the studied species and clearly need future attention.

vi Zusammenfassung

Jennifer Brunke

Auswirkungen von Habitatfragmentierung im Ökosystem tropischer Regenwald: Fallstudie zur genetischen Diversität von Kleinsäugergesellschaften am Kinabatangan Fluss auf Borneo, Sabah, Malaysia

Ein hoher Grad an Biodiversität und endemischen Arten ist in den tropischen Regenwäldern Südost Asiens beheimatet. Großflächige Abholzungen der Wälder in dieser Region haben dazu geführt, dass Habitate dezimiert und fragmentiert wurden, was eine Gefahr für den Erhalt der Artenvielfalt darstellt. Obwohl Fragmentierungseffekte bereits eingängig untersucht wurden, werden deren negative Auswirkungen weiterhin kontrovers diskutiert und können nicht artübergreifend pauschalisiert werden. Ein Grund dafür ist, dass abhängig von artspezifischen Eigenschaften sich Fragmentierungseffekte auf unterschiedliche Art und Weise manifestieren können. Artübergreifende Kenntnisse über ökologische Plastizitäten und Sensitivitäten sind daher notwendig die Auswirkungen von Fragmentierungseffekten genauer einschätzen zu können und basierend darauf effektive Schutzmaßnahmen zu entwickeln. In den tropischen Regenwäldern sind Kleinsäuger in äußerst diversen Artengemeinschaften organisiert, wodurch sich diese besonders gut eignenum Fragmentierungseffekte an einer Vielzahl von Arten mit spezifischen ökologischen Ansprüchen genauer zu untersuchen. In dieser Arbeit wurden dazu populationsgenetische Strukturen von Kleinsäugern aus dem Gebiet des Kinabatangan Flusses auf Borneo untersucht. Die Wälder entlang des Kinabatagans sind durch großflächige Rodungen und Umwandlung zu Palmölplantagen stark dezimiert wurden und bestehen zum Teil nur noch als vereinzelte Fragmente inmitten einer hochgradig anthropogen geprägten Landschaft. Dennoch ist in den verbliebenen Waldfragmenten noch eine hohe Biodiversität an Pflanzen und Tierarten vorhanden. Bisher wurden Fragmentierungseffekte hauptsächlichan Großsäuger untersucht, sind für die Kleinsäuger dieser Region jedoch noch völlig unbekannt. Im Speziellen wurden populationsgenetische Studien zu Langzeitauswirkungen der Habitatfragmentierung größtenteils vernachlässigt und wurden bislang nur am Borneo- Zwergelefanten und einigen Primaten durchgeführt. In fragmentierten Landschaften stellen geographische Merkmale oftmals zusätzliche Barrieren dar, die die Verteilung von Individuen und somit den Genfluss zwischen Population beeinträchtigen können. Für Orang-Utans beispielsweise konnte der Kinabatangan Fluss als eine solche Barriere identifiziert werden. Inwieweit der Kinabatangan und andere geographische Strukturen (z.B. Nabenflüsse, Straßen) sowie die

vii ausgeprägte Waldfragmentierung Einfluss auf die Verteilung der dort vorkommenden Kleinsäuger und deren Populationsstrukturen nimmt soll in dieser Arbeit geklärt werden. Im ersten Teil der Arbeit wurden dazu die Kleinsäuger zunächst anhand von Cytochrome b (cyt b) und 16S rRNA Sequenzen taxonomisch klassifiziert. Anhand von anschließenden molekulargenetischen Untersuchungen artspezifischer Divergenzen entlang der generierten cyt b Sequenzen, konnten schließlich Effekte verschiedener Landschaftsmerkmale als Dispersionsbarriere aufgezeigt werden. Diese waren sehr heterogen ausgeprägt und unterschieden sich je nach Taxon. Besonders in Hörnchen und Spitzhörnchen waren die Migration und der Genfluss zwischen Populationen der beiden Ufer des Kinabatangan eingeschränkt, jedoch weniger stark bei Muriden und dem Rattenigel Echinosorex gymnura. In einigen Arten der untersuchten Kleinsäuger waren zudem genetische Differenzierungen zwischen Populationen innerhalb einer Flussseite vorhanden, die darauf hindeuten, dass eine zusätzliche Einschränkung der terrestrischen Vagilität besteht. Abschließend wurden sexspezifische Ausprägungen und Verteilungsmuster von cyt b Haplotypen genutzt, um artspezifische Migrationsmuster der Kleinsäuger zu identifizieren. Ein eindeutiges sexspezifisches Muster war jedoch nur in Arten der Gattung Maxomys zu finden. Im zweiten Teil dieser Arbeit wurden schließlich Auswirkungen von Habitatfragmentierungen eingehender untersucht, indem populationsgenetische Signaturen des Borneo Langfuß- Spitzhörnchens (Tupaia longipes) erfasst und ausgewertet wurden. Hierzu wurden sowohl mitochondriale (cyt b) Sequenzen analysiert als auch Mikrosatelliten genotypisiert, um genetische Konnektivitäten bzw. Diskonnektivitäten zwischen Subpopulationen zu erfassen, die in den fragmentierten Waldgebieten entlang beider Ufer des Kinabatangans vorkommen. Während der Genfluss zwischen Populationen aus zusammenhängenden Waldfragmenten weitestgehend ungestört war, deuten ausgeprägte genetische Differenzierungen zwischen Populationen und Unregelmäßigkeiten in der Verteilung von cyt b Haplotypen auf eine genetische Fragmentierung in isolierten Fragmenten hin. Besonders limitiert war die genetische Konnektivität zwischen fragmentierten Subpopulationen, wenn diese durch großflächige Palmölplantagen getrennt waren. Einschränkungen im Migrationsvermögen bedingt durch Habitatfragmentierung konnten sowohlin männlichen als auch weiblichen T. longipes festgestellt werden. Zusammenfassend konnte in dieser Arbeit eine diverse Kleinsäugergemeinschaft in den verbleibenden Wäldern entlang des Kinabatangan bestätigt werden. Jedoch deuten geringe Fangzahlen an seltenen Habitatspezialisten, eine Überrepresentation an Habitatgeneralisten, eine unerwartete Abwesenheit von andernorts häufigen Arten sowie das Vorhandensein von invasiven Arten auf mögliche Fluktuationen innerhalb der Kleinsäugergemeinschaften hin. Beide Studien dieser Arbeit betonen die Bedeutsamkeit von zusammenhängenden Waldgebieten für die Aufrechterhaltung des Genflusses zwischen und die genetische Diversität innerhalb von

viii Kleinsäugerpopulationen in fragmentierten Landschaften. Bestimmte geographische Merkmale konnten als bedeutsame Dispersionsbarriere identifiziert werden, die den Genfluss zwischen Subpopulation entlang des Kinabatangans beeinträchtigen. Der Kinabatangan selbst stellt eine geographische Barriere dar, der den Genfluss und die Migration von Individuen zum einen negativ beeinflusst (Hörnchen und Spitzhörnchen), der aber zum anderen populationsgenetische Strukturen in positiver Weise unterstützen und formen kann (Muridae und E. gymnura). Sowohl anhand von mitochondrialen als auch nukleären Markersytemen konnten Palmölplantagen als eine weitere Dispersionsbarriere identifiziert werden, die die räumliche Verteilung von T. longipes und auch von anderen Kleinsäugerarten (identifiziert anhand von mitochondrialen Markern) entlang des Kinabatangan negativ beeinträchtigen können. Die vorhandene Straße und Nebenflüssedes Kinabatangan stellten dagegen keine Dispersionsbarriere dar. Die Effektivität von vorhandenen Barrieren scheint insbesondere von dessen Breite und Alter abhängig zu sein. Darüber hinaus, ist es möglich, dass die Fragmentierung der Wälder Veränderungen in den geschlechtsspezifischen Migrationsmustern vom Männchen und Weibchen hervorgerufen hat, was erklären würde, warum in der vorliegenden Arbeit weder anhand von maternal vererbten mitochondrialen Marker (der meisten Kleinsäuger) noch in bi-parental vererbten Mikrosatelliten (T. longipes) ein geschlechtsspezifischer Einfluss in den Migrationsmustern zu erkennen war. Es ist jedochzu erwähnen, dass für die meisten untersuchten Kleinsäugerarten der Stichprobenumfang zu gering war, um einen Barriereneffekt von Plantagen und anderen Landschaftsmerkmalen (z.B. kleineren Nebenflüssen oder Straßen) eindeutig zu bestätigen. Hierfür sind weiterführende Studien sinnvoll und notwendig, um diese Fragestellungen zweifelsfrei zu klären.

ix Chapter 1 General introduction

1.1 Southeast Asia, a leading hotspot of biodiversity Compared to other ecosystems, tropical rain forests represent the most species-rich terrestrial biome on earth (Myers et al. 2000; Barlow et al. 2018). Its distribution is limited to zones with constant warm-humid climates (about 2000 mm annual precipitation and mean monthly temperatures not < 18°C) and without a pronounced dry season (Eiserhardt et al. 2017). Between the N 23.5° and S 23.5° latitude it can be found in the Neotropics, Africa, and in the Indo-Pacific region (Eiserhardt et al. 2017). Climatic conditions paired with dynamic and complex geological events during the Pleistocene epoch generated conditions ideal for speciation and created several distinct centers of biological diversity particularly in Southeast Asia (Sodhi et al. 2004; Cannon et al. 2009; Metcalfe 2011; Leonard et al. 2015; Husson et al. 2019). As a result, one of the oldest and most diverse rain forests can be found in this region. Encompassing the countries of Myanmar (Burma), Thailand, Vietnam, Cambodia, Laos, Singapore, Malaysia, Indonesia, Brunei, and the Philippines, Southeast Asia forms a mosaic consisting mainly of islands and peninsulas. This mosaic covers only 4% of the planet’s land area, but is home to 20 – 25% of all existing plant and animal species (Woodruff 2010). Furthermore, the world’s highest proportion of endemic vascular plant, bird, and mammal species can be found within this region (Duckworth et al. 2012). Sodih et al. (2004) predicted a biodiversity crises for Southeast Asia in which about 42% of the biodiversity will be lost until 2100 because of drastic habitat reductions. Moreover, Southeast Asia overlaps with four major biogeographic regions, where exceptionally high concentrations of endemic species are confronted with severe levels of habitat degradation, so called “biodiversity hotspots” (Myers et al. 2000; Sodhi et al. 2004; Fig. 1.1). Among these, the “Sundaland hotspot” is outstanding (Fig. 1.1). This region hosts more than 25,000 (15,000 endemic) plant and 1,800 (701 endemic) vertebrate species within highly depleted habitats, making it the most threatened sub-region of Southeast Asia and an important global conservation priority (Myers et al. 2000, Sodhi et al. 2004; Wilcove et al. 2013; Hughes 2017; Verma et al. 2020).

1 Figure 1.1 The four biodiversity hotspots of Southeast Asia, as defined by Sodhi et al. (2004)

1.2 The Bornean rain forest, a vanishing ecosystem under pressure With a total land area of about 74 M ha (Gaveau et al. 2016), Borneo is the largest of the Sundaic islands, and the planet’s third largest. Borneo was originally covered with tropical rain forest, including lowland forests, hill and montane forests, freshwater and peat swamp forests, heath forests (kerangas), and mangrove forests (including nipa palm (Nypa fruticans) forests; Gaveau et al 2014). Borneo’s forests represent one of the world’s most valuable and productive tropical forests (de Bruyn 2014; Corlett 2019). However, large-scale deforestation activities have reduced the rain forests on Borneo considerably. Over the last four decades alone (1973 – 2015), Borneo has lost more than 34% of its original forest extent, and if deforestation continues unabated, most of the forest cover will be lost in 2020 (Gaveau et al. 2016; Wulffraat and Greenwood 2017). This drastic forest loss resulted predominantly from commercial logging activities and conversion to agricultural land. Particularly, lowland and peat swamp forests have been subject to logging and forest conversion (Miettinen et al. 2011; Gaveau et al. 2016). Lowland forests are dominated by dipterocarp trees, commercially valuable timber, which was and still is harvested heavily in Borneo. Reaching its peak between 1980 and 2000, more timber was harvested during these decades from Borneo than from tropical Africa, Central and South America combined (Berry et al. 2010; Gaveau et al. 2014). In the last decade (2000 – 2010), Borneo has lost lowland forest with a deforestation rate of 1.3% per year, the highest worldwide rate of deforestation next to Sumatra (Miettinen et al. 2011; Wilcove et al. 2013). Peat swamp forests experienced even a much higher rate of deforestation (2.8% per year), because they often represent the last major concentration of forests in the lowlands, and modern technologies make them increasingly accessible and attractive to land developers (Miettinen et al. 2011; Wilcove et al. 2013). This alarming trend converts important carbon sinks to carbon sources and escalates the risks of fires within these ecosystems (Page and Hooijer 2016).

2 The expansion and intensification of agriculture is the principle driver for the loss of the remaining forests on Borneo. For that purpose, forests are cleared and converted to plantations with cash crops (Gaveau et al. 2016) such as oil palm (Elaeis guineensis), rubber (Hevea brasiliensis), and pulpwood (Acacia and Eucalyptus). While plantations in other Southeast Asian regions are typically about 100 ha in size, large-scale industrial plantations with an average size of 1,000 ha are frequent on Borneo (Hughes 2018; Meijaard et al. 2018). The total area of industrial plantations reached 9.2 M ha (7.9 M ha oil palm, 1.3 M ha rubber and pulpwood) until 2015, corresponding to 12% of Borneo’s total land area. Especially in Malaysian Borneo, the loss of old-growth and secondary forests (57 – 60% of deforestation within the last decades) was associated with conversion to industry plantations (Abram et al. 2014; Gaveau et al. 2016; Meijaard et al. 2018). It has been shown that the majority of large- scale forest clearances are preceded by the expansion of road networks and that most of deforestation in Borneo occurred in the vicinity of roads (Gaveau et al. 2014; Hughes 2018; Alamgir et al. 2019; Sloan et al. 2019). Road development therefore represents another important cause for forest destruction. Further pressures arise from uncontrolled small-scale agriculture, fires, mining, and the development of dams (Page and Hooije 2016; Wulffraat and Greenwood 2017).

1.3 Impacts of deforestation and forest fragmentation Impacts of deforestation on biodiversity Compared to intact forests, vegetation density and canopy structure is altered in degraded forests and often extremely simplified in monocultures such as oil palm plantations (Sodhi et al. 2004; Meijaard et al. 2018; Corlett 2019; Luke et al. 2020). These structural modifications often cause changes in microclimate conditions (Hardwick et al. 2015) and resource availability and typically coincide with shifts in floral and faunistic community composition and abundances (Sodhi et al. 2004; Wearn et al. 2017, 2019). Especially, disturbance-intolerant and specialized taxa are prone to be replaced by ecologically flexible generalists (Edwards et al. 2014). Such changes in plant community composition can affect the functionality of ecosystems, and the value of the services which they provide can become depleted (Estes et al 2011). Simplified habitat structure and low plant diversity of monocultures usually promote assemblages dominated by a few abundant generalists, non-forest species, invasive species and pests. The majority of forest-dwelling species is absent in oil palm plantations (Fitzherbert et al. 2008; Wilcove and Koh 2010; Wilcove et al. 2013; Savilaakso et al. 2014; Yue et al. 2015; Wearn et al. 2017; Pardo et al. 2018). However, (selectively) logged forest can retain a surprisingly high level of biodiversity, although sometimes in reduced abundances (Berry et al. 2010; Wilcove et al. 2013; Wearn et al. 2017, 2019), and secondary vegetation often reaches high structural complexity and heterogeneity with successional stage (Phillips et al. 2017; Rozendaal et al.

3 2019). A recovery to levels similar to old growth forest often takes 40 – 60 years in terms of plant species richness and about 20 – 40 years in terms of vertebrate richness. Species composition even recovers at a much slower rate (Dunn 2004; Acevedo-Charry and Aide 2019; Rozendaal et al. 2019). Nevertheless, the declining extent of old growth forests demands the protection of degraded and secondary forests to maintain a substantial number of species in the long term (Dunn 2004; Wearn et al. 2017, 2019).

Impacts of forest fragmentation on dispersal and gene flow Dispersal, the permanent movement of individuals to a new location (Greenwood 1980), occurs in all animal and plant species. Since dispersal influences the spatial distribution of organisms and their genes, it has an important impact on population dynamics (Fischer and Lindenmeyer 2007; Broquet and Petit 2009; Cayuela et al. 2018). Anthropogenic generated landscape structures (e.g. plantations (Goossens et al. 2005, 2006b; Villard and Haché 2012; Scriven et al. 2017), roads and highways (Holderegger and Di Giulio 2010; Ascensão et al. 2016; Grilo et al. 2018), canals (Díaz‐Muñoz 2012)) but also natural geographic characteristics, such as waterbodies (Burney and Brumfield 2009; Rocha et al. 2014; Khanal et al. 2018), mountains (Oshida et al. 2006, Machado et al. 2018), ridges and slopes (Russo et al. 2016; Mazoch et al. 2018) or microhabitats (Geffen et al. 2004; Algar et al. 2013), may interrupt habitats and represent impenetrable barriers that constrain dispersal and the distribution of organisms within landscapes (Manel et al. 2003; Storfer et al. 2007; Holderegger and Wagner 2008; Balkenhol et al. 2017). Changes in dispersal patterns are expected to have implications for the genetic structure and the dynamics of populations as well as their persistence. For example, constraints in dispersal may reduce or interrupt gene flow, and populations become genetically more structured and less diverse (Banks et al. 2013; Haddad et al. 2015; Pardini et al. 2017). Especially in populations with a small effective population size, genetic diversity can easily get reducedby inbreeding or genetic drift (Frankham 1996; Ralls et al. 2018). Moreover, reduced levels of genetic diversity may decrease the ability of populations to adapt to environmental changes, raise the risk of local extinction and limit the persistence of populations in the long-term (Frankham 2003; 2005; Ralls et al. 2018).

Considering the vast expansion rate of oil palm plantations throughout Southeast Asia and other parts of the world, and taking into account that oil palm plantations support less forest species than logged forest, they represent the greatest immediate threats to biodiversity (Fitzherbert et al. 2008, Wilcove et al. 2013; Phillips et al. 2017). Oil palm and other monocultures represent unsuitable habitats for most forest species. Where they form part of the landscape matrix, they may act as

4 barriers to animal movements and reduce the amount of available habitats (Fitzherbert et al. 2008; Meijaard et al. 2018; Luke et al. 2020). In the process of natural forest conversion to industrial tree plantations, formerly continuous forests are reduced to smaller patches, creating a fragmented mosaic in which forest remnants are surrounded by a matrix of inhospitable landscape. Moreover, forest loss leads to forest fragments of various sizes, degree of isolation and area/edge ratio. This may decrease the ability of individuals to migrate between sub-populations and to colonise and recolonise patches that underwent stochastic extinctions (Banks et al. 2013; Haddad et al. 2015; Pardini et al. 2017). The resulting declines in effective population size and restrictions in gene flow will change genetic differentiation among and reduce genetic diversity within isolated populations. Induced by this, inbreeding depression can raise the risk of local extinction and jeopardise the long-term persistence of species within fragmented landscapes (Frankham 2005).

Not all species are equally affected by habitat fragmentation and some are at a greater risk than others. In particular, highly specialized and rare species with an inherent low individual abundance and growth rate, a high level of population fluctuation, and low dispersal capacity, are supposedto be negatively affected by habitat fragmentation (Henle et al. 2004). Furthermore, several studies suggest that population dynamics of a species is more prone to be affected by environmental alterations when dispersal strategies are asymmetric, e.g. between sexes (e.g. Stow et al. 2001, Fischer and Lindenmayer 2007; Pierson et al. 2010; Vangestel et al. 2013; Cote et al. 2017). In most mammal and bird species dispersal is sex-biased, with one sex dispersing more frequently and/or further from its natal domain (Greenwood 1980). Constrains in dispersal may cause an immigration deficit of the dispersive and a sex bias towards the philopatric sex in isolated populations. An altered sex ratio may have negative effects on social structure, demography, and microevolution, or lead to local extinction (Amos et al. 2014). However, the ability of a given species to cope with forest fragmentation and the functional connectivity of populations within fragmented landscapes depend on many species-specific traits, e.g. behaviour, social organization, reproduction, vagilityand ecological requirements (Ewers and Didham 2006; Keinath et al. 2017), and impacts on a given species are thus not readily predictable. Knowledge of specific-specific dispersal pattern and the identification of barriers that may impede the exchange of individuals and gene flow among populations are thus of high interest in conservation biology. This thesis will help to provide such knowledge by assessing impacts of landscape structures and habitat fragmentation on genetic diversity, gene flow and population structures for a suite of small mammals with different ecological requirements. Established molecular markers will be used in genetic modelling approaches to achieve this.

5 1.4 Molecular markers in conservation genetics Molecular markers are small sections of the genome that can be used as indicators of genome-wide variation and help to solve relevant questions in conservation genetics (Salgado et al. 2016). However, each marker mirrors only the demographic history of the underlying genomic region, and genetic results can thus vary between molecular markers. Today, to strengthen the reliability of data and to decrease the risk of misleading conclusions, the use of a suite of various markers is recommended (Zink and Barrowclough 2008; Toews and Brelsford 2012; Allendorf 2017). The markers used for this study are discussed in more detail in the following.

Mitochondrial DNA Mitochondrial DNA (mtDNA) is a circular extra-nuclear molecule of about 17 kbp consisting of 13 protein-coding genes, 22 transfer RNAs, two ribosomal RNAs and the control region, which initiates replication and transcription processes. Introns or non-coding spacer sequences between genes are absent. While the structure and arrangement of mitochondrial genes are relatively conserved, nucleotide mutations occur at a ten times higher rate than in single-copy nuclear DNA. This results in relatively high levels of genetic variations both within and between species (Avise et al. 1987; Avise 2009). Furthermore, mtDNA is haploid, maternally transmitted and does not undergo recombination, which means that its inheritance is effectively clonal. Its uniparental transmission clearly representsa disadvantage, because only matrilineal but not patrilineal histories of individuals and populations are reflected (Avise 2009). Nevertheless, the generally high mutation rate and the clonal inheritance allow the assessment of genealogical relationships among individuals, populations, and species, and make mtDNA a valuable and informative marker in phylogeographic and population genetic studies (Avise et al. 1987; Avise 2009; Freeland 2008; Galtier et al. 2009). Another advantage of mtDNA is that it is relatively easy to work with. Firstly, because of its high copy numbers within eukaryotic cells, mtDNA can be readily obtained from tissue and other animal matter. Secondly, due to the conserved arrangement of the mtDNA genes across vertebrates, versatile PCR primers for amplification are available for many taxonomic groups and do not need to be specifically developed for a particular species (Freeland 2008; Galtier et al. 2009). Although less polymorphic than the hypervariable control region, the protein-coding gene cytochrome b (cyt b) is a widely used mtDNA marker to answer phylogenetic and population genetic questions and is probably the best known mitochondrial gene (Esposti et al. 1993; Meyer 1994, Farias et al. 2001). Due to its widespread use, a wide range of universal primers exist, allowing PCR amplification of the cyt b locus in different organisms. Furthermore, a large number of cyt b sequences are available from many previous studies on GenBank (www.ncbi.nlm.nih.gov/genbank) for comparative purposes.

6 Microsatellites Microsatellites (msats) are short non-coding tandem repeat motives of 1 – 6 bp that are scattered throughout the nuclear genome. They are selectively neutral markers with a high allelic variability. Msats mutation rates are higher than in the rest of the genome, ranging from 10 -2 to 10-6 events per locus per generation (Oliveira et al. 2006; Selkoe and Toonen 2006; Freeland 2008). Mutations in msat loci result typically from slippage and proofreading errors during DNA replication that primarily change the number of repeats and may alter the length of the repeat sequence (Schlötterer 2000; Selkoe and Toonen 2006). As a result from high mutation rates and the tendency to either increase or decrease in sequence length, size homoplasy is possible in msats. Msats are thus less suitable for inferring evolutionary processes. However, their high mutation rates result in a high level of locus polymorphism, a suitable trait for inferring relatively recent population genetic events and quantify effects of landscape structure on gene flow and dispersal thereby and the spatial patterns of genetic variation (Freeland 2008; Allendorf 2017). Furthermore are msats biparentally transmitted and codominant. In contrast to the uniparental inheritance of mtDNA, msats reflects gene genealogies of both sexes and allows thus an unbiased assessment of whole population genetic histories. A disadvantage is the tedious and costly development of msat primers, which can rarely be transferred between taxonomic groups, and typically need to be specifically developed for each study species. Nevertheless, cross-species amplification is often possible among closely related species (Barbará et al. 2007).

1.5 Lower Kinabatangan Wildlife Sanctuary With a length of 560 km and a catchment area of 16,800 km2 (about 23% of total land area of Sabah; Fig. 1.2) the river basin of the Kinabatangan River in eastern Sabah is the largest on Borneo (Boonratana 2000a). The Lower Kinabatangan region refers to the lower catchment of the river (the lower 70 – 100 km) and its tributaries (Fig. 1.2), where it meanders through a periodically flooded area (Azmi 1998; Kler 2007). The Lower Kinabatangan floodplain is the largest in Malaysia and is one of the last forested alluvial floodplains in Asia (Kler 2007; Latip et al. 2013, 2015). This region of the Kinabatangan River is known for its diverse habitats such as limestone caves, oxbow lakes, riparian and dipterocarp forests, riverine forest, mangroves and freshwater swamp forests (Azmi 1998; Latip et al. 2013; Pimid et al. 2020). This habitats harbour a remarkably diverse wildlife with over 314 bird, 129 mammal, 101 reptile, 33 amphibian, 100 freshwater fish, and about 2500 plant species being recorded, including many endemic and endangered species like the Bornean orangutan (Pongo pygmeus), proboscis monkey (Nasalis lavartus), Borneo pygmy elephant (Elephas maximus borneensis), Sunda clouded leopard (Neofelis diardi), the rhinoceros hornbill (Buceros rhinoceros) and

7 the oriental darter (Anhinga melanogaster) (Latip et al. 2013, 2015; Abram et al. 2014; Pimid et al. 2020).

Figure 1.2 Kinabatangan River basin, adapted from Ancrenaz (2015)

However, progressive forest degradation and agricultural expansion are a major environmental problem for this region. The lowland forests of the Lower Kinabatangan district were one of the first in Sabah used for commercial timber extraction in the early 1950s. Reaching its peak in the 1970, today virtually all forests of the Lower Kinabatangan region have been logged at least once (Boonratana 2000a). The disappearance of valuable timber promoted the conversion of forest to agriculture along the Lower Kinabatangan in the late 1980s (Latip et al. 2013). The subsequent extensive expansion of oil palm plantations transformed the landscape considerably and resulted in significant forest loss (67% of forest cover lost between 1982 and 2014; DGFC 2018), degradation, and fragmentation. In order to conserve the ecological resources of the Kinabatangan floodplain, the Sabah State Government established a Wildlife Sanctuary along the Lower Kinabatangan River in 2005. The Lower Kinabatangan Wildlife Sanctuary (LKWS) comprises 27,000 ha of highly disturbed forest remnants that are linked to patches of forest reserves, virgin jungle forest reserves and private forests. The sanctuary provides a forested corridor along the lower course of the river that aims to connect the costal mangrove swamps with upriver located dry land forests (Fig. 1.2; Ancrenaz et al. 2004, Goossens et al. 2005; Latip et al. 2015). Nonetheless, significant areas of unprotected forest remain under the threat of ongoing conversion to oil palm plantations (Abram et al. 2014). In order to investigate the influence and interaction of human encroachment to the local floral and faunal wildlife, to better understand species-specific requirements, and to improve the ecological value of the landscape, several studies have been conducted in the LKWS (DGFC 2018). In particular the spatial ecology, habitat use, behaviour and health of various larger animals (e.g Bornean elephant Alfred et al. 2012; Estes et al. 2012; carnivores: Evans et al. 2016; Guharajan et al. 2017;

8 Hearn et al. 2017, 2018a; primates: Boonratana 2000b; Ancrenaz et al. 2004, 2005; Goossens et al. 2006a; Stark et al 2012; Röper et al. 2014; Klaus et al. 2017, 2018; Matsuda et al. 2020; crocodiles: Evans et al. 2016b, 2017) have been the focus of recent studies. However, studies addressing long- term effects of forest destruction and fragmentation on population genetic structures and gene flow are scarce for this landscape. Population genetic studies have been conducted only on the Bornean elephant (Elephas maximus borneensis, Goossens et al. 2016) and some primates (orangutan: Pongo pygmaeus, Goossens et al. 2005, 2006b; Jalil 2006; Jalil et al. 2008; Bruford et al. 2010; long-tailed macaque: Macaca fascicularis, Jalil 2006; Salgado 2010; proboscis monkey: Nasalis larvatus, Jalil 2006; Salgado 2010). Significant declines in genetic diversity and migration induced by anthropogenic modifications of the landscape could be revealed (Goossens et al. 2005, 2006b, 2016), and the Kinabatangan River itself represents an additional important geographic barrier affecting dispersal and gene flow in orangutans (Goossens et al. 2005; Jalil et al 2008).

1.6 Bornean small mammal community Borneo has a rich mammalian fauna with more than 240 known species. With more than 80 known species, small non-volant mammals (non-flying mammals, adult weight < 5 kg, Jayaraj et al. 2012) represent one of the largest and most diverse mammalian group on Borneo (Phillipps and Phillipps 2016). Generally, Bornean non-volant small mammals refer to various species from the taxa Eulipotyphla (gymnures and shrews), Scandentia (tree shrews), and Rodentia (rats and mice, squirrels, porcupines). Nearly half of these species are endemic to Borneo (Table 1.1) and locally they form species-rich and abundant assemblages with up to 22 different species (Nakagawa et al. 2006; Wells et al. 2007; Bernard et al. 2009; Charles and Ang 2010; Cusack 2011; Khalid and Grafe 2017; Chapman et al. 2018, 2019; Mohd-Azlan et al. 2019). For squirrels and tree shrews, Borneo represents a center of diversity, and together with murids they represent the vast majority of Bornean small mammals (Phillipps and Phillipps 2016, Table 1.1). Small mammal assemblages exploit all habitat strata and are characterised by various degrees of niche specificity and climbing activity (Wells et al. 2006). While, for example, gymnures and shrews are nocturnal invertebrate feeders, and active on the ground, rodents contain nocturnal (murids, (flying) squirrels, and porcupines) and diurnal (squirrels) species, which are found in different habitat strata from terrestrial (murids, squirrels and porcupines), scansorial (murids, squirrels) to arboreal (murids and squirrels). Next to specialised forest species, known invasives (i.e. murids such as Rattus spp.) are present within this diverse group. Tree shrews, however, are predominantly scansorial omnivores with a diurnal activity pattern, the only exception being the pen-tailed tree shrew (Ptilocercus lowii), a nocturnal and arboreal species (Emmons 2000; Phillipps and Phillipps 2016).

9 These life style and ecological distinctions make small mammal assemblages attractive to study the effects of forest disturbances and fragmentation on a variety of species with different ecological requirements. However, most studies on small mammals in Borneo have been focusing on responses of assemblages to forest disturbances such as logging or forest conversion (e.g. Wells et al. 2007; Nakagawa et al. 2006; Bernard et al. 2009; Chapman et al. 2018). The impact of forest fragmentation on Bornean small mammals is much less explored and was largely measured by changes in community composition and abundances (e.g. Charles and Ang 2010; Khalid and Grafe 2017; Mohd- Azlan et al. 2019). Long term consequences of fragmentation on the genetic diversity and gene flow among small mammal populations are mainly known from temperate and other regions with profound anthropogenic interference (e.g. Sommer 2003; Trizio et al. 2005; Redeker et al. 2006; Gauffre et al. 2008; Taylor et al. 2011; Balkenhol et al. 2013; Ćosić et al. 2013; Stephens et al. 2013; Fietz et al 2014; Kierepka et al. 2016; Russo et al. 2016; Bani et al.2017; Otero-Jiménez et al. 2018) but have been largely neglected in Bornean small mammals. Small mammals are an important key factor for the ecological functioning of tropical habitats. They provide services as pollinators, seed predators and seed dispersers of many plant and tree species, including Fagaceae, dipterocarps or figs (Shanahan et al. 2001; Wells and Bagchi 2005; Wells et al. 2009; Corlett 2017), and contribute to forest regeneration, stability and resilience (Phillipps and Phillipps 2016). Some plants even interact with small mammals on a mutualistic basis as observed for pitcher plants (Nepenthes spp.). Attracted by extrafloral nectar, rats and tree shrews deposit their faeces or (when drowning) serve as nitrogen source to the plant themselves (Clarke et al. 2009; Chin et al. 2010; Wells et al. 2011; Bauer et al. 2015). In addition, small mammals are important prey items for carnivores (Rajaratnam et al. 2007; Chua et al. 2016; Hearn et al. 2018b; Hood et al. 2019) or raptors (Puan et al. 2011; Phillipps and Phillipps 2009; Saufi et al. 2020). Furthermore, they regulate arthropod and earthworm populations as invertebrate predators (Ewers et al 2015). Therefore, knowledge about long-term effects of habitat loss and fragmentation on small mammal genetic diversity and connectivity is necessary for an effective conservation managementand landscape planning. As a result of forest degradation, 40% of Borneo’s small non-volant mammals suffer from declining population trends, and about 20% are listed on the IUCN (International Union for Conservation of Nature) Red List of Threatened Species either as Endangered, Vulnerable, or Near Threatened (IUCN 2020). High proportions of collapsing populations are in particular distinct in rodents and tree shrews (Table 1.1).

10 Table 1.1 Bornean non-volant small mammal taxa and the number of known species, number of endemic species, and number of species with an IUCN Red List classification at least as near threatened and/or with a decreasing population trend.

Order Family Species Endemics Threatened Decreasing Rodentia Muridae Rats and Mice 27 11 3 5 Sciuridae Squirrels 36 14 10 18 Hystricidae Procupines 3 1 0 1 Scandentia Ptilocercidae Pen-tailed tree shrews 1 0 0 1 Tupaiidae Tree shrews 9 7 0 8 Eulipotyphla Erinaceidae Gymnures 2 0 0 0 Soricidae Shrews 8 5 1 1

1.7 Aims and Hypotheses Small mammals play an essential role in regulating ecological processes in tropical rain forests ecosystems. A decline in biodiversity, due to deforestation, has been observed in Borneo, but the genetic consequences of forest fragmentation and other landscape barriers in Bornean small non- volant mammals are not fully understood. Therefore, the general aim of this thesis is to assess the impacts of habitat fragmentation on small mammal communities in the Lower Kinabatangan floodplain with various genetic modelling approaches. Different predictions for the impact of habitat fragmentation on genetic diversity, gene flow and structure will be tested and the effect of various geographic barriers (aquatic and terrestrial) will be evaluated (1) by partially sequencing the mitochondrial cytochrome b (cyt b) gene for all species with sufficient samples, and (2) by additionally genotyping nuclear microsatellites for a selected model species.

In the first study (Chapter 2) the importance of natural (Kinabatangan River and its tributaries) and man-made geographic structures (roads, oil palm plantations) as barriers to dispersal and gene flow is evaluated based on genetic variations in mitochondrial cyt b sequences of many different small mammals. Geographic barriers have previously been shown to reduce or interrupt gene flow, and separated populations show a larger genetic differentiation than expected under pure isolationby distance. Constraints on migration across the Kinabatangan River are thus predicted to result in low gene flow, and population genetic differentiation should be increased between the two riversides. Furthermore, it can be expected that forest fragments isolated or dissected by roads or oil palm plantation should show larger genetic discontinuities and a higher genetic differentiationthan continuous forests. Due to variations in species dispersal capacities, landscape features do not act as barriers to gene flow for all species equally, and it can be hypothesised that dispersal constraints across the Kinabatangan River are negatively correlated with species-specific swimming propensities.

11 The second study (Chapter 3) employs mitochondrial cyt b sequences and a set of nuclear microsatellite loci to evaluate in detail the impact of habitat fragmentation in comparison to the Kinabatangan River on gene flow, dispersal and the resulting genetic structures in one selected model species, the endemic Bornean tree shrews (Tupaia longipes). If relatively recent habitat fragmentation along the Kinabatangan River restricts migration in T. longipes, the genetic structure and genetic differentiation should be stronger between isolated forest fragments compared to those that are connected.

In both studies, genetic structures of males and females are analysed separately to assess differences in dispersal patterns and to improve the understanding of the respective dispersal regime, which is unknown for most study species. Dissimilarities in the dispersal behaviour of males and females are predicted to evoke differences in the genetic structure between sexes. Moreover, landscape barriers are predicted to have a stronger impact on the sex dispersing more frequently and/or over larger spatial distances.

1.8 References

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18 Chapter 2 Messing about on the river: The role of geographic barriers in shaping the genetic structure of Bornean small mammals in a fragmented landscape

Published 2019 in Conservation Genetics Volume 20, Issue 4: 691-704, https://doi.org/10.1007/s10592-019-01159-3 by Jennifer Brunke1, Ute Radespiel1, Isa-Rita Russo2, Michael W. Bruford2,3, Benoit Goossens2,3,4,5

1 University of Veterinary Medicine Hannover, Institute of Zoology, Bünteweg 17, 30559 Hannover, Germany 2 Organisms and Environment Division, Cardiff University, School of Biosciences, Sir Martin Evans Building, Museum Avenue, Cardiff CF10 3AX, United Kingdom 3 Sustainable Places Research Institute, Cardiff University, 33 Park Place, Cardiff CF10 3BA, United Kingdom 4 Sabah Wildlife Department, Wisma Muis, 88100 Kota Kinabalu, Sabah, Malaysia 5 Danau Girang Field Centre, c/o Sabah Wildlife Department, Wisma Muis, 88100 Kota Kinabalu, Sabah, Malaysia

Landscape features may restricting dispersal and gene flow, and increase demographic isolation among sub-populations. In addition, landscape features may represent potential dispersal barriers depending on species vagility. To predict the persistence of populations and to formulate adequate conservation measures it is essential to understand the ability of species to transverse landscape barriers. Using population genetic techniques we assessed the importance of physical barriers along the Kinabatangan River for a suite of non-volant small mammals. Cytochrome b sequence variation was examined for each of the 19 species sampled across both riverbanks. Haplotype networks and molecular variance analyses indicated contrasting patterns of genetic isolation between riversidesfor different taxa. Genetic isolation between riversides ranged from moderate to complete in tree shrews and squirrels, whereas no isolating effect could be detected in murids and gymnures. Although genetic divergence between forest fragments on the same side of the river could only be studied in a subset of six species, the results suggest an additional dispersal barrier for two of these studied species. While barrier effects of a paved road and tributaries could not be verified, large oil palm plantations seem to have disrupted gene flow in these species. Furthermore, the findings suggest higher genetic connectivity on the more continuously forested compared to the more fragmented riverside, and underline the importance of forest corridors as essential conservation

19 measures to maintain genetic diversity in a fragmented landscape such as that along the Kinabatangan River.

Keywords: small mammals, geographic barriers, dispersal, landscape fragmentation, gene flow, Borneo

2.1 Introduction Dispersal is the primary behavioural mechanism determining gene flow within and among populations and has substantial effects on the genetic structure. Its restriction leads to a reduction or interruption of gene flow, a decreased genetic diversity within and increased genetic differentiation among populations (Broquet and Petit 2009). Especially in populations with a low effective population size, increased inbreeding can easily occur and cause inbreeding depression, i.e. may decrease the fitness of individuals (Freeland 2008). In the long term, the evolutionary potential of the population may be lost, and the risk of extinction increases (Frankham 2005). As a consequence, the ability of species to cope with habitat changes and reduced landscape connectivity determines the distribution and viability of species (Garant et al. 2007). The use of genetic data has suggested an important role of geographic landscape features such as rivers in shaping the ecological and genetic variation of species (Geffen et al. 2004; Trizio et al. 2005; Coulon et al. 2006). Geographic features may compound the effect of fragmentation, restricting dispersal capabilities even further. Furthermore, in anthropogenically modified forest landscapes structural conditions often differ strongly from the original habitat and canopy gaps openup. Numerous species have limited tolerance to such structural variation and are unable to cross these barriers, resulting in dispersal limitations and genetic changes within and among populations (McDonald and St. Clair 2004; Heidinger et al. 2013; Edwards et al. 2014; Janecka et al. 2016; Fahrig 2017). Therefore, assessing the effects of geographic features on the distribution of species has become essential to inform conservation measures that typically aim to improve the genetic connectivity between subpopulations (Prevedello and Vieira 2010). Tropical small mammals form highly diverse and structured ecological communities and differ greatly in their life history traits and ecological requirements (Wells et al. 2007), thus making a good model system to assess the importance of geographical barriers on patterns of genetic diversity for species with different ecological requirements. The Kinabatangan River in north-eastern Borneo is the longest river in Sabah, Malaysia. Large-scale logging activities and habitat conversion in this highly productive floodplain has created a mosaic of fragmented forest and oil palm plantations. Embedded in this mosaic, forest connectivity has been

20 disrupted by landscape features such as river tributaries, plantations and a paved road. Goossens et al. (2006) demonstrated that orangutan dispersal is negatively affected by fragmentation in this landscape, leading to decreased gene flow and reduced genetic diversity. In addition, the Kinabatangan River itself represents a major barrier for dispersal and has an important role in shaping orangutan genetic structure (Goossens et al. 2005; Jalil et al. 2008). However, such knowledge is still lacking for Bornean small mammal species. Recent studies in other regions of the world have demonstrated the importance of rivers as barriers to gene flow (Oshida et al. 2011; Rocha et al. 2011; Ćosić et al. 2013), and verified the strong effect of anthropogenically modified landscapes on dispersal capabilities and genetic connectivity in small mammals (Gerlach and Musolf 2000; Trizio et al. 2005; Stephens et al. 2013; Bani et al. 2017). But most studies investigating the effect of geographic barriers on dispersal and genetic structure have mainly focused on a single species (Patton et al. 1994; Colombi et al. 2010; Nicolas et al. 2011; Ćosić et al. 2013; Russo et al. 20016), or few species from the same genus (Kennis et al. 2011; Oshida et al. 2011; Rocha et al. 2014). This study is to our knowledge the first investigating the influenceof different landscape features on migration and population connectivity in a comprehensive community of Bornean small mammals. Non-volant small mammals represent one of the largest and most diverse mammalian groups on Borneo and comprise various species of Rodentia, Scandentia, and Eulipotyphla (Payne and Francis 2007; Phillipps and Phillipps 2016). Due to the scarcity of available information on their dispersal (e.g. Emmons 2000; Wells et al. 2006a, 2008a, 2008b; Nakagawa et al. 2007; Shadbolt and Ragai 2010), nothing is known about the influence of geographic and man-made landscape features on their migration and genetic structuring. However, species affinities to water and swimming abilities should be crucial for maintaining gene flow across geographic features such as large rivers. Constraints on dispersal due to the presence of the Kinabatangan River are thus predicted to lead to a decrease in gene flow between populations from either side, and increased population genetic differentiation should exist between the two riversides. On the other hand, gene flow across the river can be expected in species which have an affinity to water and good swimming capabilities such as murid rodents (Evans et al. 1978; Cook et al. 2001; Nicolas and Colyn 2006; Santori et al. 2008). This has, for example, been confirmed in various Neotropical murid rodents (e.g. Patton et al. 1994; Colombi et al. 2010; Roratto et al. 2014). Furthermore, it can be expected that isolated or dissected forest fragments (e.g. by road or plantation) should show larger genetic discontinuities and a higher genetic differentiation than fragments connected by forest corridors. Moreover, invasive and generalist species (e.g. Rattus spp., plantain squirrels) are known to be more tolerant to anthropogenic habitat modifications (Nakagawa et al. 2006; Francis 2008; Phillipps and Phillipps 2016), and therefore

21 predicted to be less strongly affected by forest fragmentation than strictly forest dependent rodents and tree shrews (Emmons 2000; Francis 2008; Phillipps and Phillipps 2016). The aim of this study is to test the influence of the Kinabatangan River and its tributaries, and other fragmenting landscape features such as plantations, and a road on the genetic structure and gene flow of non-volant small mammals. In addition to these landscape features, life history and sex- specific reproductive strategies are important in shaping population structure (Chambers and Grant 2010). The spatio-temporal variability in resources, competition, and inbreeding avoidance are crucial drivers of dispersal (e.g. Greenwood 1980; Smale et al. 1997; Lawson Handley and Perrin 2007) and sex-specific differences in the benefits of philopatry often lead to a sex bias in dispersal (e.g. Greenwood 1980, Lawson Handley and Perrin 2007; Galliard et al. 2012; Mabry et al. 2013). Due to differences in the dispersal and reproductive behaviour of males and females, genetic structure may differ between sexes. We therefore also investigated the genetic variation in males and females separately to infer the dispersal regime for the study species. The relatively rapid mutation rate of mitochondrial (mtDNA) compared to nuclear DNA, and the absence of recombination events, which result in an effectively clonal inheritance, make mtDNA an informative and well established marker for phylogeography studies (Freeland 2008). Due to the availability of universal primers, which allow data to be obtained from a suite of different species without any a priori knowledge, cytochrome b (cyt b) provided a well-suited marker for this study. To uncover the relative importance of landscape features in shaping population structure in species with different ecological requirements, mitochondrial cyt b sequences in a variety of small mammals were examined to investigate (I) the isolating effect of the Kinabatangan River, and to assess (II) restricted migratory capabilities caused by landscape barriers along the riverbanks of the Kinabatangan River. Furthermore, to investigate possible sex-biases in the dispersal of the study species, we inferred and compared (III) potential signals of dispersal by males and females.

2.2 Material and Methods Study area and sampling Small mammals were sampled along the Kinabatangan River in eastern Sabah (05°20’ – 05°45’ N, 117°40’ – 118°30’ E), Malaysian Borneo. Running over a length of 560 km, the Kinabatangan is the longest river in Sabah, with diverse habitats, consisting of mangroves, swamp forest, riverine forest, seasonally flooded forest, riparian and mixed dipterocarp forest (Azmi 1998). Large areas of forest have been cleared and converted to agricultural landscapes (i.e. oil palm plantations) along the Kinabatangan, and the remaining forest is isolated in patches. Today, 27,000 hectares of disturbed forest are gazetted as the Lower Kinabatangan Wildlife Sanctuary (LKWS) (Ancrenaz et al. 2004; 22 Goossens et al. 2005; Fig. 2.1). Animals were sampled in 19 trapping sites (18 in forest, one in plantation (site SP in Fig. 2.1) from both sides of the river (north: 11, south: 8; Fig. 2.1). Sampling sites in close spatial proximity (≤ 7 km distance) and without separation by a geographic feature were considered as one site (differing in coloration in Fig. 2.1). Besides the Kinabatangan River, large tributaries (perennial, no overlapping of opposing branches), the main road (paved, two-lane, without game fences along its course through study area, in parts bordered by human settlements) and oil palm plantations were considered as geographic features potentially constraining landscape connectivity (Fig. 2.1). Based on this classification scheme, five sites (NA – ND, distances 4 – 23 km) were defined for the northern riverside and six (SE – SI, distances 3 – 31 km) for the south (Fig. 2.1). On the northern riverside the road acted as a potential barrier between site NA and site NB. Site NB and site ND were connected by a forest corridor varying in width and containing two sites (site NC a: width < 300 m, site NCb: width > 300 ≤ 800 m). Such a corridor was absent south of the river, where larger tributaries were also more numerous (i.e. between site SE and site SP, site SF and site SG, and between site SG and site SH, Fig. 2.1). Oil palm plantations are located between site SE and site SF, site SF and site SG as well as between site SH and site SI, and the road dissects site SE and site SP.

Figure 2.1 Map of the LKWS, locations of sampling sites (circles), and sites grouped by the fragment of origin (colouration)

23 During the dry seasons (June – November) of 2011 – 2013, animals were trapped in systematic capture-release sessions using a standardised trapping scheme. In each sampling site, traps were placed at 15 m intervals along transects of 135 m length. Per site 2 – 4 transects were established parallel at a distance of 25 – 75 m from each other, and parallel to the river, 20 – 400 m inland. On each transect 10 locally made wire mesh life traps (31 cm (length) x 21 cm (width) x 15 cm (height) or 28 cm x 18 cm x 12 cm), baited with oil palm fruits, were installed at different heights (ground, > 0 – 2 m) and on different substrates (litter, branches, vines or logs) targeting a variety of small mammals with different microhabitat preferences. In the seasons 2012 and 2013 an additional trap was placed in each transect in ≥ 4 m height above ground to target strictly arboreal species. All traps were checked twice within 24 hours on six consecutive days. Traps were kept open overnight to capture nocturnal non-volant species. Traps were checked the following morning and immediately baited again to capture diurnal taxa. Each trapping station was labelled and the coordinates were taken with a GPS device (GPSMap60CSx, Garmin Deutschland GmbH, Gräfelfing, Germany). In the three field seasons a total trapping effort of 9024 trap nights was achieved. Trapped animals were transferred to a cloth bag, weighed, sexed, and standard morphometric measurements (Yasuma et al. 2003; Payne and Francis 2007) were taken. Photos were taken from the head, back, belly, tail and the complete animal (dorsal and lateral view) and fur, tail and foot colour were recorded qualitatively for later species characterisation. A tissue sample was taken from the ear and stored in 2 ml tubes with 1.5 ml n-Lauroyl-Sarcosine buffer (Seutin et al. 1991). For individual identification in case of a recapture, each animal was marked with a pattern cut in the fur of the animal’s back before the animal was released at its point of capture. Fieldwork was conducted under the permission of the Economic Planning Unit Malaysia (Permit No.: UPE: 40/200/19/2871) and the Sabah Biodiversity Centre. All samples were transported under the permit of CITES (Export-Permit No. JHL(PB)600-3/18/1/1Jld.10/(103), Certificate No. 0602 and Export-Permit No. JHL(PB)600-3/18/1/1Jld.10/(494), Certificate No. 0689 and 0690; Import- Certificate No. E-05027/12 and E-05957/13) and the Sabah Biodiversity Centre (Export-Licence No.: JKM/MBS.1000-2/3(38)).

DNA amplification and sequence analysis DNA was extracted from a subset of tissues (n = 511) using a HotSHOT extraction protocol (Truettet al. 2000). Subsequent amplification of the mitochondrial cytochrome b (cyt b) gene was carried out using the universal primers L14727 5’ TGAYATGAAAAAYCATCGTTG ‘3 (Pääbo et al. 1988) or L14841 5’ AAAAAGCTTCCATCCAACATCTCAGCATGATGAAA ‘3 (Kocher et al. 1989) and the primer MVZ16 5’ AAATAGGAARTATCAYTCTGGTTTRAT ‘3 (Smith and Patton 1993). The cyt b gene in tree shrews (i.e. Tupaia longipes and T. Tupaia gracilis) was amplified using an optimised primer (L14841tupaia 5’

24 CCAGCYCCATCAAAYATYTCMTCATGATGAAAC ’3), which was designed based on previously generated cyt b sequences of Tupaia tana and Tupaia minor. The amplicons generated by the primer pairs L14727 – MVZ16 and L14841/L14841tupaia – MVZ16 had a length of 875 bp and 755 bp, respectively. DNA amplification was performed in a 15 µl reaction of 5 µl Multiplex PCR Master Mix (Qiagen), 0.1 µl Q-solution (Qiagen), 0.6 pmol of each primer, and 0.5 – 1 µl template DNA. Reactions were amplified with an initial denaturation at 95°C for 15 minutes, followed by 35 cycles of 45 seconds at 95°C, 1:30 minutes at 47°C and 1 minute at 72°C, and a final extension at 72°C for 10 minutes. PCR products were sent to Eurofins (Ebersberg, Germany) for sequencing. All samples were sequenced in forward direction. Subsequent sequence alignment and editing was performed in Sequencher version 4.9 (GeneCodes) and BioEdit version 7.2.5 (Hall 1999).

Taxonomic classification Unambiguous species assignment of captured animals on the basis of morphometric data alone was not always possible. Therefore, phylogenetic analysis based on cyt b and 16S rRNA sequencing was conducted whenever necessary for taxonomic classification. Details on the methodological approach for the taxonomic clarification are in Appendix 2.6.1.

All cyt b sequences generated for taxonomic clarification were used in later phylogeographic analyses (n = 385). In addition, from tree shrews, squirrels and gymnures at least 20 samples (10 from each riverside, whenever possible) were sequenced at the cyt b locus (n = 126) for subsequent phylogeographic analyses.

Analyses of genetic diversity and population structure Mitochondrial haplotypes were identified and intraspecific haplotype (h) and nucleotide (π) diversity were estimated for each species using DnaSP version 5.10.1 (Librado and Rozas 2009). A Mann- Whitney U Test was conducted in Statistica version 6.1 (StatSoft Inc. 2004) to assess differences in haplotype and nucleotide diversity between riversides, whenever possible (α = 0.05). Intraspecific haplotype sharing and genetic distances between species at sites on either side of the river were visualised using a minimum-spanning network computed in Arlequin version 3.5 (Excoffier and

Lischer 2010). Based on AMOVA and φST-statistics, implemented in Arlequin v. 3.5 (1,000 permutations, α = 0.05), genetic variation between the two riversides was estimated for all species sampled in large numbers on both sides of the river (n ≥ 10 per riverbank). In order to identify effects of the river, sites where assigned to their respective riverside for AMOVA (three hierarchical levels: riverside, site, individual) analysis. For the assessment of potential barriers within each riverside, an

25 AMOVA (site, individual) was conducted for the northern and southern riverside separately, and significant pairwise φST-signals between all sites with n ≥ 3 individuals (Table 2.1) were explored.

Because of its potential transitional position between site NC a and site ND, site NCb was not included in these analyses.

Sex-biased dispersal We investigated sex as a determinant of gene flow in all species for which males and females could be sampled in sufficient numbers (at least eight individuals for each sex). Since mitochondrial markers are uniparentally inherited via the mother, maternally related males and females share the same haplotype. However, immigrants may carry a different, less frequent haplotype into a population. Consequently, haplotype diversity should be higher in the dispersing sex and unique haplotypes within sites can be assumed to be a signal of (recent) immigration. Furthermore, genetic differentiation between sites should be lower in the more dispersed sex. In addition, haplotype diversity was compared between males and females in three ways: first, haplotype diversity was determined as the overall number of haplotypes in males and females. Second, the number of sex-specific haplotypes was compared between males and females. Third, the overall number of unique haplotypes (= only present in one animal per site) was determined for males and females separately. Only sites in which both sexes were sampled were included in this last analytical step. Genetic differentiation between sites was calculated for each sex separately (forall sites with n ≥ 3) using pairwise φST-statistics, and a Mann-Whitney U Test (Statistica version 6.1, StatSoft Inc. 2004) to detect significant signals between sexes (α = 0.05), whenever possible.

2.3 Results Small mammal species richness along the Kinabatangan and their genetic diversity A total of 1,185 non-volant small mammals were sampled in the three field seasons between 2011 and 2013. Based on phenotypic characteristics and/or phylogenetic analysis (Appendix 2.6.1), animals could be taxonomically assigned to 19 different taxa of rodents (murids and squirrels), tree shrews, and gymnures. While rodents showed high species diversity with nine different murid and four squirrel species, only one gymnure was trapped. Tree shrews were represented by five different taxa (Table 2.1). From the 14 species trapped on both sides of the river (Table 2.1) a 589 – 791 bp fragment of the cyt b gene was amplified and sequenced. Within this DNA fragment no insertions or deletions could be detected. The number of variable sites (s) varied from one in Rattus exulans to 48 in Maxomys whiteheadi. The highest number of haplotypes could be identified in M. whiteheadi (34), and only 26 two haplotypes were identified in R. exulans, Rattus rattus, and Rattus tanezumi (Appendix 2.6.2). Overall haplotype diversity ranged from h = 0.167 (R. rattus) to h = 0.967 (Callosciurus notatus), while nucleotide diversity varied from π = 0.0005 (R. exulans) to π = 0.0178 (T. tana). R. exulans, R. rattus, and R. tanezumi showed low haplotype and nucleotide diversity but most other species had high haplotype diversity (Appendix 2.6.3). Comparing populations from the northern and the southern riverbank, significant differences in haplotype and nucleotide diversity could not be detected in any of the analysed species (Appendix 2.6.3).

Table 2.1 Analysed cyt b sequences. Given is the number of sequences per site, the total number of analysed sequences for each riverside and the overall number of analysed sequences for each species, respectively

North South

Species NA NB NCa NCb ND Total SE SP SF SG SH SI Total All Chiropodomys sp. - - - - 1 1 ------1 Maxomys surifer ------2 14 - 8 24 24 Maxomys whiteheadi 10 14 12 21 32 89 19 - 21 41 5 7 93 182 Niviventer cremoriventer 2 6 12 1 3 24 5 - - 3 - 3 11 35 Rattus exulans - 1 2 - - 3 2 1 1 1 - - 5 8 Rattus rattus - 2 2 1 - 5 3 2 1 1 - - 7 12 Rattus tanezumi 1 3 1 - 2 7 2 10 - - 1 - 13 20 Rattus sp. - 5 - 1 13 19 3 - - 2 23 1 29 48 Sundamys muelleri 6 3 9 4 7 29 10 - 1 4 8 3 26 55 Callosciurus notatus 2 3 2 1 3 11 5 - 2 2 2 2 13 24 Callosciurus prevostii - 3 - - 1 4 1 - 1 - - - 2 6 Sundasciurus hippurus - - 1 - 4 5 ------5 Sundasciurus lowii 3 - - - 7 10 2 - 2 4 - 2 10 20 Ptilocercus lowii ------1 1 - - 2 2 Tupaia gracilis - - - - 4 4 3 - - - - - 3 7 Tupaia longipes 4 3 - 1 3 11 3 - 2 5 - 2 12 23 Tupaia minor - - 3 1 - 4 ------4 Tupaia tana 3 2 1 1 3 10 3 - 2 3 1 1 10 20 Echinosorex gymnura - 4 - - 4 8 2 - - 3 - 2 7 15

Genetic differentiation across the Kinabatangan River Interspecific differences in geographic haplotype distribution and the isolating effect of the river were examined by comparing the species-specific minimum-spanning networks (Fig. 2.2 and Appendix 2.6.4). The haplotype network patterns can be broadly grouped into three categories. Sundasciurus lowii, Tupaia tana, and Tupaia gracilis formed distinct haplotype clusters on either side of the river. (Network type 1, Fig. 2.2a). No haplotype sharing was observed among groups between the two riversides in these three species. The haplotype networks of Tupaia longipes, Callosciurus notatus, and Callosciurus prevostii showed no such grouping of haplotypes into distinct clusters. However, haplotype sharing between different sides of the river was never observed in C. prevostii and occurred only once in C. notatus and T. longipes (Network type 2, Fig. 2.2b). The third type (Fig. 2.2c and Appendix 2.6.4) of network was seen in the murid species M. whiteheadi, Niviventer

27 cremoriventer, Sundamys muelleri, Rattus sp., and in Echinosorex gymnura. They showed no distinct haplotype clustering whether north or south of the river with frequent haplotype sharing between sites from both riversides. The rat species R. exulans, R. rattus, and R. tanezumi most likely also belong to this third type, but only two haplotypes could be identified in each species, one of which was shared between the two riversides, while the other was not (Fig. 2.2c). Nine species were sampled in sufficient numbers (at least 10 individuals per riverside) to be included in the analysis of molecular variance between riversides (Table 2.2). Pairwise genetic differences between riversides were high and significant only in the squirrel S. lowii and the tree shrew T. tana (both Network type 1) and the highest percentages of molecular variance were also explained by the riversides in these two species (Table 2.2). In addition, in the whole dataset, the percentage of genetic variance partitioned within local groups was low and significant in these two species, allowing us to infer generally low gene-flow between sites for them. In all other species pairwise φ ST- statistics and AMOVA analysis revealed no clear partitioning effect of the Kinabatangan River on molecular variance (Table 2.2).

Figure 2.2 Three broad categories of minimum-spanning network: Network type 1 (a), Network type 2 (b), Network type 3 (c). In each network, each haplotype is represented as a circle. The diameter of the circle corresponds to haplotype frequency, with smallest circles representing singletons. Mutational steps between haplotypes are indicated as vertical bars.

28 Genetic differentiation along the Kinabatangan River Haplotypes were distributed widely along the river in E. gymnura and in most murid species except in Maxomys surifer, where the distribution of haplotypes was typically restricted to single locations, and only one haplotype was shared between sites on the southern riverside (Appendix 2.6.4). Similarly, no haplotype sharing between sites was found in S. lowii, C. prevostii, and T. gracilis (Fig. 2.2). Analysing the two riversides separately, significant genetic variation among sites within riversides was found in six species, but only in S. lowii on both riversides (Table 2.2). Based on significant pairwise genetic differences between sites, we could infer the impactof geographical features, other than the Kinabatangan River itself, on gene flow for six members of the resident small mammal community that could be sampled in large enough numbers. Significant differences could be detected on the southern riverside between site SF and site SG, site SE and site SF, and between site SH and site SI. Interestingly, these sites are separated by an oil palm plantation (Table 2.3). No significant differences were found between sites separated by a large tributary (site SG and site SH). Likewise, between site NA and site NB, which are separated by a road, no significant genetic differentiation could be found. Furthermore, between sites without any putative geographic barrier no significant genetic differentiation existed, even if spatially distant from each other (e.g. site NB and site ND) (Table 2.3).

Table 2.2 Pairwise differences (φST) between riversides and intraspecific percentage of genetic variance between riversides, within riversides, and within sites are given for the whole population, and for both riversides, respectively. Species are listed in order of their associated haplotype network category (Network type 1 – 3) as described in Fig. 2.2

Percentage of variance All North South Pairwise Between

φST sites, (between Between within Within Between Within Between Within Species riversides) riversides riversides sites sites sites sites sites Network Type 1 Sundasciurus lowii 0.84396** 78.62+ 15.75** 5.63** 75.23** 24.77 71.13** 28.87 Tupaia tana 0.82422** 80.85** 7.82* 11.33** 46.75* 53.25 33.23 66.77 Network Type 2 Callosciurus notatus 0.06036+ -2.09 34.81** 67.27** 43.84* 56.16 23.97+ 76.03 Tupaia longipes 0.10149+ 5.40 9.01 85.59+ 4.91 95.09 16.63 83.37 Network Type 3 Maxomys whiteheadi 0.01050+ -0.59 3.82** 96.77** 1.28 98.72 5.55** 94.45 Niviventer cremoriventer -0.03064 -15.34 38.73** 76.61** 27.76** 72.24 53.57+ 46.43 Rattus tanezumi 0.28795+ 1.46 65.22+ 33.32 34.62 65.38 0.00 0.00 Rattus sp. 0.05740+ -3.04 11.36 91.68 -14.55 114.55 25.37+ 74.63 Sundamys muelleri 0.01426 -3.01 11.76+ 91.25+ -3.18 103.18 22.09* 77.91

+ ≤ 0.1, * ≤ 0.05, ** ≤ 0.01

29 Table 2.3 Pairwise genetic differences (φST) between sites. Separating geographic features (road, plantation, tributary) and euclidean distances between sites are provided for all investigated pairings.

Road Large Tributary + Tributary Small Large distance Road Tributary + Plantation + Plantation distance distance + Plantation

NA – NB SG – SH SF – SG SE – SF NB – NCa NB – ND SH – SI Species (9.71 km) (3.11 km) (5.74 km) (5.92 km) (4.26 km) (15.51 km) (15.89 km) Maxomys whiteheadi 0.03602 -0.03250 0.06687** 0.09985** 0.06790 0.04376 -0.08650 Niviventer cremoriventer - - - - 0.11749 0.22439 - Rattus sp. ------0.14553 - Sundamys muelleri -0.24784 0.28973 - - -0.20732 0.06808 0.32948* Callosciurus notatus ------0.22449 - Tupaia longipes 0.01887 - - - - 0.26316 - * ≤ 0.05, ** ≤ 0.01

Signals of male- vs. female-biased dispersal Based on haplotype analyses a signal of sex-biased dispersal could be detected in two species (Appendix 2.6.5). In M. surifer and M. whiteheadi the overall number of haplotypes, the number of sex-specific haplotypes, and the number of unique haplotypes at a given site (“haplotype singletons”) was higher in males than in females (Appendix 2.6.5), indicating a male-biased dispersal in these species. In contrast, haplotypes were rather evenly distributed between sexes in N. cremoriventer, Rattus sp., S. muelleri, C. notatus, S. lowii, T. longipes, and T. tana (Appendix 2.6.5). Differences between the sexes can be most likely explained by the different sample sizes for both sexes in these species. However, a sex-bias in dispersal could not be verified in any of the analysed species based on pairwise φST-statistics (Appendix 2.6.5).

2.4 Discussion Small mammal diversity along the Kinabatangan River and their genetic diversity The composition of Bornean small mammal communities has been described at several locations throughout Borneo (Nor 2001; Wells et al. 2007, 2014; Nakagawa et al. 2006; Bernard et al. 2009; Charles and Ang 2010) and often comprises between 12 (Bernard et al. 2009; Charles and Ang 2010) and 22 (Nakagawa et al. 2006) non-volant small mammals of the orders Rodentia (murids, squirrels and porcupines), Scandentia (tree shrews), and Eulipotyphla (gymnures and shrews). To our knowledge, the present study is the first verifying the presence of 19 different non-volant small mammal species in habitats along the Kinabatangan River. The composition of the small mammal communities along the Kinabatangan River is similar to that of other locations studied in Borneo, and

30 indicates a rather high α-diversity with 19 different species, which includes both species native and potentially invasive to Borneo. Some species were only identified using molecular techniques which confirm the presence of cryptic species including R. tanezumi, in addition to R. rattus. Despite the high levels of habitat fragmentation along the Kinabatangan, mtDNA diversity was relatively high in most species, except in the Rattus species R. exulans, R. rattus, and R. tanezumi (only two haplotypes) (Pagès et al. 2013; Thomson et al 2014). High haplotype diversity coupled with low nucleotide diversity can indicate a population bottleneck followed by rapid growth and parallel accumulation of mutations (Grant and Bowen 1998). Such a scenario was previously invoked for orangutans along the LKWS (Jalil et al. 2008). For the Rattus species R. exulans, R. rattus and R. tanezumi, however, low haplotype and low nucleotide diversity may rather indicate a recent founder event, and the presence of only two haplotypes, one of which being high in frequency, may even imply a very limited number of founders. The Pacific rat (R. exulans) as well as black rats (Oceanic form: R. rattus, Asian form: R. tanezumi) are known to be strongly associated with human settlements (Roberts 1991; Payne and Francis 2007; Aplin et al. 2011; Banks and Hughes 2012), and could very likely have invaded the landscape around LKWS more recently as a consequence of the forest loss starting in the 1950 and human habitation (Latip et al. 2013). The conversion of forest along the Lower Kinabatangan to agriculture has provided suitable habitats for these species and their presence in the plantation sampling site supports their adaptation to anthropogenically altered environments.

Connectivity of Bornean small mammals across the Kinabatangan River Haplotype networks and genetic variance analyses indicate contrasting levels of genetic isolation between riversides within species sampled on both sides of the Kinabatangan River. Clear genetic separation between riversides, indicating a strong influence of the Kinabatangan River as a barrier for dispersal, was found in S. lowii, T. tana, and T. gracilis, and was supported for those species trapped on one riverside only such as M. surifer, S. hippurus, and T. minor. Moderate genetic isolation was found between riversides in C. notatus, C. prevostii, and T. longipes, indicating a moderate barrier function of the Kinabatangan River. No genetic separation between riversides indicates an unimpeded dispersal across the Kinabatangan, and this was found in the remaining murid rodents and in the moonrat E. gymnura. Consequently, the Kinabatangan River represents an important landscape feature for shaping the genetic structure and constraining dispersal for some species, notably in squirrels and tree shrews. Sporadic reports of squirrels crossing water bodies in North America and Asia (e.g. Applegate and McCord 1974; Pauli 2005; Brunke pers. obs.) show that certain squirrel species are able to swim. However, the presence of only one haplotype (out of 24 samples) shared between riversides in C.

31 notatus, implies that this species crosses the river only rarely. The same may also be true for the tree shrew T. longipes, where only one haplotype (among 23 samples) was shared between riversides. Although no observations exist on the swimming capabilities of tree shrews, they generally prefer dense understorey and avoid open areas without cover that may expose them to predators (Emmons 2000). In contrast, most murids are known to be good swimmers (e.g. Evans et al. 1978; Cook et al. 2001; Nicolas and Colyn 2006; Santori et al. 2008; Brunke pers. obs.), and migration across rivers have been confirmed in murid rodents elsewhere (e.g. Patton et al. 1994; Colombi et al. 2010; Roratto et al. 2014; Russo et al. 2016). Therefore, migration across the river in murids is not surprising. Interestingly, this study allowed us to infer swimming and river crossing in wild Bornean moonrats for the first time. Limited knowledge of E. gymnura ecology is mainly on captive individuals (Liat 1967; Gould 1978), but aquatic feeding activities have been observed, including swimming and diving, and morphological adaptations to this behavioural trait have been identified (Gould 1978). Additionally, stomach content analyses indicated a partially aquatic diet (fish and crabs, Liat 1967). For species with an affinity to water such as murids and the moonrat, the Kinabatangan River may even facilitate dispersal. When species with swimming capabilities (murids and moonrat) use the river, they may eventually disperse over longer distances by passive drifting downriver than if they would only move by terrestrial locomotion. Future studies are needed to evaluate this possible enhancing effect of the river on dispersal distances in more detail. The mixed effects of large rivers on dispersal were shown also in comprehensive studies on Neotropical bird communities (e.g. Capparella 1988, 1991; Burney and Brumfield 2009). Capparella (1988, 1991) suggested an important role of species-specific differences of ecological niches for migratory propensities across rivers. Here, understory species are the least likely to cross large light gaps such as rivers, compared to species inhabiting the forest canopy and open areas, which probably are more accustomed to light gaps. In our study, the river acted as a barrier mainly for the diurnal species, whereas all but one nocturnal species (all murids besides M. surifer) showed signals of dispersal across the river. It can be hypothesised that diurnal species such as squirrels and tree shrews, which move preferably in dense cover of the forest (Emmons 2000, Wells et al. 2006a), may also avoid open areas, such as larger rivers, as a predator avoidance strategy, while nocturnal rats may be safe to cross the river in the cover of the darkness.

Connectivity of Bornean small mammals along the Kinabatangan River A variety of landscape features occur along the Kinabatangan River, including large tributaries, a road and oil palm plantations, which may potentially affect dispersal and genetic connectivity for small mammals. In general, low genetic differentiation coupled with a broad geographic distributionof

32 haplotypes implies relative unrestricted migration among sites. This was typically found in the murid species. Furthermore, dispersal across the river may help to reduce the effect of terrestrial barriers and may help to explain the high genetic connectivity in murids. High genetic differentiation among sites along both riversides was only found unambiguously in S. lowii, indicating strong limitations in its terrestrial vagility, coupled with a low dispersal across the river. Both, low terrestrial and low aquatic dispersal, may lead to an especially accented genetic isolation of this species along the river. No obvious effect of large tributaries on migration could be detected in any species. As the size and/or the width of rivers are likely to determine the extent to which such a barrier is permeable for small mammals (Russo et al. 2016), it is possible that the Kinabatangan River represents the only river barrier in the area. Furthermore, water levels fluctuate throughout the year and may allow occasional crossing of tributaries by otherwise constrained species during the dry season, which may be sufficient to maintain detectable gene flow in such species. Likewise, the one major road seems not have affected inferred dispersal in any of the small mammal species, although a barrier effect of major roads has been reported for other small mammals in other regions of the world (Gerlach and Muslof 2000; Goosem 2002; Rico et al. 2007). In particular, wide roads (> 30 m) represent an almost impermeable barrier for small mammals (Gerlach and Muslof 2000; Rico et al. 2007). It is possible that the width of the road (about 7 m road surface) is not enough to have resulted in genetic isolation or that the time span since the construction of the road (ca. 20 years before this study) may not have been long enough to result in genetic differentiation. Although sample sizes were small in most species, in species with a high number of samples distinct levels of genetic differentiation existed between sites separated by plantations, while sites connected by forest corridors of the same distance showed no genetic differentiation. This indicates that landscape features (i.e. oil palm plantations) have an impact on the dispersal of non-volant small mammals along the Kinabatangan River, and emphasises the importance of forest corridors for population connectivity, which evidently increases animal movements and gene flow between patches (Bennett 1990; Haddad et al. 2003). Future studies, comprising larger sample sizes and/or employing genetic markers with higher resolving power (e.g. microsatellites or SNPs) would clearly be needed to clarify this question.

Evidence for sex-specific dispersal patterns in Bornean small mammals Physical barriers restricting migration of individuals and shaping genetic structures should have a greater impact on the dispersing sex than on the philopatric sex. However, dispersal patterns have not been studied in most Bornean small mammal species until now. Wells et al. (2008a) detected differences in the movement patterns of Bornean small mammals based on mark-recapture analysis, and suggested an influence of sex for M. surifer, M. whiteheadi, N. cremoriventer, S. muelleri, and T.

33 longipes. Additionally, Munshi-South (2008) described a female-biased dispersal in T. tana on the basis of microsatellite analysis. Although male-biased dispersal is the dominant dispersal pattern in most mammals due to higher benefits of philopatry for females (Greenwood 1980), we found signals of male-biased dispersal only in the murids M. surifer and M. whiteheadi, but this was solely based on higher haplotype diversity in males than in females. These findings do not exclude the possibility that females might occasionally disperse too, but possibly less frequently and over shorter distances than males. No clear signal of sex-biased dispersal could be detected in N. cremoriventer, Rattus sp., S. muelleri, C. notatus, S. lowii, T. longipes, and T. tana, because haplotypes were rather evenly distributed between sexes in these species. Nevertheless, findings for the squirrel species are congruent with other studies (Waser and Jones 1983; Wauters and Dhondt 1992; Smale et al. 1997), which suggested dispersal by both sexes, at least for C. notatus and S. lowii. It needs to be acknowledged that the rather small sample sizes have limited the scope of the haplotype and φST-analysis, and that larger sample sizes are needed to detect signals of sex-biases in dispersal, unambiguously. Furthermore, because of its uniparental inheritance through the maternal line, mtDNA has some limitations, and differences in dispersal rates among sexes may manifest differently in the genetic structure of males and females. Therefore, further studies encompassing a range of paternally and bi-parentally inherited markers, in addition to maternally inherited marker, are clearly necessary to complete the understanding of possible proximate and ultimate factors that determine and influence the dispersal of males and females in these rather cryptic animals. Especially, because large gaps still exist in the knowledge of the socioecology and behaviour of most Bornean small mammal species.

Conservation implications A diverse small mammal community was confirmed in the present study. However, extensive land conversion to oil palm plantations, habitat fragmentation and the large Kinabatangan River create physical barriers which may constrain the connectivity of small mammal populations by restricting dispersal. The genetic analyses carried out here found some support for this effect in some but not all studied species. The Kinabatangan River acts as an important geographic feature shaping the genetic structure of Bornean small mammals in very different ways by either restricting (in squirrels and tree shrews) or potentially even promoting dispersal (in murids and gymnures). Species able to cross the river (i.e. murids) may genetically even benefit from its presence, since the use of the river for migration may mitigate the effects of other barriers along the river. On the other hand, species which seem not to cross the river (S. lowii, T. tana, M. surifer) are expected to suffer much more from forest fragmentation along both riversides. In particular, S. lowii showed very limited gene flow along the

34 river, but the scarcity of information about its ecological requirements limits speculations about the restricting habitat factors. Although findings were limited to a small number of species, this study revealed that plantations affected dispersal and gene flow in small mammals more than any other studied barrier beside the Kinabatangan River, such as its tributaries or a paved road. The higher genetic connectivity along the northern compared to the southern riverside in most of the small mammals is most likely due to the presence of forest corridors along the northern border of the river. Given that not all small mammal species can cross the river, such corridors should also be established on the highly fragmented southern riverside to increase population connectivity. This would constitute an important conservation measure especially for those species the Kinabatangan represents a major barrier for dispersal (such as S. lowii, T. tana, and M. surifer). Finally, many studies have shown that invasive species can have an impact on the persistence of the native fauna, and are even capable of replacing it (Banks and Hughes 2012; Wells et al. 2014). Moreover, invasive species have been identified as the second most important cause of biodiversity loss after habitat destruction (Abdelkrim et al. 2005), and as a major vector of diseases for humans and wildlife alike (Meerburg et al. 2009; Aplin et al. 2011). This study has confirmed the presence of three invasive rat species (R. exulans, R. rattus, and R. tanezumi) in both anthropogenic modified and forest habitats of the LKWS, and other studies confirmed these rat species in forest habitats elsewhere on Borneo (e.g. Nakagawa et al. 2006; Wells et al. 2006b, 2007, 2014), which emphasises their potential to invade into forest environments. It is difficult to predict which of the native small mammal species will persist or suffer from competition with these invasive species. However, previous studies demonstrated that larger species are often dominant over smaller ones (Banks and Hughes 2012) and this pattern of competitive dominance has been confirmed for black rats in other regions (e.g. Harris and Macdonald 2007; Harper and Cabrera 2010; Guo et al. 2017). Potential implications of these invasive rats for the native fauna need to be urgently assessed by establishing monitoring programmes across different sites and both riversides. Eventually, effective conservation measures need to be developed that include management decisions on the invasive species but also on those that may be negatively affected by them to ensure the long-term survival of the native species and to preserve the biodiversity within the forest along the Kinabatangan River.

35 Acknowledgements We thank the Malaysian Economic Planning Unit and the Sabah Biodiversity Centre for permission to carry out research in Sabah. We are grateful to Audrey Adella Umbol for her help with local authorities and Jumrafiah Abd. Shukor for accepting the role of the Malaysian counterpart. We would like to thank the research assistants at the Danau Girang Field Centre for their support in the field. Among them we specially thank Samsir bin Laimun, Petrieadi Ambo Tola, and Saroto bin Payar for their endless help. We are also grateful to Baharudin bin Resake and Simon Amos for sharing their climbing skills. We are furthermore grateful to Josephine D’Urban-Jackson, Rebecca Lawrence and all the students offering their help during field and lab work. For financial support of JB we thank the Calenberg-Grubenhagensche Landschaft, Danau Girang Field Centre, and the University of Veterinary Medicine Hannover.

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39 2.6 Appendix

Appendix 2.6.1 Taxonomic clarification Materials and Methods Phylogenetic analyses were conducted whenever necessary for taxonomic clarification basedon mitochondrial cytochrome b (cyt b) (Murids: n = 385, Scandentia: n = 56) and 16S ribosomal RNA (16S rRNA) sequences (Scandentia: n = 69). Due to the scarcity of available reference sequences for the cyt b region in tree shrews, the 16S rRNA gene was sequenced additionally. For partial amplification of the 16S rRNA gene the universal vertebrate primer 16S ar-L (5’ CGCCTGTTTATCAAAAACAT ‘3) and 16S br-H (5’ CCGGTCTGAACTCAGAT CACGT ‘3) (Hedges 1994, Smith & Wheeler 2004) were used, resulting in amplicons of 595 bp in length. The same reaction as described for cyt b were used for DNA amplification of the 16S rRNA, but thermal cycling was set for 15 minutes at 95°C for initial denaturation, followed by 35 cycles of 45 seconds at 95°C, 1:30 minutes at 55°C and 1 minute at 72°C, and a final extension of 10 minutes at 72°C. All newly generated (haplotype) sequences were deposited in GenBank under the accession no. MK111862 – MK112036 (Appendix 2.6.1.1c and Appendix 2.6.1.1d).

All sequences were aligned with reference sequences from GenBank (Appendix 2.6.1.1a and Appendix 2.6.1.1b) in Sequencher version 4.9 (GeneCodes). In order to test for the best-fit nucleotide substitution model, a likelihood ratio test was performed in jModelTest version 2.1.10 (Guindon and Gascuel 2003; Darriba et al. 2012) under the Akaike Information Criterion for murids, and tree shrews, respectively. Models were subsequently applied in phylogenetic analyses using MEGA version 6 (Tamura et al. 2013). Neighbour-joining trees were constructed separately for murids, and tree shrews. Sampled individuals were assigned to their respective species according to their position in the phylogenetic gene tree and the underlying bootstrap values (based on 1,000 bootstrap replications).

Phylogenetic analyses of squirrels and gymnures were not necessary, as an unambiguous taxonomic clarification could be reached based on phenotypic characteristics.

Results Due to the large number of murid samples, three different phylogenetic trees were constructed (NJ Muridae): NJ Maxomys cyt b (Maxomys spp.), NJ Niviventer-Chiropodomys cyt b (Niviventer spp. and Chiropodomys spp.), and NJ Rattus-Sundamys cyt b (Rattus spp. and Sundamys spp.; Appendix 2.6.1.2). Unambiguous classification to a species was not possible for samples of two different murid

40 taxa, neither using phenotypic nor genetic analyses. Nevertheless, phenotypic characteristics and their monophyletic position in the phylogenetic tree allowed the categorisation of these samples as Rattus sp. and Chiropodomys sp., respectively (see tree NJ Rattus-Sundamys cyt b and NJ Niviviventer-Chiropodomys cyt b; Appendix 2.6.1.2).

In order to strengthen the rather weak support of monophyletic groups in the phylogenetic 16S rRNA tree (NJ Tupaia 16S rRNA), we conducted an additional phylogenetic tree based on cyt b sequences for tree shrews (NJ Tupaia cyt b; Appendix 2.6.1.3).

References Darriba D, Taboada GL, Doallo R, Posada D (2012) jModelTest 2: more models, new heuristics and parallel computing. Nature Methods 9: 772 GeneCodes Sequencher 4.9 Sequence analysis software ed: Gene Codes Corporation, Ann Arbour, MI, USA Available: http://www.genecodes.com. [Accessed 2013 Dec 12] Guindon S, Gascuel O (2003) A simple, fast and accurate method to estimate large phylogenies by maximum-likelihood. Syst Biol 52: 696-704 Hedges SB (1994) Molecular evidence for the origin of birds. Proc Natl Acad Sci 91: 2621-2624 Smith WL, Wheeler WC (2004) Polyphyly of the mail-cheeked fishes (Teleostei: Scorpaeniformes): evidence from mitochondrial and nuclear sequence data. Mol Phylogenet Evol 32: 627-646 Tamura K, Stecher G, Peterson D, Filipski A, Kumar S (2013) MEGA6: Molecular Evolutionary Genetics Analysis version 6.0. Mol Biol Evol 30: 2725-2729

41 Appendix 2.6.1.1a Cyt b reference sequences used for taxonomic analyses of the murids

Species Locus Accession No. Muridae Chiropodomys gliroides cyt b EU349740 cyt b KJ607274 Maxomys ochraceiventer cyt b JF436983 cyt b JF436984 Maxomys rajah cyt b KC878135 cyt b JF436997 Maxomys surifer cyt b HQ877134 cyt b JF437004 Maxomys tajuddinii cyt b KC878151 Maxomys whiteheadi cyt b HQ877147 cyt b HQ877150 cyt b HQ877151 cyt b KC878153 Niviventer cremoriventer cyt b JF437007 cyt b JF436994 Niviventer rapit cyt b DQ191483 cyt b HQ877098 Rattus argentiventer cyt b HM217362 cyt b JN675488 Rattus exulans cyt b DQ191486 cyt b HM217424 Rattus norvegicus cyt b DQ439840 cyt b HM217481 Rattus rattus cyt b JN675612 cyt b JQ823506 Rattus tanezumi cyt b EF186503 cyt b JQ814266 Rattus tiomanicus cyt b JF436975 cyt b JF436986 Sundamys muelleri cyt b AM408340 cyt b DQ191490 cyt b JF437012 Callosciurus notatus (outgr.) cyt b AB499913 Hylopetes spadiceus (outgr.) cyt b DQ093189

Appendix 2.6.1.1b 16S rRNA reference sequences used for taxonimic analyses of tree shrews

Species Locus Accession No. Ptilocercidae Ptilocercus lowii 16S rRNA JF795291 Tupaiidae Dendrogale melanura 16S rRNA JF795293 Tupaia dorsalis 16S rRNA JF795305 Tupaia gracilis 16S rRNA JF795309 Tupaia longipes 16S rRNA JF795311 Tupaia minor 16S rRNA JF795313 Tupaia montana 16S rRNA JF795315 Tupaia picta 16S rRNA JF795318 Tupaia splendidula 16S rRNA JF795319 Tupaia tana 16S rRNA JF795322 Nycticebus coucang (outgr.) 16S rRNA JF795323

42 Appendix 2.6.1.1c Cyt b sequences used for taxonomic and phylogeographic analyses. For each haplotype the Sample ID, Haplotype ID, Haplotype frequencies, and GenBank-Accession No. are given

Haplotype Accession Haplotype Accession Sample ID Haplotype freq. No. Sample ID Haplotype freq. No. MURIDAE (cyt b) Niviventer cremoriventer Chiropodomys sp. KINA106 Nc 7 1 MK111909 KINA618 - 1 MK111862 KINA525 Nc 8 2 MK111910 KINA19 Nc 9 2 MK111911 Maxomys surifer KINA881 Nc 10 1 MK111912 KINA92 Ms 1 13 MK111863 KINA129 Nc 11 1 MK111913 KINA1146 Ms 2 4 MK111864 KINA130 Nc 12 4 MK111914 KINA1150 Ms 3 1 MK111865 KINA219 Nc 13 1 MK111915 KINA430 Ms 4 3 MK111866 KINA19 Nc 9 2 MK111911 KINA1168 Ms 5 2 MK111867 KINA881 Nc 10 1 MK111912 KINA70 Ms 6 1 MK111868 KINA129 Nc 11 1 MK111913 KINA130 Nc 12 4 MK111914 Maxomys whiteheadi KINA219 Nc 13 1 MK111915 KINA759 Mw 1 9 MK111869 KINA717 Mw 2 7 MK111870 Rattus exulans KINA639 Mw 3 2 MK111871 KINA361 Re 1 2 MK111916 KINA427 Mw 4 1 MK111872 KINA60 Re 2 6 MK111917 KINA398 Mw 5 4 MK111873 KINA1036 Mw 6 1 MK111874 Rattus rattus KINA797 Mw 7 1 MK111875 KINA151 Rr 1 1 MK111918 KINA1074 Mw 8 3 MK111876 KINA62 Rr 2 11 MK111919 KINA703 Mw 9 1 MK111877 KINA1153 Mw 10 3 MK111878 Rattus tanezumi KINA685 Mw 11 1 MK111879 KINA400 Rt 1 2 MK111920 KINA901 Mw 12 5 MK111880 KINA368 Rt 2 18 MK111921 KINA883 Mw 13 22 MK111881 KINA10 Mw 14 40 MK111882 Rattus sp. KINA736 Mw 15 1 MK111883 KINA360 Rattus 1 26 MK111922 KINA1058 Mw 16 2 MK111884 KINA1120 Rattus 2 1 MK111923 KINA640 Mw 17 2 MK111885 KINA620 Rattus 3 12 MK111924 KINA170 Mw 18 4 MK111886 KINA776 Rattus 4 8 MK111925 KINA871 Mw 19 4 MK111887 KINA93 Rattus 5 1 MK111926 KINA720 Mw 20 1 MK111888 KINA424 Mw 21 1 MK111889 Sundamys muelleri KINA422 Mw 22 2 MK111890 KINA186 Sm 1 5 MK111927 KINA393 Mw 23 5 MK111891 KINA1067 Sm 2 1 MK111928 KINA902 Mw 24 2 MK111892 KINA802 Sm 3 7 MK111929 KINA502 Mw 25 3 MK111893 KINA499 Sm 4 2 MK111930 KINA756 Mw 26 5 MK111894 KINA444 Sm 5 1 MK111931 KINA1004 Mw 27 2 MK111895 KINA1143 Sm 6 1 MK111932 KINA649 Mw 28 1 MK111896 KINA45 Sm 7 2 MK111933 KINA1130 Mw 29 6 MK111897 KINA831 Sm 8 1 MK111934 KINA2 Mw 30 34 MK111898 KINA216 Sm 9 12 MK111935 KINA471 Mw 31 1 MK111899 KINA993 Sm 10 1 MK111936 KINA301 Mw 32 1 MK111900 KINA890 Sm 11 2 MK111937 KINA792 Mw 33 2 MK111901 KINA378 Sm 12 1 MK111938 KINA467 Mw 34 3 MK111902 KINA733 Sm 13 3 MK111939 KINA13 Sm 14 3 MK111940 Niviventer cremoriventer KINA410 Sm 15 1 MK111941 KINA357 Nc 1 1 MK111903 KINA104 Sm 16 1 MK111942 KINA11 Nc 2 16 MK111904 KINA404 Sm 17 11 MK111943 KINA1188 Nc 3 1 MK111905 KINA242 Nc 4 2 MK111906 KINA15 Nc 5 1 MK111907 KINA1149 Nc 6 2 MK111908

43 Appendix 2.6.1.1c continued Haplotype Accession Haplotype Accession Sample ID Haplotype freq. No. Sample ID Haplotype freq. No. SCIURIDAE (cyt b) TUPAIIDAE (cyt b) Callosciurus notatus Tupaia gracilis KINA520 Cn 1 2 MK111944 KINA896 Tg 1 1 MK111984 KINA553 Cn 2 2 MK111945 KINA599 Tg 2 3 MK111985 KINA995 Cn 3 1 MK111946 KINA1029 Tg 3 3 MK111986 KINA764 Cn 4 3 MK111947 KINA1101 Cn 5 1 MK111948 Tupaia longipes KINA253 Cn 6 1 MK111949 KINA446 Tl 1 6 MK111987 KINA1066 Cn 7 2 MK111950 KINA1142 Tl 2 1 MK111988 KINA1152 Cn 8 1 MK111951 KINA235 Tl 3 1 MK111989 KINA316 Cn 9 1 MK111952 KINA214 Tl 4 1 MK111990 KINA167 Cn 10 1 MK111953 KINA370 Tl 5 1 MK111991 KINA1158 Cn 11 1 MK111954 KINA1062 Tl 6 4 MK111992 KINA836 Cn 12 2 MK111955 KINA377 Tl 7 2 MK111993 KINA415 Cn 13 2 MK111956 KINA914 Tl 8 2 MK111994 KINA237 Cn 14 1 MK111957 KINA630 Tl 9 2 MK111995 KINA897 Cn 15 1 MK111958 KINA391 Tl 10 1 MK111996 KINA466 Cn 16 2 MK111959 KINA542 Tl 11 2 MK111997

Callosciurus prevostii Tupaia minor KINA176 Cp 1 1 MK111960 KINA57 Tm 1 1 MK111998 KINA521 Cp 2 2 MK111961 KINA56 Tm 2 1 MK111999 KINA1059 Cp 3 1 MK111962 KINA760 Tm 3 1 MK111200 KINA1016 Cp 4 1 MK111963 KINA63 Tm 4 1 MK111201 KINA532 Cp 5 1 MK111964 Tupaia tana Sundasciurus hippurus KINA392 Tt 1 3 MK111202 KINA29 Sh 1 1 MK111965 KINA504 Tt 2 1 MK111203 KINA996 Sh 2 1 MK111966 KINA38 Tt 3 1 MK111204 KINA627 Sh 3 1 MK111967 KINA383 Tt 4 1 MK111205 KINA589 Sh 4 1 MK111968 KINA579 Tt 5 2 MK111206 KINA961 Sh 5 1 MK111969 KINA909 Tt 6 2 MK111207 KINA118 Tt 7 2 MK111208 Sundasciurus lowii KINA1100 Tt 8 1 MK111209 KINA238 Sl 1 3 MK111970 KINA287 Tt 9 1 MK111210 KINA986 Sl 2 2 MK111971 KINA196 Tt 10 4 MK111211 KINA583 Sl 3 4 MK111972 KINA799 Tt 11 1 MK111212 KINA913 Sl 4 1 MK111973 KINA175 Tt 12 1 MK111213 KINA278 Sl 5 1 MK111974 KINA306 Sl 6 1 MK111975 ERINACEIDAE (cyt b) KINA1173 Sl 7 2 MK111976 Echinosorex gymnura KINA676 Sl 8 1 MK111977 KINA551 Eg 1 1 MK111214 KINA112 Sl 9 1 MK111978 KINA989 Eg 2 1 MK111215 KINA177 Sl 10 2 MK111979 KINA945 Eg 3 6 MK111216 KINA1057 Sl 11 1 MK111980 KINA103 Eg 4 1 MK111217 KINA1064 Sl 12 1 MK111981 KINA1182 Eg 5 1 MK111218 KINA541 Eg 6 5 MK111219 PTILOCERCIDAE (cyt b) Ptilocercus lowii KINA670 Pl 1 1 MK111982 KINA461 Pl 2 1 MK111983

44 Appendix 2.6.1.1d Tree shrew 16S rRNA sequences used for taxonomic analysis. For each haplotype the Sample ID, Haplotype ID, Haplotype frequencies, and GenBank-Accession No. are given

Haplotype Accession Haplotype Accession Sample ID Haplotype freq. No. Sample ID Haplotype freq. No. PTILOCERCIDAE (16S rRNA) Tupaia minor Ptilocercus lowii KINA57 Tm 1 4 MK112029 KINA670 Pl 1 2 MK112020 Tupaia tana TUPAIIDAE (16S rRNA) KINA127 Tt 1 6 MK112030 Tupaia gracilis KINA799 Tt 2 1 MK112031 KINA896 Tg 1 7 MK112021 KINA128 Tt 3 2 MK112032 KINA1179 Tt 4 1 MK112033 Tupaia longipes KINA330 Tt 5 7 MK112034 KINA84 Tl 1 10 MK112022 KINA392 Tt 6 1 MK112035 KINA203 Tl 2 6 MK112023 KINA735 Tt 7 5 MK112036 KINA391 Tl 3 3 MK112024 KINA1142 Tl 4 1 MK112025 KINA108 Tl 5 8 MK112026 KINA214 Tl 6 3 MK112027 KINA81 Tl 7 2 MK112028

45 Appendix 2.6.1.2 Neighbour joining trees of murids: Maxomys spp. (a), Chiropodomys spp and Niviventer spp. (b), Rattus spp and Sundamys spp. (c). Detailed images online at https://doi.org/10.1007/s10592-019-01159-3 (Supplementary Material 1)

a c

46 Appendix 2.6.1.3 Neighbour joining trees of tree shrews (cyt b [a] and 16S rRNA [b]). Detailed images online at https://doi.org/10.1007/s10592-019-01159-3 (Supplementary Material 2)

47 Appendix 2.6.2 Length of analysed cyt b sequences, number of variable sites (s), transition (ti) and transversion (tv) ratio, the number of haplotypes found in all analysed sequences, and sequences sampled on the northern and southern riverside, respectively

No. of haplotypes Seq. length Species [bp] s ti:tv All North South Maxomys whiteheadi 589 48 43:5 34 18 24 Niviventer cremoriventer 640 26 22:4 13 10 5 Rattus exulans 791 1 1:0 2 2 1 Rattus rattus 764 6 6:0 2 1 2 Rattus tanezumi 746 5 3:2 2 2 1 Rattussp. 633 8 7:1 5 3 5 Sundamys muelleri 638 29 25:4 17 13 11 Callosciurus notatus 745 31 27:4 16 7 10 Callosciurus prevostii 751 20 19:1 5 3 2 Sundasciurus lowii 634 28 23:5 12 4 8 Tupaia gracilis 676 5 5:0 3 2 1 Tupaia longipes 675 23 20:3 11 8 4 Tupaia tana 646 32 26:6 12 6 6 Echinosorex gymnura 634 15 15:0 6 5 3

Appendix 2.6.3 Haplotype (h ± SD) and nucleotide diversity (π ± SD) of cyt b sequences, are given together with Z and p- values of the Mann-Whitney U Test for all sampled individuals and for individuals from the two riversides, respectively. For missing values (-) statistical comparison was not possible due to small sample sizes

Species h π All North South Z p All North South Z p Maxomys 0.896 0.855 0.909 0.0086 0.0085 0.0086 whiteheadi ± 0.013 ± 0.025 ± 0.015 -1.358 0.222 ± 0.0002 ± 0.0003 ± 0.0004 -0.313 0.841 Niviventer 0.782 0.783 0.782 0.0096 0.0107 0.0074 cremoriventer ± 0.069 ± 0.081 ± 0.107 1.414 0.229 ± 0.0013 ± 0.0016 ± 0.0015 0.707 0.629 Rattus 0.429 0.667 0.000 0.0005 0.0008 0.0000 exulans ± 0.169 ± 0.314 ± 0.000 - - ± 0.0002 ± 0.0004 ± 0.0000 - - Rattus 0.167 0.000 0.286 0.0013 0.0000 0.0022 rattus ± 0.134 ± 0.000 ± 0.196 - - ± 0.0011 ± 0.0000 ± 0.0015 - - Rattus 0.189 0.048 0.000 0.0013 0.0032 0.0000 tanezumi ± 0.108 ± 0.171 ± 0.000 - - ± 0.0007 ± 0.0012 ± 0.0000 - - Rattus 0.629 0.620 0.579 0.0046 0.0044 0.0045 sp. ± 0.053 ± 0.061 ± 0.089 - - ± 0.0003 ± 0.0004 ± 0.0006 - - Sundamys 0.892 0.867 0.855 0.0107 0.0093 0.0120 muelleri ± 0.023 ± 0.052 ± 0.048 0.245 0.905 ± 0.0009 ± 0.0014 ± 0.0012 -0.367 0.730 Callosciurus 0.967 0.909 0.962 0.0096 0.0110 0.0079 notatus ± 0.019 ± 0.066 ± 0.041 0.245 0.905 ± 0.0009 ± 0.0012 ± 0.0009 -0.122 0.905 Callosciurus 0.933 0.833 1.000 0.0119 0.0129 0.0107 prevostii ± 0.122 ± 0.222 ± 0.500 - - ± 0.0026 ± 0.0034 ± 0.0053 - - Sundasciurus 0.937 0.778 0.956 0.0173 0.0035 0.0055 lowii ± 0.033 ± 0.091 ± 0.059 - - ± 0.0010 ± 0.0005 ± 0.0007 - - Tupaia 0.714 0.500 0.000 0.0038 0.0007 0.0000 gracilis ± 0.127 ± 0.265 ± 0.000 - - ± 0.0007 ± 0.0004 ± 0.0000 - - Tupaia 0.901 0.945 0.682 0.0084 0.0095 0.0066 longipes ± 0.040 ± 0.054 ± 0.102 0.884 0.400 ± 0.0009 ± 0.0009 ± 0.0015 1.061 0.400 Tupaia 0.937 0.889 0.844 0.0178 0.0069 0.0034 tana ± 0.033 ± 0.075 ± 0.103 0.000 1.000 ± 0.0011 ± 0.0009 ± 0.0008 1.528 0.200 Echinosorex 0.762 0.857 0.714 0.0050 0.0063 0.0036 gymnura ± 0.081 ± 0.108 ± 0.127 - - ± 0.0015 ±0.0024 ± 0.0008 - -

48 Appendix 2.6.4 Minimum spanning network of M. whiteheadii, M. surifer, S. hippurus, and T. minor

Maxomys whiteheadi

Maxomys surifer

Tupaia minor

Sundasciurus hippurus

49 Appendix 2.6.5 Sex-specific sampling and distribution of cyt b haplotypes. Listed are the number of sampled males and females, the number of overall haplotypes for males and females, the number of haplotypes seen only in males or females, the number of single male and female individuals with unique haplotypes in sites in which both sexes were sampled, and the number of sites in which both sexes could be sampled. Furthermore, results of the pairwise φST-statistic (mean ± SD) together with the Z and p-values of the Mann-Whitney U Test is given for each analysed species

Sex-specific Haplotype

Individuals Haplotypes haplotypes singletons Sites Pairwise φST both Species females males females males females males females males sexes females males Z p Maxomys surifer 13 11 3 6 0 3 1 2 3 - - - - 0.052 0.029 Maxomys whiteheadi 88 94 22 30 4 12 16 28 10 ± 0.104 ± 0.089 1.157 0.247 Niviventer cremoriventer 18 16 8 8 5 5 8 7 7 - - - - 0.351 -0.089 Rattussp. 20 28 3 4 1 2 1 4 4 ± 0.492 ± 0.057 0.655 0.700 0.122 0.286 Sundamys muelleri 32 23 13 10 7 4 15 12 9 ± 0.305 ± 0.350 -1.003 0.315 Callosciurus notatus 13 11 10 9 7 6 9 8 9 - - - - Sundasciurus lowii 11 9 8 8 4 4 3 4 5 - - - - Tupaia longipes 11 12 7 8 3 4 4 5 7 - - - - Tupaia tana 12 8 9 6 6 3 3 3 4 - - - -

50 Chapter 3 Dispersal and genetic structure in a tropical small mammal, the Bornean tree shrew (Tupaia longipes), in a fragmented landscape along the Kinabatangan River, Sabah, Malaysia

Published 2020 in BMC Genetics Volume 21, Issue 43, https://doi.org/10.1186/s12863-020-00849-z by Jennifer Brunke1, Isa-Rita M. Russo2, Pablo Orozco-terWengel2, Elke Zimmermann1†, Michael W. Bruford2,3, Benoit Goossens2,3,4,5, Ute Radespiel1

1 University of Veterinary Medicine Hannover, Institute of Zoology, Buenteweg 17, 30559 Hannover, Germany 2 Organisms and Environment Division, Cardiff University, School of Biosciences, Sir Martin Evans Building, Museum Avenue, Cardiff CF10 3AX, United Kingdom 3 Sustainable Places Research Institute, Cardiff University, 33 Park Place, Cardiff CF10 3BA, United Kingdom 4 Sabah Wildlife Department, Wisma Muis, 88100 Kota Kinabalu, Sabah, Malaysia 5 Danau Girang Field Centre, c/o Sabah Wildlife Department, Wisma Muis, 88100 Kota Kinabalu, Sabah, Malaysia † deceased

Constraints in migratory capabilities, such as the disruption of gene flow and genetic connectivity caused by habitat fragmentation, are known to affect genetic diversity and the long-term persistence of populations. Although negative population trends due to ongoing forest loss are widespread, the consequence of habitat fragmentation on genetic diversity, gene flow and the genetic structure has rarely been investigated in Bornean small mammals. To fill this gap in knowledge, we used nuclear and mitochondrial DNA markers to assess genetic diversity, gene flow and genetic structure in the Bornean tree shrew, Tupaia longipes, that inhabits forest fragments of the Lower Kinabatangan Wildlife Sanctuary, Sabah. Furthermore, we used these markers to assess dispersal regimes in male and female T. longipes. In addition to the Kinabatangan River, a known barrier for dispersal in tree shrews, the heterogeneous landscape along the riverbanks affected the genetic structure in this species. Specifically, while in larger connected forest fragments along the northern riverbank genetic connectivity was relatively undisturbed, patterns of genetic differentiation and the distribution of mitochondrial haplotypes in a local scale indicated reduced migration on the strongly fragmented southern riverside. Especially, oil palm plantations seem to negatively affect dispersal in T. longipes.

51 Clear sex-biased dispersal was not detected based on relatedness, assignment tests, and haplotype diversity. This study revealed the importance of landscape connectivity to maintain migration and gene flow between fragmented populations, and to ensure the long-term persistence of species in anthropogenically disturbed landscapes.

Keywords: Forest fragmentation, migration, Tupaia longipes, Borneo, microsatellites, cytochrome b, population structure, genetic differentiation, sex-biased dispersal

3.1 Introduction Deforestation of tropical rain forests causes severe problems for the maintenance of biodiversity and ecological functions worldwide (Achard et al. 2002; Sodhi et al. 2010; Edwards et al. 2014). Agricultural expansion, logging, and urbanisation often results in a matrix of altered terrain surrounding original habitat patches. Landscape matrix features and inherent ecological and behavioural plasticity determine the permeability of the surrounding landscapes for a given species (Edwards et al. 2014; Prevedello and Vieira 2010). Species unable to penetrate the surrounding matrix may experience limits to their movement between habitat patches. Restricted migration between populations results in a reduction in gene flow and connectivity which may lead to a decrease in viability and persistence of isolated populations (Edwards et al. 2014; Broquet and Petite 2009). However, the extent to which habitat fragmentation has a negative effect on the genetic structure and persistence of animal populations remains debated (Fahrig 2017). While some studies have demonstrated a pronounced effect of habitat fragmentation on species’ population genetic structure (e.g. Stow et al. 2001; Sommer 2003; Goossens et al. 2006; Stephens et al. 2013; Janecka et al. 2016; Bani et al. 2017), others have failed to detect these effects (e.g. Watts et al. 2006; Benedick et al. 2007; Mbora and McPeek 2010; Janecka et al. 2016). In this context, a better understanding of how and the degree to which fragmentation affects species will help to understand demographic processes that are triggered by fragmentation events. One of the highest rates of deforestation in the tropics has occurred on Borneo (Sodhi et al. 2010). Much of the original large forested areas on Borneo have already been lost due to logging or forest conversion to agriculture (Miettinen et al. 2010; Gaveau et al. 2016). The remaining forests are restricted to isolated patches in many locations. Along the Kinabatangan River of north-eastern Borneo, much of the original lowland dipterocarp forest was cleared or converted within the last 30 – 40 years (Goossens et al. 2006; Latip et al. 2013). Nowadays, oil palm plantations dominate the landscape along the river and forest is restricted to isolated patches. Despite this extensive land conversion in the Kinabatangan floodplain, a highly diverse small mammal community remains in the 52 forest remnants along the Kinabatangan River (Chapter 2). Diverse responses to landscape features along the Kinabatangan River itself have recently been reported for a suite of small mammals in this anthropogenically modified region (Chapter 2). Dispersal and gene flow have been shown to be limited by the Kinabatangan River predominantly in squirrels and tree shrews. However, signals of high genetic differentiation between populations have suggested additional limitations in terrestrial vagility along riverbanks for these taxa. The consequences of habitat fragmentation on dispersal and genetic diversity are still largely unknown for Bornean non-volant small mammals. The relative scarcity of systematic data on the effect of fragmentation in Bornean small mammals is alarming,as they provide important ecosystem functions and services for their natural habitats (Jones and Safi 2011). They are important seed dispersers, pollinators, invertebrate and seed predators, and are prey for larger predators (Emmons 2000; Shanahan et al. 2001; Wells et al. 2009; Jones and Safi 2011). Along the Kinabatangan River, habitat connectivity differs considerably on both riversides. While forest connectivity is high on the northern riverbank, it is disrupted (mainly by oil palm plantations) on the southern side. This setting provides a heterogeneous landscape highly suitable for estimating and comparing migratory activities and genetic structures under different habitat configurations. The Bornean tree shrew (Tupaia longipes Thomas 1893, Scandentia, Tupaiidae) is endemic to Borneo, is widespread throughout the island and is known to inhabit fragmented forest patches such as those along the Kinabatangan River (Chapter 2). Although listed by the IUCN as Least Concern (IUCN 2019), T. longipes has a declining population trend throughout its range due to deforestation and habitat degradation, and is therefore listed in CITES Appendix II (UNEP-WCMC 2014). It is restricted to habitats with dense understory, and occurs in virgin forest, logged forest (Wells et al. 2007), but also in tree plantations with dense understory (Nakagawa et al. 2006; Phillipps and Phillipps 2016). In habitats without dense understory T. longipes is largely absent (Bernard et al. 2009). Being a terrestrial surface gleaner, it feeds mainly on insects and other arthropods, but also on fruits. It is a very agile tree shrew species with large home ranges and large daily travel distances (up to 2 km) compared to other tree shrews (Emmons 2000). Based on behavioural observations, a facultative monogamy is the dominant social system suggested for tree shrews, including T. longipes (Martin 1968; Kawamichi and Kawamichi 1979, 1982; Emmons 2000). Various olfactory and acoustic signals govern their social interactions (Binz and Zimmermann 1989; Schehka et al. 2007). Although female- biased dispersal is known from other tree shrew species (Munshi-South 2008), and Wells et al. (2008) suggested that sex-biased dispersal may be found in a variety of Bornean small mammals including T. longipes, no information exist on its dispersal regime so far. Here we analysed and compared nuclear and mitochondrial genetic diversity and genetic structure in T. longipes subpopulations from both sides of the Kinabatangan River (Fig. 3.1) to assess constraints

53 in migration and gene flow due to relatively recent habitat fragmentation. If this fragmentation limits migration in T. longipes, genetic discontinuities and genetic differentiation should be higher in landscapes comprising isolated forest fragments, compared to those that are contiguous and uniform. Furthermore, habitat alterations may affect dispersal of males and females differently,if they possess sex-biased dispersal (Banks et al. 2007). We therefore investigated sex-specific dispersal in males and females. By identifying potential disruptions in gene flow, we will compare the importance of landscape modifications for generating genetic structure and will provide knowledge for the developmentof effective conservation measures and landscape management plans that help to improve population connectivity in anthropogenically disturbed landscapes.

Figure 3.1 Map of the Lower Kinabatangan Wildlife Sanctuary (LKWS) showing the distribution of sampling locations (circles) and respective forest sites (coloration). The smaller black square highlights the study area within the LKWS.

3.2 Material and Methods Sampling The study was conducted in 18 forest sites and one plantation site along the Lower Kinabatangan Wildlife Sanctuary (LKWS) in eastern Sabah between 2011 and 2013 (Fig. 3.1). Sampling locations in close spatial proximity and without separation by a geographic feature were considered as one site for most analyses, as shown in Figure 3.1. For the present study, forest sites NCa and NCb were pooled to site NC due to small sample sizes. Small ear tissue biopsies (obtained from the ear pinnae) were collected from T. longipes individuals that were live-trapped in the LKWS. All individuals were released back to the wild at their individual capture locations after handling. A more detailed description of the study area, sites and the sampling procedure can be found in Chapter 2. Sampling and handling protocols were reviewed and approved by the Institute of Zoology, University of Veterinary Medicine Hannover, Germany, Cardiff University, UK, and the Danau Girang Field

54 Centre, Malaysia. Official statements from ethical committees are not required under German and UK law for research carried out abroad. All field protocols reported in this study adhered to the legal requirements of Malaysia and the state of Sabah. All methods were officially approved by the Economic Planning Unit Malaysia (Permit No.: UPE:40/200/19/2871) and the Sabah Biodiversity Centre. This research also adhered to the guidelines of the American Society of Mammalogists (ASM; Animal Care and Use Committee, 2011) for the ethical handling of animals. Samples were transported under the permits of CITES [Malaysian Export-Permit No. JHL(PB)600- 3/18/1/1Jld.10/(103), Certificate No. 0602 and Export-Permit No. JHL(PB)600-3/18/1/1Jld.10/(494), Certificate No. 0689 and 0690; German Import-Permit No. E-05027/12 and E-05957/13] and the Sabah Biodiversity Centre [Export-Licence No.: JKM/MBS.1000-2/3(38)].

Molecular methodology DNA was extracted from T. longipes tissue biopsy samples from 116 individuals following a HotSHOT extraction protocol (Truett et al. 2000). All samples were genotyped using microsatellite (msat) primers described by Munshi-South and Wilkinson (2006), and Liu and Yao (2013). Out of 14 available primers, only eight msat loci were applicable on the T. longipes samples used in this study (Appendix 3.6.1). All forward primers were fluorescently labeled and PCR reactions were performed in 10 µl with 5 µl Multiplex PCR Master Mix (Qiagen, Hilden, Germany), 0.1 µl Q-solution (Qiagen), 0.2 pmol of each primer, and 1 µl template DNA. DNA amplification was carried out in three multiplexes (M1 – M3) and two single reactions (S1 – S2; Appendix 3.6.1) with an initial denaturation at 95°C for 15 minutes, followed by 35 cycles of 45 seconds at 95°C, 1:30 minutes at varying annealing conditions (Appendix 3.6.1) and 1 minute at 72°C, and a final extension at 72°C for 10 minutes. PCR products were analysed at Dundee Biosciences, Scotland. Allele length was determined using GeneMapper version 5.0 (Applied Biosystems, USA). Homozygote samples were reanalysed at least twice to minimize genotyping errors. A subset of 60 samples was sequenced at the mitochondrial cytochrome b (cyt b) locus to support the msat-analyses. For the sequencing, an optimised primer (L14841tupaia) together with the primer MVZ16 was used as described in Chapter 2. All cyt b haplotype sequences were deposited in GenBank under the accession number MK111987 – MK111997 and MT013304 – MT013306 (Appendix 3.6.5).

Genetic diversity, genetic differentiation and population structure For each microsatellite locus the presence of null alleles was assessed using the software Microchecker v 2.2.3 (van Oosterhout et al. 2004). Genetixv 4.05.2 (Belkhir et al. 1996) and FSTAT v

2.9.3.2 (Goudet 1995) were used to estimate the observed (Ho) and the unbiased expected

55 heterozygosity (He), and the fixation coefficient (FIS) as an indicator for inbreeding. These estimates were obtained for all loci, for each site, and for each side of the river (Table 3.1). Hardy-Weinberg equilibrium (HWE) was tested for all loci and sites with 10 000 permutations using Genetix v 4.05.2. Allelic richness was calculated with FSTAT v 2.9.3.2 by calculating the standardised allelic richness for each locus and analysing the respective average per site and for each side of the river. Linkage disequilibrium (LD) among all pairs of loci was estimated using the correlation coefficient of Weir (1979) and significant departures from LD were assessed by 10 000 permutations using Genetix v 4.05.2. Pairwise genetic differentiation between forest sites was explored with Wright’s F-statistic according to Weir and Cockerham (1984) implemented in Genetix v 4.05.2 with 10 000 permutations. Pairwise

FST-values between sites (with n ≥ 10) from the same riverside and from different riversides were assessed and compared to the overall set of pairwise comparisons. Conversely, interindividual relatedness was used as a measure of genetic similarity between forest sites. For this, the relatedness coefficient (r) of Queller and Goodnight (1989) was calculated in Kingroup v 2.0 (Konovalov et al. 2004) for all possible pairwise comparisons. Mean relatedness was calculated for all dyads, and for the northern and southern subset separately. Between riverside comparisons were calculated, and within riversides, within and among site comparisons were assessed for all forest sites with n ≥ 10. Significances between riversides (α = 5%) was tested for genetic differentiation and relatedness with a Mann-Whitney U Test implemented in R v 3.5.0 (R Core Team 2018). In order to investigate population genetic structure in T. longipes along the Kinabatangan River, Bayesian clustering was used to assign individuals to population clusters based on their genotypes without prior information on their sampling sites. The analysis was performed in STRUCTURE v 2.3.4 (Pritchard et al. 2000; Falush et al. 2003) with k values ranging from 2 to 8. Each value of k was tested 20 independent times, under an admixture model with correlated allele frequencies between clusters and with each STRUCTURE run lasting 500 000 iterations of the Markov chain Monte Carlo (MCMC) algorithm discarding the first 20% steps as burn-in. To determine the most probable number of population clusters, the likelihood of the data (LnPr(k); Pritchard et al. 2000) and its rate of change (Δk; Evanno et al. 2005) were inspected. For further examinations of potential sub-structuring along both riversides of the Kinabatangan River, individuals from the northern and southern riversides were analysed separately. Genetic diversity within T. longipes was assessed by identifying mitochondrial haplotypes and by estimating haplotype (h) and nucleotide (π) diversity using DnaSP v 5.10.1 (Librado and Rozas 2009). Haplotype genealogies were visualised with a minimum-spanning network computed in Arlequin v 3.5. (Excoffier and Lischer 2010).

56 Migration rates To estimate the recent migration rates between riversides and between forest sites within each riverside, a MCMC analysis was run for each dataset using the software BayesASS v. 3 (Wilson and Rannala 2003). For each analysis, 10 million MCMC iterations were performed discarding the first 20% as burn-in. For each dataset three independent runs with different seed numbers were carried out, and convergence of the three runs was assessed by comparing the posterior estimates of each parameter (all three runs for each test gave the same results).

Sex-biased dispersal A possible sex-bias in dispersal was assessed by comparing male and female relatedness together with differences in male and female population assignment indices. The relatedness coefficient (r) of Queller and Goodnight (1989) was calculated for all possible pairwise comparisons in Kingroup v. 2.0 (Konovalov et al. 2004). Mean relatedness was calculated for all male/male and female/female dyads and for males and females within and among sampling locations within riversides. The number of related dyads was determined for each sex. For each dyad the likelihood of a relationship of r ≥ 0.25 (parent – offspring, full siblings, aunt/uncle – nice/nephew, half siblings) was calculated in Kingroup against the null-hypothesis of being unrelated, estimated using 10 000 simulations from the allele frequency data. Dyads were classified as related when respective likelihood ratios reached α = 5% significance level. Furthermore, inter-individual spatial proximities of related dyads were assessed by measuring the closest (Euclidian) distance of two related individuals based on trapping data implemented in the Animal Movement extension in ArcView GIS 3.3 (ESRI, Inc.). The relatedness analysis was complemented with corrected assignment index (AIc) of Goudet et al. (2002). The assignment index is centered on zero and gives the probability of an individual’s genotype occurring in the sampled sub-population compared to that by chance alone (Goudet et al. 2002). While negative AIc values indicate a genotype less likely than average to occur in the sampled population and characterises possible immigrants, a positive AIc value indicates probable residents. Therefore, the dispersing sex should be characterised by a lower mean AIc with a higher variance compared to the more philopatric sex (Goudet et al. 2002). Individual AIcs were calculated using the R package hierfstat v 0.04-22 (Goudet and Jombart 2015) and mean AIc (mAIc) and variances (vAIc) were analysed for each sex separately. The two riversides were inspected separately to infer possible physical barriers on movements. A Mann-Whitney U Test was used to detect significant sex differences in relatedness and mAIc, and Levene’s test to detect differences between sexes in vAIc. All tests were performed using basic packages in R v 3.5.0 (2018). Finally, (matrilineal) mitochondrial markers represent another tool to infer sex-biased dispersal. Maternally related males and females share the same haplotypes, differences in haplotype diversity,

57 the presence of sex-specific haplotypes and unique haplotypes within sampling locations (haplotype singletons), represent thus signals of immigration, and should be higher in the dispersing sex. Therefore, mitochondrial haplotype diversity was compared between males and females to infer sex- biased dispersal in T. longipes according to Chapter 2.

3.3 Results Genetic diversity and population structure along the Kinabatangan River Individuals of T. longipes were present in all except one forest site (site SH), and were absent in the plantation site (site SP, Fig. 3.1). These sites were thus not included in the following analyses. All eight loci (Appendix 3.6.1) were polymorphic, but null alleles were present in locus TB 14. The number of alleles per locus varied between 4 and 36 and allelic richness between 2.670 and 7.115 (Table 3.1). Locus TB 14 had a significant Hardy-Weinberg equilibrium (HWE) departure across the whole dataset and most forest sites and thus was removed from further analyses (Table 3.1). A significant departure from HWE was present in locus TB 8 for site NA and site ND, however, as no overall HWE departure was observed this locus was kept. While a significant overall departure from HWE could be detected in locus TB 18, HWE-departures were not evident at individual forest sites for this locus, hence it was retained. Testing for Linkage disequilibrium (LD) with Bonferroni correction, seven pairwise comparisons were significant but with no consistent pattern across remaining loci.

The values of observed (Ho) and expected heterozygosity (He) as well as the inbreeding coefficient

(FIS) were not significantly different between riverbanks (Nnorth = 4, Nsouth = 4; Ho: p = 0.110; He: p =

0.149; FIS: p = 0.773; Table 3.1), while allelic richness was significantly lower south of the river (N north =

4, Nsouth = 4, p = 0.021; Table 3.1). Across the 60 mitochondrial cytochrome b (cyt b) sequences (693bp length), 28 variable sites (23 transitions, 5 transversions) and 14 haplotypes were identified. Haplotype diversity (h) ranged from

0.587 to 1.000, and nucleotide diversity (π) from 0.004 to 0.012 (Table 3.1). Neither haplotype (Nnorth

= 4, Nsouth = 4; p = 0.149) nor nucleotide diversity (Nnorth = 4, Nsouth = 4; p = 0.186) differed significantly between the two riversides (Table 3.1).

58 Table 3.1 Genetic characteristics of analysed loci (upper half) and sites (lower half). Number of alleles per locus (Na), the size range of each locus, allelic richness, unbiased expected heterozygosity (H e), observed heterozygosity (Ho), fixation index

(FIS), deviations from HW equilibrium, total number of analysed samples (n), number of males and females, number of sequenced samples and haplotypes, haplotype diversity (h) and nucleotide diversity (π) for each site and riverside, respectively. Size range Allelic Sites out of Locus Na [bp] richness He Ho FIS HWE Js 22 7 172-186 2.670 0.496 0.483 0.027 Js 183 11 134-144 4.373 0.794 0.759 0.044 Js 188 15 182-201 4.542 0.789 0.759 0.039 TB 8 4 404-420 2.831 0.621 0.595 0.043 NA, ND TB 14 36 457-569 7.115 0.966 0.482 0.503** all, exc. SI TB 15 31 284-344 6.601 0.942 0.905 0.040 TB 16 16 170-203 4.174 0.697 0.655 0.060 TB 18 32 408-544 6.452 0.933 0.888 0.049* Males/ Allelic Sequences/ Area 2 Site n Females richness He Ho FIS Haplotypes h π [km ] NA 11 6/5 4.911 0.725 0.727 -0.004 8/6 0.893 0.008 22.17 NB 18 10/8 5.105 0.788 0.786 0.003 7/4 0.714 0.007 41.60 NC 5 4/1 4.571 0.698 0.771 -0.119 2/2 1.000 0.012 9.66 ND 17 9/8 5.013 0.731 0.714 0.024 10/5 0.800 0.008 73.17 North 51 29/22 4.962 0.761 0.748 0.017 27/11 0.889 0.009 146.60 SE 11 8/3 4.257 0.740 0.753 -0.019 6/3 0.733 0.005 47.99 SF 21 9/12 3.902 0.685 0.674 0.018 7/4 0.810 0.005 1.25 SG 29# 14/14 3.888 0.689 0.714 -0.037 17/3 0.581 0.004 17.14 SI 4 3/1 3.286 0.628 0.571 0.103 3/2 0.667 0.012 16.14 South 65# 34/30 3.895 0.718 0.699 0.026 33/6 0.695 0.006 82.52 All 116# 63/52 4.414 0.753 0.720 0.044** 60/14 0.878 0.008 229.12 *p ≤ 0.05, **p ≤ 0.01, #one individual of unknown sex

FST-values varied between 0.0040 (between site NA and ND) to 0.0847 (between site ND and SE) and most were significant (N = 15, mean FST = 0.0445; Appendix 3.6.2). On average, FST-values were higher between sites located on different riverbanks (N = 9, mean FST = 0.0575) than on the same riverside

(N = 6, mean FST = 0.0249), with, FST-values between southern sites (N = 3; FST = 0.0353) higher than between northern sites (N = 3, mean FST = 0.0144; p = 0.050; Appendix 3.6.2 and Appendix 3.6.3). A similar pattern was found in the mean relatedness between forest sites, with values ranging from -0.1054 (between site ND and SE) to 0.0532 (between site NA and ND; Appendix 3.6.2). Overall no relatedness exists between the 5671 analysed dyads (r = -0.0021), and between dyads from different riversides (N = 2806, r = -0.0467), but mean relatedness was relatively high within riversides (N = 1015, r = 0.0416). Relatedness was higher in the southern (N = 1830, r = 0.0616) than in the northern subset (N = 1035, r = 0.0064; p < 0.001; Appendix 3.6.2 and Appendix 3.6.3). In particular, within southern forest sites a high mean relatedness was found between dyads (N north = 344, rnorth = 0.0247;

Nsouth = 671, rsouth = 0.1113; p < 0.001), but also among forest sites mean relatedness was higher in the south (N = 1159, r = 0.0329) than in the north (N = 691, r = -0.0028; p < 0.001; Appendix 3.6.2 and Appendix 3.6.3).

59 Although the low number of sites along each riverbank prevented an unbiased isolation-by-distance analysis, no obvious correlation between pairwise genetic differentiation (FST) or mean relatedness and geographic (Euclidian) distance was observed on either riverbank (Appendix 3.6.2). However, northern riverside FST-values were lower and mean relatedness was higher between the most remote sites (site NA and ND), while the opposite was observed between the most distant southern forest sites (site SE and SG; Appendix 3.6.2). STRUCTURE analysis suggested the existence of two population clusters within T. longipes with the highest likelihood (LnP(D): -3112.2; Fig. 3.2). As the method of Evanno et al. (2005) is prone to produce biases toward k = 2 (Janes et al. 2017), the next highest Δk values at k = 3 and k = 6 were also considered. However, both showed no further geographically meaningful clustering of individuals. At k = 2 (used for further analyses) 103 out of 116 individuals could be assigned to one particular cluster with a q > 80% probability (Appendix 3.6.4). Individuals from the northern riverside were almost exclusively assigned to cluster I. On the southern riverside individuals from sites upstream were mostly assigned to cluster II, while those from sites downstream were assigned to either cluster (Fig. 3.2, Appendix 3.6.4). A progressive partitioning approach for the northern and southern sample subsets yielded solely two sub-clusters in the southern subset. However, these provided no further geographically meaningful information.

Figure 3.2 Spatial distribution of mitochondrial cyt b haplotypes (above) and Bayesian STRUCTURE plot (below) showing the membership of individuals for k = 2 clusters (based on nuclear microsatellite genotypes)

60 Among the 14 mitochondrial haplotypes, 11 were found in samples from the northern side and only six occurred in the southern subset (Table 3.1, Appendix 3.6.5). The haplotype network revealed only three haplotypes (Tl 1, Tl 8, Tl 9) that were shared between riversides (Fig. 3.2, Appendix 3.6.6). On the northern riverside one haplotype (Tl 7) was shared between all sites and another haplotype (Tl 11) occurred in spatially distant sites (sites NB and ND, Fig. 3.2). On the southern riverside three haplotypes (Tl 1, Tl 6, Tl 14) were shared between the adjacent sites SE, SF, and SG, but did not occur in the more remote site SI. Conversely, two haplotypes (Tl 2, Tl 8) were identified in site SI which occurred nowhere else on the southern riverside (Fig. 3.2).

Migration and gene flow across and along the Kinabatangan River Results obtained with BayesASS indicated a very low proportion of individuals migrating per generation (about one individual per generation in either direction) between riversides (from south to north: 0.0161 ± 0.0137; from north to south: 0.0092 ± 0.0084). Within riversides, mean migration rates among northern forest sites was 0.0617 (± 0.0221), and on the southern riverside 0.0531 (± 0.0364). However, migration rates were not evenly distributed among sites. On the northern riverside, high migration rates (> 0.15) were calculated for site NB to all other sites (Table 3.2). On the other hand, site NB had the highest proportion of residents among all sites and thus did not receive many immigrants (Table 3.2). On the southern riverside the sites SF and SG were the sources of most individuals migrating to other sites with SE and SI being the main acceptors from these fragments, respectively (Table 3.2).

Table 3.2 Migration rates between pairs of forest sites from the northern (NA – ND) and southern (SE – SI) riverside. The values for each forest site (row) are the proportion of migrants (± SD) deriving from another site (column). Migration values ≥ 0.10 are highlighted in bold

Migrants NA NB NC ND SE SF SG SI Residents 0.2014 0.0171 0.0180 0.0185 0.0221 0.0209 0.0175 0.6845 NA ± 0.0383 ± 0.0164 ± 0.0170 ± 0.0174 ± 0.0204 ± 0.0196 ± 0.0166 ± 0.0169 0.0135 0.0135 0.0148 0.0141 0.0172 0.0386 0.0136 0.8746 NB ± 0.0128 ± 0.0129 ± 0.0151 ± 0.0135 ± 0.0159 ± 0.0361 ± 0.0129 ± 0.0436 0.0261 0.1526 0.0253 0.0255 0.0258 0.0271 0.0255 0.6921 NC ± 0.0245 ± 0.0452 ± 0.0236 ± 0.0238 ± 0.0235 ± 0.0249 ± 0.0236 ± 0.0236 0.0133 0.2309 0.0133 0.0140 0.0151 0.0177 0.0133 0.6823 ND ± 0.0129 ± 0.0337 ± 0.0127 ± 0.0132 ± 0.0150 ± 0.0196 ± 0.0126 ± 0.0164 0.0181 0.0233 0.0178 0.0174 0.1509 0.0689 0.0179 0.6857 SE ± 0.0170 ± 0.0213 ± 0.0172 ± 0.0164 ± 0.0619 ± 0.0566 ± 0.0170 ± 0.0180 0.0118 0.0163 0.0117 0.0122 0.0126 0.0737 0.0113 0.8505 SF ± 0.0114 ± 0.0155 ± 0.0112 ± 0.0117 ± 0.0121 ± 0.0747 ± 0.0110 ± 0.0760 0.0093 0.0153 0.0093 0.0092 0.0098 0.1065 0.0092 0.8313 SG ± 0.0091 ± 0.0144 ± 0.0091 ± 0.0091 ± 0.0095 ± 0.0831 ± 0.0089 ± 0.0824 0.0277 0.0455 0.0277 0.0278 0.0279 0.0298 0.1189 0.6946 SI ± 0.0258 ± 0.0348 ± 0.0253 ± 0.0256 ± 0.0258 ± 0.0292 ± 0.0476 ± 0.0259

61 Reconstruction of dispersal in males and females Sex-specific relatedness was assessed in a total of 1953 male/male dyads and 1326 female/female dyads. The overall mean relatedness was low in both sexes. However, while within sampling locations no difference in mean relatedness exists between males and females, among sampling locations (within riversides) higher relatedness exists in male than in female dyads (Table 3.3). The overall number of related dyads (r ≥ 0.25) was similar for males (N = 105, 5.4%) and females (N = 91, 6.9%), but related males were spatially more distant than related females. This pattern was also present within riversides when comparing related males and females from different sampling locations (Table 3.3). Although no significant differences were detected in overall mAIc of males and females in the assignment tests (Table 3.3), the overall male mAIc was positive, suggesting higher proportions of male residents, while the overall female mAIc was negative, indicating higher proportions of female immigrants. In accordance with their negative mAIc, female AIc values showed a higher variance than those of males (Table 3.3). In a subsample of 28 males and 31 females mitochondrial haplotype diversity was assessed as a further determinant of sex-biased dispersal. Although the overall number of haplotypes (males = 11, females = 13), and the number of sex-specific haplotypes (males = 1, females = 3), were slightly higher in females, the number of unique haplotypes at a given site (haplotype singletons: males = 11, females = 10) was rather evenly distributed between sexes, and no clear sex-specific dispersal pattern could be inferred.

62 Table 3.3 Relatedness (r, mean ± SD), amount of related dyads (r ≥ 0.25) and interindividual distance (mean ± SD) for all related male and female pairwise comparisons. Within riversides results of within and among sampling location comparisons are given. The mean (mAIc ± SD) and variance (vAIc) of corrected assignment indices are given for the overall male and female subset. Parameters differing significantly between sexes are given in bold (details of statistical tests are inAppendix 3.6.3)

Males Females related related No. of dyads distance No. of dyads distance dyads r [%] [km] mAIc vAIc dyads r [%] [km] mAIc vAIc ** ** ** ** -0.0032 7.59 0.0310 ** -0.0205 4.24 -0.0419 ** All 1953 ±0.1885 5.38 ±7.92 ±1.6967 2.8790 1326 ±0.1942 6.86 ±5.94 ±2.4774 6.1375 within riversides, within sampling 0.1187 0.06 0.1302 0.06 locations 94 ±0.2101 23.40 ±0.05 81 ±0.2250 27.16 ±0.05 within riversides, among sampling 0.0407** 8.82* 0.0130** 4.95* locations 873 ±0.1811 7.67 ±7.65 585 ±0.1977 9.23 ±5.93 *p ≤ 0.05, **p ≤ 0.01

63 3.4 Discussion The influence of the Kinabatangan River and the landscape on genetic diversity and structure The present study corroborates some findings of Chapter 2 and confirms the river as an important barrier to gene flow in T. longipes. For example, when ignoring the river by pooling the northern and southern samples, a significant excess of homozygosity was visible in the whole dataset, most likely a result of the Wahlund effect (Freeland 2008). Since this effect is absent within the northern and southern subsamples, T. longipes from the two riversides of the Kinabatangan River can be assumed to belong to two genetically distinct sub-populations. This assumption is further supported by the results of the STRUCTURE analysis, in which samples were allocated to two population clusters, with the two clusters largely representing one riverside each. However, the migration rates assessed with BayesASS, the sharing of three haplotypes between riversides and the allocation of some southern individuals to the northern population cluster indicate some degree of genetic exchange between the two riverside subpopulations. Occasional crossings of the river are therefore likely to occur. Genetic differentiation among populations can also be generated without geographic isolation,for example, as a result of past colonization processes (He et al. 2013). The observed patterns of high mtDNA haplotype and low nucleotide diversity suggest an influence of such processes onthe population structure in T. longipes (Grant et al. 1998). Most likely, historical habitat contractions and fragmentation during Pleistocene glaciation, and a later colonization of the Kinabatangan floodplain from glacial refuges has shaped such signals in T. longipes, as proposed for many other animal and plant species of this region (e.g. Barkman and Simpson 2001; Garthorne-Hardy et al. 2002; Cannon and Manos 2003; Jalil et al. 2008). In addition to historical population processes, recent anthropogenic modification (i.e. forest conversion to oil palm plantations) in this region markedly shaped the landscape within the last 30 – 40 years (Goossens et al. 2006; Latip et al. 2013). The shaped landscape differed considerably between riversides. While on the northern side of the Kinabatangan River forests are still connected by a forest corridor, the forest fragments on the southern riverbank are separated mainly by oil palm plantations. Considerable differences in the remaining forest size within the study area (north = 147 km2, south = 83 km2; Table 3.1) further underline the stronger fragmentation of forests in the south. Considering the short generation time of 1 – 2 years in tree shrews with 1 – 2 litter per year (sometimes three, in times of supermasting; Emmons 2000; Munshi-South et al. 2007), genetic patterns in the studied T. longipes population may already reflect this recent landscape fragmentation. For example, low FST-values, high relatedness and shared haplotypes suggested ongoing gene flow even between more distant sites in the contiguous forest on the northern riverside and less in the fragmented southern forest. However, migration rates assessed with BayesASS suggested some (between some forest sites rather low) gene flow between study sites on both riversides. On the northern riverside, site NB seems to act as an 64 important source of immigrants to the other northern sites, possibly because of its central position. Although less pronounced than on the northern riverside, migration also took place between sites on the southern riverbank. Again, migrants originated predominantly from central sites (i.e., site SF and SG). A triangulation study of T. longipes, which was carried out on two animals in one forest fragment along the Kinabatangan River (forest site SG), revealed home range sizes of 13 – 16 ha (Brunke et al. unpublished data; Appendix 3.6.7). Furthermore, high distances traveled per day (~ 2 km) are known for this species (Emmons 2000). Bowman et al. (2002) showed that home range size co-varies with maximum dispersal distances in mammals. Based on the underlying mathematical relationship, a maximum dispersal distance of about 16 km could be suggested for T. longipes. In our study, neighboring sites within each riverside were 6 – 22 km (Appendix 3.6.2) apart from each other and therefore ranged within this potential maximum dispersal distance of T. longipes. This together with a rather low sample size might be reasons why no obvious isolation-by-distance effect (although not explicitly tested) could be detected on both sides of the river. On the other hand, the signals of decreased connectivity and genetic diversity on the southern riverside may be interpreted as a consequence of more intense and/or earlier landscape fragmentation on the southern side of the river. For example, it cannot be finally decided whether migration between distant sites in the north is still ongoing under the current degreeof fragmentation, or if the inferred migration rates may be partly a signal reflecting pre-fragmentation connectivity. While the southern forest sites already started to get fragmented by ongoing landscape modifications (i.e. expansion of oil palm plantations) in early 1980s, forests on the northern side largely retained their connectivity until mid to late 1990s (Appendix 3.6.8). This longer period of forest connectivity north of the river may also explain the lower genetic differentiation among the northern sites. Following the principles of the coalescence theory, high frequency haplotypes represent old alleles, while haplotype singletons and haplotypes with low frequency are likely to have evolved rather recently (Freeland 2008). Moreover, “old” haplotypes are expected to show a broader spatial distribution than more recent ones, because their carriers had longer times to disperse (Freeland 2008). The comparison of haplotype frequencies and the spatial distribution patterns of haplotypes on both riversides indeed suggest longer forest connectivity in the north. Conversely, the earlier fragmentation processes on the southern riverside may have shaped genetic structure by a relatively early reduction of migration between forest patches, resulting in higher genetic differentiation between sites, a lower number of haplotypes, and in an accumulation of “older” haplotypes on this side of the river.

65 In addition, the presence of unique haplotypes in the isolated site SI (and the absence of these haplotypes in other southern sites), the high genetic relatedness within forest sites, and the heterogeneous representation of genetic clusters on the southern side suggests that plantations have further reduced genetic connectivity. Other potential barriers such as roads or tributaries did not seem to have influenced gene flow in T. longipes, possibly because they are not wide enough to affect dispersal in this species (Rico et al. 2007; Russo et al. 2016). These results are congruent with findings in murids of this region (Chapter 2) in which signals of genetic differentiation in mtDNA sequences existed between sites separated by large plantations. Furthermore, other studies have demonstrated that T. longipes is absent in habitats with poor understory structures, such as oil palm plantations (Bernard et al. 2009), and this was also the case in our study (site SP; Fig. 3.1). As it is known that T. longipes prefers to move in habitats with dense understory (Emmons 2000), a disrupting effect of oil palm plantations (in which dense understory is largely absent) on connectivity is therefore highly likely.

Dispersal of males and females along the Kinabatangan River Sex-biased dispersal should be reflected in the relatedness structure within populations because the dispersing sex will transfer alleles that do not correspond to the local allele distribution and thus can be identified by a lower overall relatedness compared to the more philopatric sex (Prugnolle and de Meeus 2002). Overall relatedness was low in both sexes, but a higher mean relatedness in males than in females from different locations, suggesting female-biased dispersal inT. longipes. In contradiction, however, suggest the higher distance of related males a farther reaching dispersal in males. The results of the assignment test were also ambiguous: both sexes had quite low and similar mAIc values (= low likelihoods to be resident), whereas females had higher vAIc values (= more variability) which would be expected for the dispersing sex. To detect a signature of sex-biased dispersal using mAIc, actual differences between the sexes should be substantial (not less than 80:20; Goudet et al. 2002). Following this guideline, T. longipes would have a rather unbiased dispersal system. The evenly distribution of haplotypes between sexes further support this. Tupaia longipes forms strong pair bonds with a pronounced intrasexual territoriality, and female home ranges are large (Emmons 2000; Brunke et al. unpublished data, Appendix 3.6.7), males are thus not able to monopolise more than one (sometimes two) female(s). Munshi-South (2008) hypothesised that the energetically-expensive absentee maternal care system, observed in tree shrews (Martin 1968; Emmons 2000), restricts the ability to rear young on poor-quality territories, and produces intense competition between females for resources (Munshi-South et al. 2007). In order to avoid competition and to ensure their own reproductive success, females may be driven to disperse from their natal site to settle in a distant, high-quality territory. Female dispersal may be

66 driven by this pattern in T. longipes, whereas the farther male dispersal may be a consequence of female dispersion and intense male-male competition for access to females in T. longipes. However, this hypothesis needs to be explicitly tested to assess the driving factors for male and female dispersal in this species.

Conclusion This study demonstrates that the Kinabatangan River and other landscape features have most likely shaped the genetic structure of T. longipes populations in historic and in recent timescales. Moreover, it demonstrates that anthropogenic forest conversion starting 30 – 40 years ago (Goossens et al. 2006; Latip et al. 2013) can already be detected by increased genetic differentiation and a decrease in gene flow between populations of T. longipes. In particular, signals of restricted gene flow were pronounced on the strongly fragmented southern riverside, possibly as an additive effect of earlier and stronger landscape modifications (i.e. expansion oil palm plantations). There have been contrasting results on the effectiveness of corridors to reduce effects of habitat fragmentation (Beier and Noss 1998; Mech and Hallett 2001; Haddad et al. 2003). However, Bruford et al. (2010) showed that the establishment of forest corridors in the Kinabatangan area can help to retain demographic stability of isolated orangutan (Pongo pygmaeus) subpopulations. It has been demonstrated that corridors need to be sufficiently wide to reduce the vulnerability to edge effects and increase its structural habitat heterogeneity (MacDonald 2003; Haddad 2008). Lees and Peres (2008) proposed a minimum corridor width of about 400 m to provide appropriate habitats for various species. The corridor width on the northern Kinabatangan riverbank ranged between 140 – 2,000 m, and thus should serve well to maintain migration and to reduce genetic isolation between demes. Moreover, it contributes in the enlargement of effective forest expanse in the north. On the southern side a corridor is largely absent, only a few forest trees with scattered availability connect some of the southern patches along the Kinabatangan River. In our study, this seems not to be effective in providing sufficient habitat connectivity. Possibly because, corridors should reflect natural landscape features, which do not typically take the shape of narrow habitat strips (MacDonald 2003; Gilbert-Norton et al. 2010). Although some degree of gene flow may still be maintained between the southern forest fragments up to the present, the negative effect of habitat fragmentation and isolation on genetic diversity might increase further in future generations by the ongoing agricultural expansion in this region (Abram et al. 2014; Gaveau et al. 2016). This is an alarming trend by which the long-term viability of T. longipes populations may become jeopardized. Considering that according to the IUCN (2019) more than half of the small mammal species confirmed for this region (Chapter 2) show a decline throughout their range, it is important to ensure

67 gene flow between isolated demes that may enable them to avoid the negative effects of demographic isolation. Our study also tried to unravel sex-specific dispersal patterns, to estimate impacts of habitat fragmentation on male and female dispersal. An overall high interindividual relatedness in the T. longipes sub-population could already be verified within the fragmented forests in the south, and inbreeding or kin competition may occur when small habitat patches remain isolated (Banks et al. 2007; Walker et al. 2008). Moreover, to avoid inbreeding or competition pressure modifications in dispersal patterns of males and females may occur with substantial long-term changes in within- population social and demographic dynamics. Whether the dispersal pattern found in males and females of T. longipes along the Kinabatangan River is already influenced by the fragmentation cannot be said for certain, and further comparative studies, both in connected and fully isolated habitats are clearly necessary to investigate the social and behavioral consequences of forest fragmentation and disrupted dispersal patterns in this species. An effective landscape management approach combining the preservation of forest remnants with an effective corridor planning to ensure connectivity between isolated patches will be essential to mitigate ongoing fragmentation effects and ensure long-term survival of animal populations within already altered habitats.

Acknowledgements We thank the Malaysian Economic Planning Unit and the Sabah Biodiversity Centre for permission to carry out research in Sabah. We are grateful to Audrey Adella Umbol for her help with local authorities and Jumrafiah Abd. Shukor for accepting the role of the Malaysian counterpart. We would like to thank the research assistants at the Danau Girang Field Centre for their support in the field. Among them we specially thank Samsir bin Laimun, Petrieadi Ambo Tola, Saroto bin Payar, and Baharudin bin Resake for their endless help. We are furthermore grateful to Josephine D’Urban- Jackson, Rebecca Lawrence and all the students offering their help during field and lab work. The Calenberg-Grubenhagensche Landschaft and the German Academic Exchange Service (DAAD) supported this study during data collection by providing a travel grant for JB. The Danau Girang Field Centre and the Institute of Zoology, University of Veterinary Medicine Hannover, provided financial support during the design of the study, data collection and during data analyses.

68 3.5 References

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71 3.6 Appendix

Appendix 3.6.1 Amplification characteristics of single (S) and multiplex (M) reactions and the respective microsatelliteloci used in T. longipes Annealing Reaction temperature Loci S1 65°C (-1°C) x 35 circles Js22a S2 68°C (-1°C) x 9 circles, TB14 b 58°C x 26 circles M1 58°C x 35 circles Js183 a, TB18 b M2 56°C x 35 circles Js188 a, TB8 b 64°C (-1°C) x 5 circles, M3 TB15 b, TB16 b 58°C x 30 circles aMunshi-South and Wilkinson 2006, bLiu and Yao 2012

Appendix 3.6.2 Upper table: Genetic ([FST-values], upper matrix) and geographic (Euclidean) distances ([km], bottom matrix) among forest sites. Lower table: Mean relatedness (r, upper matrix) and geographic (Euclidean) distances ([km], bottom matrix) among and within forest sites.

Distances

[km]/FST NA NB ND SE SF SG NA - 0.0158 0.0040 0.0438** 0.0396** 0.0475*** NB 9.71 - 0.0235** 0.0544*** 0.0569*** 0.0497*** ND 22.36 15.51 - 0.0847*** 0.0732*** 0.0680*** SE 4.20 7.02 21.78 - 0.0239* 0.0440*** SF 8.56 1.66 16.67 5.92 - 0.0381*** SG 13.58 4.23 11.61 11.58 5.74 - Distance Within [km]/r NA NB ND SE SF SG sites NA - -0.0256 0.0532 -0.0323 0.0118 -0.0092 0.0490 NB 9.71 - -0.0222 -0.1027 -0.0562 -0.0577 -0.0253 ND 22.36 15.51 - -0.1054 -0.0463 -0.0387 0.0712 SE 4.20 7.02 21.78 - 0.0294 0.0049 0.0640 SF 8.56 1.66 16.67 5.92 - 0.0488 0.1184 SG 13.58 4.23 11.61 11.58 5.74 - 0.1140 * ≤ 0.05, ** ≤ 0.01, *** ≤ 0.001

72 Appendix 3.6.3 Results from MWU test (Observed (Ho) and expected heterozygosity (He), inbreeding index (FIS), allelic richness, haplotype (h) and nucleotide (π) diversity, relatedness (r), on the northern and southern riverside, and relatedness (r), interindividual distances of related dyads (r ≥ 0.25), mAIc in males and females) or Levene’s test (vAIc in males and females). Tests were conducted for the whole sets and for within and among forest site/sampling location comparisons. Significant differences are in bold North vs South Within Among forest forest ALL sites sites MWU Test Z p Z p Z p Ho 1.5972 0.110 He 1.4434 0.149

FIS -0.2887 0.773 allelic richness 2.3094 0.021 h 1.4434 0.149 π 1.3229 0.186

FST -1.9640 0.050 r -7.2962 < 0.001 -6.0751 < 0.001 -4.1245 < 0.001 Males vs Within Among Females sampling sampling ALL locations locations MWU Test Z p Z p Z p r -3.1440 0.002 0.4100 0.682 -3.0987 0.002 distance among related dyads -2.9150 0.004 0.0704 0.944 -3.2379 0.0012 mAIc -0.0618 0.953 Males vs Females ALL Levene’s Test F p vAIc 9.6273 0.002

Appendix 3.6.4 Relative proportion of individuals per site and riverside, respectively, which were assigned to one of the clusters under k = 2 (LnP(D): -3112.2), with a probability of q > 80% Cluster I Cluster II (n = 60) [%] (n = 43) [%] NA 81.82 9.09 NB 100.00 0.00 NC 100.00 0.00 ND 94.12 0.00 North 94.12 1.96 SE 9.09 81.82 SF 0.00 80.95 SG 27.59 55.17 SI 27.59 55.17 South 18.46 64.62

73 Appendix 3.6.5 Cytochrome b Haplotype haplotypes in T. longipes (n = 60) frequency and their spatial distribution Accession

(upstream → downstream) along Haplotype Total NA1 NA2 NB1 NB2 NC1 NC4 ND1 ND2 ND3 SE1 SE2 SF SG1 SG2 SI No. the northern (NA – ND) and southern Tl 1 13 - - - - 1 - - - - 1 1 3 5 2 - MK111987 (SE – SI) riverside Tl 2 2 ------2 MK111988 Tl 3 3 2 1 ------MK111989 Tl 4 2 - 1 1 ------MK111990 Tl 5 2 1 - 1 ------MK111991 Tl 6 14 ------2 1 2 8 1 - MK111992 Tl 7 7 1 - - 1 - 1 1 1 2 ------MK111993 Tl 8 2 ------1 ------1 MK111994 Tl 9 4 ------2 - 1 - - 1 - - - MK111995 Tl 10 1 1 ------MK111996 Tl 11 5 - - 3 1 - - - - 1 ------MK111997 Tl 12 1 ------1 ------MT013304 Tl 13 1 - 1 ------MT013305 Tl 14 3 ------1 1 - 1 - MT013306

Appendix 3.6.6 Haplotype network based on cyt b sequences

74 Appendix 3.6.7 Home range area, the maximum (D max) and minimum (D min) home range diameter for the two T. longipes observed with the triangulation method Number of Area D max D min Sex individuals [ha] [m] [m] Male 1 15.6 621.43 305.86 Female 1 13.4 675.71 264.23

Appendix 3.6.8 Landscape changes along the Kinabatangan River between 1985 and time of study (2013). Study sites are marked with green lines (Pictures: Google Earth Pro V 7.3.2, 12/1985 – 12/2013, Kinabatangan River, lat.: 5.465153° long.: 118.071165°. [01/2020])

1985 1990

1995 2000

2005 2013

75 Chapter 4 General Discussion

The forests along the Lower Kinabatangan River have been (and still are) highly impacted by habitat loss and fragmentation. Logging for commercial timber extraction, from early 1950s to late 1980s, and later on deforestation and conversion of forests to plantations shaped this landscape considerably (Boonratana 2000; Latip et al. 2013, 2015; Pimid et al. 2020). Today, oil palm plantations dominate the landscape along the Lower Kinabatangan River, and natural habitats mainly remain as isolated patches. Nevertheless, a remarkably diverse floral and faunal wildlife could be retained within the remaining forests along the Kinabatangan River (Latip et al. 2013, 2015; Abram et al. 2014; Pimid et al. 2020). However, deforestation and the fragmentation of forests impose major threats to the local wildlife. In particular, species dependent on natural habitats for their long-term survival, and unable to disperse between fragments, easily get isolated and their populations become prone to local extinction. While fragmentation effects on larger mammals found in this area have been studied in more detail (e.g. elephants: Alfred et al. 2012; Estes et al. 2012, primates: Ancrenaz et al. 2004, 2005; Goossens et al. 2005, 2006; Stark et al 2012, Röper et al. 2014; Klaus et al. 2017, 2018; Matsuda et al. 2020 or carnivores Evans et al. 2016; Guharajan et al. 2017, Hearn et al. 2017, 2018), such effects on non-volant small mammal assemblages remained largely unknown. Identifying species’ vulnerability to habitat modifications is of high conservation concern and crucial for effective landscape management. In the present thesis two studies were presented which used genetic modelling approaches to explore consequences of habitat fragmentation and the impact of various biogeographical barriers found within this landscape on the partitioning of genetic diversity in non- volant small mammals. In the following, key aspects will be discussed and conclusions will be drawn from these two studies to evaluate the consequences of forest destruction on these taxa.

4.1 Composition of non-volant small mammal assemblages at the Lower Kinabatangan River Although non-volant small mammal assemblages were studied throughout Borneo (e.g. Nakagawa et al. 2006; Wells et al. 2007; Bernard et al. 2009; Charles and Ang 2010; Cusack 2011; Khalid and Grafe 2017; Chapman et al. 2018, 2019; Mohd-Azlan et al. 2019), the composition of the small mammal community in habitats along the Kinabatangan River remained unknown until now. In Chapter 2 the composition of the small mammal communities in the fragments along the Kinabatangan River was briefly described, but a more detailed account of all captured animals and species per study site is provided in Appendix 4.5.1 The underlying patterns are subsequently discussed.

76 4.1.1 Impact of forest fragmentation on community composition Crucial for the assessment of small mammal assemblages and for the studies presented in Chapter 2 and Chapter 3 was the taxonomic classification of captured individuals. Most studies investigating Bornean small mammals (Nakagawa et al. 2006; Wells et al. 2007; Bernard et al. 2009; Charles and Ang 2010; Cusack 2011; Khalid and Grafe 2017; Chapman et al. 2018, 2019; Mohd-Azlan et al. 2019) solely used phenotypic characteristics (e.g. Payne and Francis 2007) for species identifications. However, identification on phenotypic characteristics alone can be tedious and might resultin misidentifications. In particular, murids and tree shrews have shown to be cryptic and hardto identify based on morphometric measurements alone. Moreover, some Rattus species are known to be distinguishable only by molecular genetic methods (Appendix 4.5.3, Rattus rattus). Compared to fragmented forests elsewhere on Borneo (e.g. Bernard et al. 2009; Charles and Ang 2010; Khalid and Grafe 2017; Mohd-Azlan et al. 2019) a highly diverse small mammal community was retained along the Kinabatangan. Applying phylogenetic in addition to morphometric analyses for taxonomic clarifications (Chapter 2), the present study was the first confirming the existence of at least 19 different non-volant small mammal species in the lower Kinabatangan area (Chapter 2 and Table 4.1 or Appendix 4.5.1). The small mammal assemblages consisted of murids (9 species, 52% of all captures), squirrels (4 species, 25% of all captures), tree shrews (5 species, 22% of all captures) and gymnures (1 species, 1% of all captures). Furthermore, some small mammals like for example the giant squirrel (Ratufa affinis) or the endemic Bornean pygmy squirrel (Exilisciurus exilis) could regularly be observed (Appendix 4.5.1), but were never found within traps, suggesting a higher α- diversity than confirmed in this study. Locally four (in site NC) to 14 (in site SE) small mammal species coexist within forest fragments along the Kinabatangan (Appendix 4.5.1). Structural heterogeneity and niche availability within habitats have been suggested to determine the coexistence of species (Delciellos et al. 2016; Wearn et al. 2017, 2019), and may explain variations in the composition of local small mammal assemblages within forest fragments along the Kinabatangan. In particular, larger areas allow the coexistence of various species because more physical space and resources are available (Pardini et al. 2005, 2018; Haddad et al. 2015). Indeed, the average number of coexisting species was highest within the large forest fragments ND (73.13 km2) and SE (47.99 km2), while a noticeable low number of species may result from a rather low structural heterogeneity within the small forest fragment SH (1.2 km 2; Appendix 4.5.1). In general, common residents of disturbed and undisturbed forests found throughout Borneo dominated the small mammal assemblages. Nearly all can be considered as habitat generalists (Table 4.1 and Appendix 4.5.3). Rare Bornean species, which typically have specific habitat requirements such as Chiropodomys sp., Sundasciurus hippurus, Ptilocercus lowii, Tupaia gracilis, were also scarce

77 in forest fragments of the Lower Kinabatnagan River and occurred in much lower frequencies (Table 4.1). Some otherwise common species like Callosciurus prevostii, Tupaia minor, or Echinosorex gymnura occurred also in lower frequencies, most likely because of specific habitat requirements (Table 4.1 and Appendix 4.5.3). Other studies found a similar dominance of habitat generalists in disturbed and fragmented habitats compared to undisturbed forests on Borneo (e.g. Wells et al. 2007; Charles and Ang 2010; Cusack 2011; Khalid and Grafe 2017). Deforestation and fragmentation have been demonstrated to alter habitat structures and resource availability and thereby create conditions favorable for some and less suitable for other species (Wearn et al. 2017). In particular, habitat specialists are prone to be negatively affected, while more flexible habitat generalists may remain unaffected or can be even positively affected by habitat alterations (Edwards et al. 2014; Loveridge et al. 2016; Chapman et al. 2018; Pardini et al. 2018), and thus have the potential to replace less adaptable species. However, forest fragments along the Kinabatangan River seem to have retained the capacity to provide key resources that still allow the coexistence of habitat generalist and species with specific requirements, at least in some forest fragments (Appendix 4.5.1). Again, predominantly in larger forest fragments (i.e., site ND, SE, SG) species with specific requirements were found, supporting the suggestion of higher structural heterogeneity and niche availability within these fragments, and emphasises the need of (large) structurally diverse forest habitats to maintain species diversity along the Kinabatangan River. Especially when considering that more than half of all the species occurring in the Lower Kinabatangan area show a decreasing population trend and two species are considered as Near Threatened and Vulnerable by the IUCN (Table 4.1). Other species commonly found in disturbed as well as undisturbed forests throughout Borneo were, however, not captured nor observed in the forests along the Lower Kinabatangan River at all (Table 4.1). Although, most of them could be considered as habitat generalists, which should allow to adapt easily to changes within habitats, it is possible that an erosion of structural diversity and niche availability encouraged a replacement by other competing species with overlapping ecological requirements, like the abundant M. whiteheadi, S. muelleri or C. notatus. Such a scenario was, for example, proposed for a fragmented forest in Brunei, where invasions of S. muelleri and C. notatus most likely have resulted in the displacement of other native forest species (Charles and Ang 2010). Details about ecological requirements and niche utilization are, however, scarce for most of the Bornean small mammal species. All the more are follow-up studies needed to throw light on the requirements of the generalists and specialist forest species to evaluate causes for local extinctions but also to ensure the long-term persistence of the small mammals inhabiting forest remnants along the Lower Kinabatangan River.

78 Furthermore, more studies are needed to clarify whether species not captured during this study may indeed be absent (or occurring in non-detectable frequency) in the Kinabatangan area. The methods used could have affected the result. Several factors including the census method, trap arrangement and trapping protocol (Pearson and Ruggiero 2003; Astúa et al. 2006; Bovendorp et al. 2017), trap type (Anthony et al. 2005; Thompson and Thompson 2007; Santos-Filho et al. 2015; Ardente et al. 2017; Cepelka et al. 2019), or bait (Astúa et al. 2006; Hice and Velazco 2013; Harkins et al. 2019) can influence the trapping success and bias the conclusions regarding community composition. Local made wire-mesh live traps baited with peanut butter, banana or oil palm fruits were standard in various studies to capture Bornean small mammals (e.g. Nakagawa et al. 2006; Wells et al. 2007; Bernard et al. 2009; Charles and Ang 2010; Cusack 2011; Khalid and Grafe 2017; Chapman et al. 2018, 2019; Mohd-Azlan et al. 2019) and were used to trap small mammals along the Kinabatangan River in the present study (Chapter 2). Other effective and widely used alternatives to sample small mammals are Sherman and pitt-fall traps (e.g. to capture shrews, Pardini et al. 2005; Molur and Singh 2009; Balkenhol et al. 2013; Kelt et al. 2013; Esselstyn et al. 2019; Palmeirim et al. 2019), or camera traps to verify possible trap-shy species and species too large for commonly used traps (e.g. porcupines, de Bondi et al. 2010; Burton et al. 2015, Hobbs and Brehme 2017; Brozovic et al. 2018; Mohd-Azlan et al. 2018; Palmeirim et al. 2019). Furthermore, can a combination of several (frugivorous and insectivorous/carnivorous) baits enhance inventory integrity (Hice and Velazco 2013; Harkins et al. 2019). Due to logistical reasons these methods, have not been used in this study, but a combined usage of different trap-types and a variety of different baits should be considered to improve the number of trappable species in further studies.

79 Table 4.1 Bornean small mammals captured or sighted within forest fragments along the Lower Kinabatangan River and species common throughout Borneo but absent from the Kinabatangan (no capture/sighting). For all species general abundance, abundance in undisturbed and disturbed (logged/secondary forest) forests, their ecology and space use as known from literature, the category and population trend as declared in the IUCN red list of threatened species. For all species found in the Lower Kinabatangan landscape details are given in Appendix 4.5.3 Abundance IUCN undisturbed disturbed Space pop Species n Borneo* forest* forest* Ecology use category trend Captured species Murids Chiropodomys sp. 2 NA NA NA seed specialists A NA NA Maxomys surifer 27 common common common generalist, avoids open areas T LC ↙ Maxomys whiteheadi 206 common common common generalist T VU ↙ Niviventer cremoriventer 148 common common common generalist, forest dependent T LC ? Rattus exulans 8 common (i) rare (less) common generalist T LC → Rattus rattus 12 common (i) rare (less) common generalist T LC → Rattus tanezumi 20 common (i) rare (less) common generalist T LC ↗ Rattus sp. 48 NA NA NA NA NA NA NA Sundamys muelleri 186 common rare common generalist T LC ↙

Squirrels Callosciurus notatus 232 common rare common generalist A LC ↗ Callosciurus prevostii 6 common common common depends on intact mid- to high canopy layers A LC ↙ Sundasciurus hippurus 5 rare rare rare specialist of hard nuts, large territories, A NT ↙ forest dependent Sundasciurus lowii 68 common common common i.a., selective bark feeder S LC →

Tree shrews Ptilocercus lowii 2 rare rare rare insectivorous A LC ↙ Tupaia gracilis 7 rare (e) rare rare forest dependent, very large territories S LC ↙ Tupaia longipes 126 common (e) common common generalist, prefers habitat with dense understorey, S LC ↙ large territories Tupaia minor 4 common common (less) common depends on vines and/or intact mid-storey layer A LC ↙ Tupaia tana 146 common common common generalist, prefers habitat with dense understorey S LC ↙

Gymnures Echinosorex gymnura 16 common common common prefers high-quality and moist habitats T LC ?

80 Table 4.1 continued

Abundance IUCN undisturbed disturbed Space Pop Species n Borneo* forest* forest* Ecology use category trend Sightings Exilisciurus exilis NA common (e) common common bark feeder A DD ? Ratufa affinis NA common common common active in tall trees A NT ↙

Common Bornean species not captured/sighted along the Kinabatangan River Murids Leopoldamys sabanus - common common common generalist, large territories S LC → Maxomys rajah - common (less) common common generalist T VU ↙

Flying squirrels Hylopetes spadiceus - common common common ? A LC ? Iomys horsfieldi - common common common ? A LC → Petaurista petaurista - common common common generalist? A LC ↙ Petinomys setosus - common common common ? A VU ↙

Porcupines Hystrix brachyura - common common common generalist T LC ↙ Thecurus crassispinis - common (e) common common generalist T LC → Trichys fasciculata - common common common generalist T LC ↙

Shrews Crocidura foetida - common (e) common common insectivorous, prefers open areas T LC ? *Abundances according Phillipps and Phillipps 2016; (e) = endemic, (i) = invasive; A = arboreal, S = scansorial, T = terrestrial; LC = Least Concern, NT = Near Threatened, VU = Vulnerable, DD = Data Deficient, ↙ = decreasing population, ↗ = increasing population, → = stable population, ? = unknown

81 4.1.2 Presence and impact of invasive species The taxonomic analyses conducted in the current study have verified the presence of four Rattus species (Rattus exulans, R. tanezumi, R. rattus and Rattus sp.; Table 4.1) in the Lower Kinabatangan area. The species R. exulans, R. tanezumi and R. rattus are listed among the most successful invasive species worldwide, and responsible for considerable reductions of the native fauna in numerous locations where they have been introduced (e.g. Aplin et al. 2003; Harris and MacDonald 2007; Russel et al. 2008; Banks and Hughes 2012; Harper and Bunbury 2015; Loveridge et al. 2016; GISD 2020a, 2020b). Invasive species tend to be habitat generalists able to thrive in a wide variety of habitats and environments and are easily able to invade disturbed ecosystems. A rather low mitochondrial genetic diversity (Chapter 2) and the low number of sampled individuals within the studied forest patches (Appendix 4.5.1) indicate a rather recent introduction of these three Rattus species in the Kinabatangan area. Furthermore, these species were the only ones found in the plantation site, suggesting an introduction of these omnivorous generalists as a consequenceof human-induced landscape modifications and the expansion of agriculture in this region. On Borneo, commensal Rattus species were predominantly found in housing areas (Wells et al. 2014), but were also verified in primary and secondary forest sites (this study; Nakagawa et al. 2006; Wells et al. 2006a, 2014; Cusack et al. 2015; Loveridge et al. 2016; Chapman et al. 2018, 2019; Mohd-Azlan et al. 2019), highlighting their potential to invade forest habitats. However, compared to the other small mammals, invasive Rattus species were found in rather low abundance within forest sites. Elsewhere on Borneo, Rattus rattus for example, occurred frequently in the surrounding matrix but was absent or less frequent in forest fragments (Charles and Ang 2010; Wells et al. 2014; Cusack et al. 2015; Loveridge et al. 2016). Beside structural determinants (Cusack et al. 2015; Loveridge et al. 2016) the presence of native competitive species, like S. muelleri (Charles and Ang 2010) was suggested to prevent R. rattus from invading forest fragments. Possibly, the presence of S. muelleri in high abundance throughout the forests along the Kinabatangan River likewise still prevents a stronger invasion of Rattus species into forest habitats, but this hypothesis needs future investigation. In addition to the three commensal Rattus spp., a fourth Rattus species was present within forests sites along the Kinabatangan River. Although, the morphometric dataset overlaps with that of some other Rattus species, the definite species assignment for this monophyletic Rattus lineage (Chapter 2, see also Appendix 4.5.3) remained unclear. Whether it represents an undescribed Rattus species or belongs to R. argentiventer or R. tiomanicus that are both known to occur on Borneo, needs to be verified in the future. Compared to the three mentioned invasive Rattus species, this Rattus sp. was found more regularly within forest fragments but was absent in the plantation site (Appendix 4.5.1). Furthermore, was it found in partial sympatry with all three other Rattus species and with S. muelleri (Appendix 4.5.1). Mitochondrial haplotype genealogies suggested that Rattus sp. is a species with

82 longer residence in the area than the other Rattus spp. In conclusion, whether the population of this species is native, stable or threatened by the co-existence with the other Rattus spp. and S. muelleri (or other species having overlapping ecological requirements) cannot be answered here, and this lack of knowledge emphasises the need for more comprehensive investigations on ecological dependencies and evolutionary trajectories of this cryptic species.

4.2 Impacts of landscape structures along the Kinabatangan River on species distribution, gene flow and population structure Restrictions in dispersal and gene flow can reduce the genetic variability within isolated populations, and increase the risk of local extinction by decreasing the ability of a species to adapt to environmental changes in the long-term (Frankham 2005; Ralls et al. 2017). In particular, small mammals with low dispersal and colonisation potential should be prone to adverse fragmentation effects and their persistence becomes jeopardized in fragmented landscapes. Therefore, it is important to determine potent barriers decreasing the ability of individuals to migrate and colonise, and to reduce gene flow between populations. The highly fragmented ecosystems along the Lower Kinabatangan River represented an ideal setting for studying fragmentation effects and for identifying important barriers to gene flow. First of all, geographic features (natural: Kinabatangan River and its tributaries; anthropogenic: two-lane paved road, oil palm plantations) are present in this landscape, which were previously emphasized as effective dispersal barriers in other small mammal species (e.g. Gerlach and Musolf 2000; Rico et al. 2007; Fitzherbert et al. 2008; Kennis et al. 2011; Nicolas et al. 2011; Oshida et al. 2011; Savilaakso et al 2014; Yue et al. 2015; Ascensão et al. 2016; Russo et al. 2016; Ji et al. 2017; Grilo et al. 2018; Mazoch et al. 2018). Second, the landscape along the Lower Kinabatangan River differs considerably between riversides in terms of forest connectivity. While large areas of forest remain connected via a forest corridor on the northern riverside, a connecting corridor was absent and forests were more fragmented on the southern side (Fig. 2.1 and Fig. 3.1). This allows the direct comparison of migratory activities, gene flow, and population structure in different habitat settings.

4.2.1 Impacts of anthropogenic landscape modifications Oil palm plantations The landscape along the Kinabatangan River is dominated by oil palm plantations. Especially, on the southern side of the Kinabatangan River, forests are disrupted by this monoculture. Oil palm plantations have a much simpler construction than forests, with fewer canopy layers, less cover, and less diverse structural elements (Fitzherbert et al. 2008; Foster et al. 2011; Meijaard et al. 2018; Luke

83 et al. 2020). Therefore, many taxonomic groups show decreased species richness and abundances in oil palm plantations compared to primary and secondary forests (Foster et al. 2011; Savilaakso et al. 2014; Yue et al. 2015; Wearn et al. 2017; Pardo et al. 2018). In contrast to tree plantations, which can retain a considerable number of animal species (Stuebing and Gasis 1989; Nakagawa et al. 2006; Mang and Brodie 2015), species diversity and abundances are typically low in oil palm plantations (Bernard et al. 2009; Chapman et al. 2019; Mohd-Azlan et al. 2019). So far, only few studies assessed non-volant small mammal assemblages in plantation sites on Borneo (e.g. Stuebing and Gasis 1989; Nakagawa et al. 2006; Bernard et al. 2009; Chapman et al. 2019; Mohd-Azlan et al. 2019). In particular, generalist species were known to occur in oil palm plantations. For example, beside some forest-dwelling species like M. surifer, M. whiteheadi and S. muelleri (Bernard et al. 2009; Chapman et al. 2019; Mohd-Azlan et al. 2019), some Rattus spp., and C. notatus were recorded in this kind of monoculture (e.g. Hafidzi 1998; Andru et al. 2013; Phua et al. 2018; Chapman et al. 2019; Mohd- Azlan et al. 2019). In the current study, only commensal species (R. exulans, R. rattus, R. tanezumi) were captured in the plantation site (Chapter 2 and Appendix 4.5.1). However, the sampling for the present study was conducted only in one plantation in about 1 km distance to the next forest edge. While murids were mainly sampled adjacent to forest edges in other studies (Bernard et al. 2009; Chapman et al. 2019; Mohd-Azlan et al. 2019), M. whiteheadi was also present further inside an oil palm plantation in the study of Chapman et al. (2019). The proximity to forest edges may explain the presence of most native species within plantations, because some species may use plantations during temporary foraging but return regularly to the forest for other activities (Yue et al. 2015; Chapman et al. 2019). Furthermore, was the plantation site of the present study near (about 150 m distance) human habitations, which may partly explain the lack of native small mammals in the current study. The non-volant small mammal diversity within oil palm plantations along the Kinabatangan River may therefore have been partly underestimated in this study, and further studies investigating the small mammal fauna in varying distance to the forest edge are needed to identify species that are able to use this monoculture landscape at least temporarily.

In the current study the importance of oil palm plantations as effective dispersal barrier was inferred with a genetic approach (Chapter 2 and Chapter 3). Significant genetic variation in mtDNA sequences among southern forests suggested a negative impact of forest fragmentation on gene flow in M. whiteheadi, S. muelleri, S. lowii, and T. longipes, but also in N. cremoriventer (Chapter 2 and Chapter 3). For all other small mammal species, however, sample sizes were too small to unambiguously evaluate impacts of oil palm plantations on population structure. Interestingly, two of the species constrained by plantations (M. whiteheadi and S. muelleri) were observed in oil palms in the past (Bernard et al. 2009 Chapman et al. 2019; Mohd-Azlan et al. 2019). This discrepancy could be

84 explained, if these murids occasionally forage in plantations close to the forest edge but do not penetrate further and do not use it for migratory movements. However, the presence of M. whiteheadi in plantation interior (Chapman et al. 2019) suggests that this species may be able to traverse oil palm plantations. Irregular crossings of oil palm plantations may retain some degree of gene flow among forest fragments and may explain the significant but rather weak genetic differentiation in M. whiteheadi. Furthermore, M. whiteheadi and S. muelleri can actively cross the Kinabatangan and may thereby partly mitigate the disrupting effects of oil palm plantations. This mechanism represents an alternative explanation, why genetic differentiation between forestson the southern riverbank was significant but still rather weak in both murids. Likewise, high genetic differentiation suggests restrictions in migratory movements across southern forest fragments in N. cremoriventer. This murid was frequently found in primary and secondary forest (Nakagawa et al. 2006; Wells et al. 2007; Charles and Ang 2010; Khalid and Grafe 2017), where it was most active in the canopy layer (Wells et al. 2006b; Phillipps and Phillipps 2016). Within oil palm habitats it was never observed (Bernard et al. 2009; Chapman et al. 2019; Mohd-Azlan et al. 2019). Possibly, a lack of sufficient canopy resources within oil palm monocultures may prevent N. cremoriventer to use plantations as habitat and may restrict migration between forest fragments. Signals of restricted movements between fragmented habitats were also present in S. lowii. Moreover, high genetic differentiation between populations was obvious across both riversides, indicating overall limitations in migration capabilities (Chapter 2). In other studies S. lowii was found in virgin and disturbed forests but not outside forest habitats (Wells et al. 2007; Bernhard et al. 2009; Cusack 2011; Chapman et al. 2019; Mohd-Azlan et al. 2019). Feeding mainly on bark from certain trees (Whitten and Whitten 1987; Abdullah et al. 2001) this small squirrel species is furthermore rather selective in its diet. This specialisation might preclude an utilisation of oil palm plantations as habitat but also for movements. In addition, dispersal in this species may be locally restricted, possibly associated to the availability of certain resources, which could explain why overall strong genetic differentiation was present on both riversides.

Further genetic parameters such as FST, relatedness, and gene flow were calculated based on a set of multilocus microsatellite genotypes to define the extent of genetic connectivity between populations in T. longipes (Chapter 3). While low FST-values, high relatedness and shared haplotypes suggested ongoing gene flow even between more distant sites in the continuous forest on the northern riverside, a heterogeneous representation of genetic clusters on the southern riverside, high relatedness within southern forest sites, and the presence of unique haplotypes in the isolated forest SI (and the absence of these haplotypes in other southern sites) found in T. longipes implied that aggravated forest fragmentation on the southern riverside caused reductions in genetic connectivity, and identified large oil palm plantations as a limiting factor.

85 Roads Roads dissect natural habitats and have been shown to affect animal behavior (Holderegger and Di Guilio 2010; Kleinschroth and Healey 2017). Several studies on small mammals demonstrated that roads can inhibit movements and cause genetic isolation between populations on both sides (e.g. Gerlach and Muslof 2000; Goosem 2002; Rico et al. 2007; de Redon et al. 2015; Ascensão et al. 2016; Ji et al. 2017; Grilo et al. 2018). A highly frequented two-lane paved road (without fences; build ca. 20 years before this study) traverses the landscape of the Lower Kinabatangan River in north-south direction, dissecting forests on both riversides. The importance of the road as potential dispersal barrier was assessed in the current study (Chapter 2 and Chapter 3). Unfortunately, sample sizes were too small, and direct investigations of the effects of this potential barrier were thus not possible for the majority of the investigated taxa. However, in species sampled in sufficient numbers (M. whiteheadi, S. muelleri, and T. longipes), dispersal was not impaired by the road (Chapter 2 and Chapter 3). This result was unexpected, since roads acted as an almost impermeable barrier for small mammals in other studies (Gerlach and Muslof 2000; Rico et al. 2007; de Redon et al. 2015; Ascensão et al. 2016). However, those roads were at least 30 m wide, whereas the road crossing the Kinabatangan has a width of about 7 m only. Hence, it is quite likely that it is not wide enough to effectively prevent movements of small mammals. On the other hand,a behavioral study stated that forest roads of 10 m width were rarely crossed by Bornean brown spiny rats (Maxomys rajah; Shadbolt and Ragai 2010), indicating the potential even of narrow roads to inhibit movements of Bornean small mammals. Road crossing was suggested to be related to the propensity to explore new areas (Ji et al. 2017; Grilo et al. 2018). Species with stronger habitat specialisation are restricted to few habitat types and should thus show stronger road avoidance than generalist species, which are more prone to exploit new environments. Hence, it is possible that the generalist ecology of M. whiteheadi, S. muelleri, and T. longipes (Table 4.1) account for an unimpaired dispersal across the Kinabatangan Road. Furthermore, it is possible that 20 years since road construction were not long enough to manifest in genetic differentiation. Similar to this study, no genetic differentiation between observedvole populations separated by a 25-year-old motorway was found in a study of Gauffre et al. (2008). Interestingly, since effects of forest fragmentation, starting 30 – 40 years before this study, were detectable in the population genetic structures of the studied small mammals. This may imply that genetic effects of landscape changes may need > 20 years to become recognizable. Moreover, it may provide important information for further genetic studies, since generation times for most Bornean small mammals are scarcely known. Nevertheless, impacts of the Kinabatangan Road on the movements of small mammals cannot be completely rejected and need further genetic and

86 observational investigations to clarify the importance of the road to influence animal movements and population connectivity.

4.2.2 The effectiveness of corridors in fragmented landscapes In fragmented landscapes habitat corridors can enhance habitat connectivity and facilitate movements between otherwise disrupted populations. Nevertheless, the efficacy of corridors has been controversially debated in the literature (e.g. Beier and Noss 1998; MacDonald 2003; Falcy and Estades 2007; Haddad 2008; Gilbert-Norton et al. 2010; Christie and Knowles 2015; Resasco 2019). In the current study different measures of genetic differentiation between demes revealed unimpeded dispersal and gene flow on the northern but not on the southern riverside in the Bornean tree shrew (T. longipes; Chapter 3) and some other small mammals (e.g. M. whiteheadi, S. muelleri; Chapter 2). The forest corridor present on the northern riverside most likely contributes to this stronger population connectivity. On the other hand the distribution patterns of individuals and haplotypes in several of the other studied small mammals suggest rather negative effects of narrow habitat strips on migratory movements. For example, R. exulans, R. tanezumi, Rattus sp., C. prevostii, S. lowii, T. gracilis, T. minor, T. tana, and E. gymnura were absent in sites located in the narrowest part of the forest corridor (i.e. site NCa2 and NCa3, width: 140 – 200m; Appendix 4.5.1) but were present in other northern sites. Furthermore, haplotype sharing was constricted to each deme located on either side of the corridor in T. minor, T. tana, and in S. lowii (Appendix 4.5.2), indicating restrictions to dispersal possibly caused by a limited suitability of the corridor at least for these taxa. However, small sample sizes preclude further statements and further studies will be needed to confirm whether these species do indeed actively avoid rather narrow dispersal corridors. Habitat heterogeneity have been shown to increase with corridors width, influence habitat utilisation and facilitate the coexistence of species within such habitats (MacDonald 2003; Haddad 2008; Lees and Peres 2008; Gilbert-Norton et al. 2010; Zimbres et al. 2017; Beier 2018; Resasco 2019). To reduce the effect of habitat edges and to increase structural diversity within corridors, authors already requested them to be wide enough (> 400 m; Lees and Peres 2008; Zimbres et al. 2017). Indeed, more species were present in wider parts of the northern corridor (i.e. NC a1 and NCb width: > 600m) than in its narrower parts (Appendix 4.5.1), and habitat heterogeneity might have been higher there. Low species diversity and the dominance of generalist species (e.g. M. whiteheadi, S. muelleri, C. notatus) may indeed indicate a reduced structural diversity within the narrower parts of the corridor. Increased ecological diversity and resource variability within the corridor may enhance the number of traversing species. For example, some arboreal species (e.g. Chiropodomys sp., N. cremoriventer, T. minor, P. lowii, and squirrels) are known to depend on sufficient canopy connectivity (Wells et al. 2004 and 2006b), and tree shrews like T. longipes and T. tana are known to preferably roam in dense

87 undergrowth vegetation (Emmons 2000; Wells et al. 2006b). Details about species-specific microhabitat requirements and niche utilization are, however, lacking for most small mammal species.

4.2.3 Impacts of the Kinabatangan River and riverine structures Kinabatangan River In Chapter 2 and Chapter 3 the Kinabatangan River was identified as an important biogeographic barrier, inducing different levels of genetic isolation and differentiation between sub-populations of various small mammal species. The only terrestrial possibility to cross the Kinabatangan River via the road bridge located within the study area (Fig. 2.1 and Fig. 3.1). The road itself was not identified as a dispersal barrier, and it cannot be completely excluded that some small mammals use the bridge occasionally. Nonetheless, it is extremely unlikely for most of them to cross the Kinabatangan River this way, because it would require traversing human settlements plus a surrounding plantation on the southern end. Plantations were identified as unsuitable habitat and migration barrier for most of the studied species (Chapter 2 and Chapter 3). Only commensal Rattus species were confirmed to occur in plantations or near human settlements and should therefore be potentially able to use the bridge. For all other species, the most probable way of crossing of the river would be by swimming, floating or rafting. In a comparative study of three primate species along the Kinabatangan River (Jalil 2006), genetic differentiation between populations from different riversides was pronounced in the orangutan (Pongo pygmaeus), while population genetic structures suggested regular crossings of the river in long-tailed macaques (Macaca fascicularis) and proboscis monkeys (Nasalis larvatus). Compared to orangutans, long-tailed macaques and proboscis monkeys are able to utilise wet habitats and are known to be good swimmers (Matsuda et al. 2008, 2011, 2018; Otani et al. 2012; Mohammad and Wong 2019). Similarly, an ineffectiveness of rivers as dispersal barrier was found in small mammals with good swimming abilities (e.g. Katuala et al. 2008; Kennis et al. 2011; Kalinin and Kupriyanova 2016). Although pronounced swimming abilities are known for various species of small mammals (Evans et al. 1978; Hickman and Machiné 1986; Cook et al. 2001; Tapisso et al. 2013; Torres et al. 2020), this behavioural trait is largely unknown for most of the Bornean small mammals. Like in the primate study (Jalil 2006) good swimming abilities in small mammals were mostly attributed to inhabitants of wetlands or regularly inundated habitats (e.g. Cook et al. 2001; Nicolas and Colyn 2006; Santori et al. 2008; Torres et al. 2020). So far, an affinity to wetter habitats was only described for the moonrat and S. muelleri (see Appendix 4.5.3) and for some Rattus species (Aplin et al. 2003; Russel et al. 2008; Wells et al. 2014; GISD 2020a; personal observations).

88 Since all study sites along the Lower Kinabatangan floodplain are regularly inundated after heavy rainfall, all small mammal species inhabiting this area are periodically confronted by a wet habitat. While arboreal species or those with scansorial abilities (Table 4.1) should be able to use pathways in upper strata to circumvent wet areas, the suite of terrestrial species in these habitats (Table 4.1) may be restricted to those with good swimming abilities. Indeed, haplotype distribution and measures of genetic differentiation revealed strong to moderate levels of genetic separation between subpopulations from different riverbanks primarily in arboreal and scansorial species such as squirrels and tree shrews. In murids, however, no genetic isolation was observed among riverbanks, and the same was true for the moonrat (E. gymnura). In general, rather irregular patterns of mitochondrial haplotype distribution were observed along the river, suggesting regular and/or active crossings by swimming murids and moonrats, while patterns in squirrel and tree shrew haplotype distribution indicate a less frequent (if any) and possibly rather accidental crossing of the river. Although good swimming abilities might enhance the propensity of a species to cross waterbodies such as rivers and may be a prerequisite for terrestrial species to survive in these seasonally inundated habitats, the frequency of actual river crossings can by no means be predicted from this. For example, for both terrestrial Maxomys species swimming abilities can be proposed, but haplotype distribution and low genetic differentiation between riversides suggest frequent crossing of the river in M. whiteheadi, while the presence of M. surifer on only one riverside suggests restrictions in movements across the Kinabatangan. Morphological variations between the two species (Appendix 4.5.3) indicate adaptations to divergent life styles and ecological niches, and may explain why M. whiteheadi crosses the river more frequently than M. surifer. More specifically, in a comparative study M. surifer, but not M. whiteheadi, was considered a gap avoider (Cusack et al. 2015). The avoidance of gaps as a constraint for movements across the river was proposed in Chapter 2 for other species and may also be a possible explanation for the lack of crossings in M. surifer. Furthermore, compared to M. surifer, M. whiteheadi is one of the most abundant species trapped within the study area (Table 4.1 and Appendix 4.5.1), and in extreme fruiting and flowering seasons, population density can considerably increase in M. whiteheadi (Nakagawa et al. 2007), while populations remain relatively stable during masting events in M. surifer (Phillipps and Phillipps 2016). A periodical and massive increase in population size in M. whiteheadi may trigger emigration events (Gaines and McClenaghan 1980; Matthysen 2005; Almeida et al. 2015; Harman et al. 2020) and stimulate animals to overcome barriers including rivers in order to escape intraspecific competition (Bohdal et al. 2016). Further investigations will be necessary to confirm this hypothesis. Interestingly, the ability to cross the river by swimming seems also to facilitate movements along the river. Especially, in species with obvious swimming affinities (i.e. M. whiteheadi, S. muelleri, Rattus spp. and E. gymnura), sub-populations were genetically less differentiated along riversides. Possibly,

89 the use of the river for migration across and along the river may reduce the effects of other barriers including distance along the river. Nevertheless, signals of genetic differentiation were present even in swimming species and suggest that the effects of extensive obstacles, such as large oil palm plantations, can be mitigated but not be evaded completely (as discussed above). Movements and gene flow across the river was particularly restricted in arboreal and scansorial taxa such as squirrels and tree shrews. However, different levels of crossing propensities were found within squirrels and tree shrews, respectively. While haplotype genealogies in the networks of S. lowii and T. tana suggest early genetic separations of northern and southern sub-populations without any haplotype sharing between riversides, haplotype sharing across the river indicates some genetic connectivity between sub-populations in the past for C. notatus and T. longipes. Although active swimming might be a rare event in these species, it cannot be completely excluded and represents a possible explanation for haplotype sharing across riversides in C. notatus and T. longipes as proposed in Chapter 2. Nevertheless, these findings raise the question why such differences exist within these taxonomic groups. Morphology and ecological requirements are for example quite similar in T. tana and T. longipes. However, investigations on tree shrew mobility, conducted prior to the current study within one Kinabatangan forest (unpublished data; Appendix 3.6.7 and Appendix 4.5.3) and elsewhere on Borneo (Emmons 2000), revealed differences in movement patterns. While large home ranges (~ 15 ha) suggest high mobility in T. longipes (see also Chapter 3), smaller home ranges (~ 6 ha) propose a lower vagility in T. tana. Additionally, T. longipes displays a stronger intrasexual territoriality than T. tana (Emmons 2000). Higher vagility and territoriality may coincide with more movements and roaming of non-resident individuals that may lead to more frequent active or accidental river crossings in T. longipes. In contrast, the comparably low dispersal propensity of T. tana may in the long term lead to stronger genetic isolation of sub-populations within and across the river. In the two squirrel species morphology differed considerably. Moreover, they are known to use distinct habitat strata (Abdullah et al. 2001; Wells et al. 2006b). While C. notatus uses all strata up to 10 m, S. lowii is more scansorial and seldom found above 5 m. It is possible that differences in resource specialisation determine dispersal behaviour and river crossing in the two species. For example, S. lowii is more specialized in its diet than the generalist C. notatus, which is known to exploit numerus tree species for feeding (Abdullah et al. 2001). However, whether differences in ranging patterns and territory size exist, is largely unknown for these two squirrels. Yet, the overall high genetic differentiation detected inS. lowii for both riversides suggest restricted terrestrial dispersal, which may be a result of low general vagility, and may explain the strong genetic isolation of sub-populations within and across the river. Probably, a comparably higher vagility may lead to sporadic river crossings in C. notatus, but this hypothesis needs to be investigated in further studies.

90 Tributaries Within the study area three large perennial tributaries (typical width ca. 30 m) of the Lower Kinabatangan River run through the landscape (between site SE and site SF, site SF and site SG, and between site SG and site SH, Fig. 2.1 and 3.1), which were considered to constrain landscape connectivity. In contrast to the Kinabatangan River, no effects of tributaries on gene flow were detected in any of the studied small mammal species, although sample sizes were small in most cases. However, even for those species studied in sufficient numbers (M. whiteheadi, S. muelleri, T. longipes) no effect of tributaries could be observed. Although a lack of power of the statistical tests cannot be entirely excluded, it is likely that the effectiveness of large tributaries as dispersal barrier in the Kinabatangan area is rather weak. These findings may be explained by the smaller size of tributaries. In contrast to the very wide Kinabatangan River (width at study site: ca. 200 m), tributary width was an order of magnitude smaller (see above). Under these conditions, branches of large trees with wide canopies from opposite sides may even meet each other in the middle and may act as canopy bridges facilitating movements of arboreal and scansorial species (e.g. T. longipes). In addition, terrestrial species may cross the tributaries easily by swimming, in particular those for which not even the Kinabatangan acted as barrier (e.g. M. whiteheadi, S. muelleri). Finally, although perennial, water levels fluctuate throughout the year, and some tributaries fall almost dry during very low tides. All of these aspects are possible and non-exclusive explanations for the maintenance of gene flow across tributaries, and the impact of these landscape features on the genetic structure of non-volant small mammals may therefore be rather negligible.

Historic changes in river course Riverine landscapes are highly dynamic. In strongly meandering rivers, such as the Kinabatangan River, fluvial processes regularly produce meander cut-offs and form oxbow lakes. Oxbow lakes are common within the study area on both riversides, indicating frequent changes in the river course over time which might have influenced animal movements across the river. For example, a small mammal population may initially have been located on one riverside inside a river bend. If this river bend was cut off by a sudden change in the river course, e.g., following a major flooding event, the respective population was relocated to the other riverside. Consequently, such an event may generate sudden bursts of gene flow across the river, but without individuals ever swimming and dispersing from one riverside to the other. As all species were studied in vicinity to the river, such historical and natural translocations should have evenly affected the genetic structure of all study species. However, while in species that can actively cross the river, the effects of meander cut-offs should be less consequential, such events may explain signals of past gene flow across the river in species that do not typically cross the river. For example, the river was identified to restrict

91 movements in C. notatus and T. longipes, but haplotype sharing was nevertheless observed between riverbanks in these species (Chapter 2 and Chapter 3). Moreover, in T. longipes genetic similarities between riversides were visible when examining nuclear DNA (msats). Bayesian STRUCTURE analyses, for example, revealed two genetic population clusters, with one cluster dominating one riverside each (Chapter 3; Fig. 3.2), but downstream populations were genetically more mixed on the southern riverside where oxbow lakes occur more frequently (i.e. site SG and site SI). Since oxbow lakes were present along both riverbanks, such a “mixed” structure should normally be visible on both sides. Unfortunately, a historic record is not available for the oxbow lakes in the area. However, it was shown that the higher forest connectivity on the northern riverside allows unimpeded gene flow between northern T. longipes demes, which might lead to a quicker dilution effect after a local release of some southern “immigrants” caused by an episode of river course change. In contrast, such events may cause rather local genetic effects in the southern sub-population as a result of stronger restrictions in gene flow between isolated forest patches.

4.2.4 Genetic inference of historical landscape changes Genetic effects of landscape changes need time to become recognizable in the genome of animals. Depending on the species’ vagility, generation time, and the filtering effect of the new barrier, it can take tens to hundreds of generations until a new gene flow barrier is detectable in population genetic structures (Gauffre et al. 2008; Landguth et al. 2010; Ascensão et al. 2016; Maigret et al. 2020). Therefore, genetic structures do mainly reflect past rather than contemporary connectivity conditions, especially, in recently altered landscapes (Holderegger and Di Guilio 2010; Epps and Keyghobadi 2015). Although small mammals have relatively short generation times (Pacifici et al. 2013), which range between several months (e.g. rats: Ewer 1971; GISD 2020a, 2020b) and up to one year (e.g. tree shrews: Emmons 2000), it can be argued that the rather recent landscape modifications along the Kinabatangan such as oil palm expansion (starting 30 – 40 years ago) or road construction (about 20 years ago) may not yet have generated their full signature in the genomes of the study species. Indeed, despite the high levels of habitat fragmentation along the Kinabatangan River, the overall genetic diversity was still quite high in most small mammal species (except in the invasive R. exulans, R. rattus, R. tanezumi). Possibly, these high levels of genetic diversity may still reflect past genetic diversity as proposed for the orangutan population inhabiting this area (Goossens et al. 2005, 2006). Compared to the recent landscape modifications, the river has influenced movements in small mammals already in the distant past, which could partly explain why signals of impaired dispersal were much more pronounced across than along the Kinabatangan River in some species. It has been suggested that genetic markers with higher resolution power, like microsatellites (msats) or single

92 nucleotide polymorphisms (SNPs) represent an alternative (to mitochondrial sequence data) to detect signals of recent genetic isolation (Holderegger and Wagner 2008; Landguth et al. 2010; Salgado et al. 2016; Allendorf 2017). Therefore, mtDNA analysis was complemented with msat genotyping in the second study which focused on one species only, T. longipes (Chapter 3). Compared to the mtDNA dataset, signatures of a reduced gene flow due to forest fragmentation were indeed much clearer in the microsatellite dataset. Similar to an orangutan study in the same area (Goossens et al. 2006), this study showed that despite high genetic diversity, contemporary forest conversion has already decreased gene flow between populations of small mammals. Goossens et al. (2006) also diagnosed a recent population collapse of more than 95% for the orangutans along the Kinabatangan River. It is possible that habitat fragmentation may also have caused population declines in small mammals. However, the assessment of demographic size changes and the detection of genetic bottlenecks was beyond the scope of this study. Future studies would be highly desirable and necessary to disentangle and evaluate the consequences of historic and recent landscape changes on small mammal populations on Borneo.

4.2.5 Impact of sex-specific dispersal patterns Species-specific space use and the resulting population genetic structures are likely to be influenced by sex. Particularly, an asymmetry in dispersal rates or distances between the sexes will generate different patterns of population structure for males and females. In general, the philopatric sex should exhibit a more pronounced genetic structure than the dispersing sex. For example, a pronounced geographic mtDNA haplotype pattern along the river was suggested to be influenced by strong female philopatry in Kinabatangan long-tailed macaques, rather than by landscape barriers (Jalil 2006). Likewise, female philopatry may pose an alternative explanation for the strong genetic differentiation within riversides found in S. lowii. However, based on haplotype distribution in males and females of this species (Chapter 2), no clear sex-bias was found in S. lowii, suggesting that gene flow in S. lowii might indeed be determined mostly by habitat structures rather than by sex-biased dispersal. A detailed evaluation of the haplotype distribution for males and females revealed that dispersal was not sex-biased in the majority of the studied small mammal species, (i.e. N. cremoriventer, Rattus sp., S. muelleri, C. notatus, S. lowii, T. longipes, T. tana). Only the Maxomys rats M. surifer and M. whiteheadi showed signals of male-biased dispersal, the typical mammalian dispersal mode (Greenwood 1980). Although sample size was limited for most studied small mammals, and final statements on the dispersal regimes of these species can therefore not be made, the conclusion remained consistent when evaluated again with a larger sample set of bi-parentally inherited msats in the case of T. longipes (Chapter 3).

93 However, some studies demonstrated that fragmentation can affect the dispersing sex to a greater degree than the philopatric sex, and thus may alter patterns of sex-biased dispersal (e.g. Stow et al. 2001; Banks et al. 2005; Walker et al. 2008; Pierson et al. 2010; Cote et al. 2017; Ferreira da Silva et al. 2018; Klass et al. 2019). Such an effect may explain why this study did not find evidence of female- biased dispersal in T. tana (Chapter 2), although this dispersal pattern was described for this species based on mtDNA and msat data in a large continuous forest (Munshi-South 2008). If female T. tana in a fragmented landscape cannot move as far and regularly from their natal home ranges as they would normally do, female-biased dispersal may no longer be detected on the genetic dataset. Such an effect would also explain the general scarcity of signals of sex-biased dispersal in the studied small mammal species. However, a current lack of information on the dispersal regimes for most studied species living in continuous natural habitats precludes a further evaluation of this hypothesis. Future comparative studies on dispersal behavior in contiguous vs. fragmented habitats are needed for many small mammals to fully evaluate the impact of habitat connectivity on space use, movements and the socioecology of Bornean small mammals.

4.3 Conservation implications Extensive anthropogenic modifications impacted the landscape along the Kinabatangan River and reduced natural habitats considerably. This study indicates that habitat fragmentation jeopardises both species and genetic diversity of the small mammal communities in the Kinabatangan landscape. In particular, species that are not able to cross the river may become (or are already) genetically isolated in forest patches north or south of the lower river catchment of the Kinabatangan River (e.g. S. lowii, T. tana). Furthermore, it can be proposed that species that were found in low abundances and with a patchy distribution (e.g. M. surifer, C. prevostii, S. hippurus, P. lowii, T. minor, T. gracilis; Appendix 4.5.1) within the study area may already represent isolated island populations that raise conservation concerns. Further encroachment of the remaining Kinabatangan forests thus likely increase genetic isolation and may lead to local extinction events in future generations. On the other hand, this study has demonstrated that forest corridors along the northern riverbank do effectively connect forest habitats and can improve genetic connectivity among populations. Conservation strategies that complement the protection of intact habitats with efficient corridor planning leading to the reconnection of formerly isolated forest fragments would thus secure gene flow and thereby guarantee long-term viable populations. It further implies the necessity to preserve and extend the Lower Kinabatangan Wildlife Sanctuary, for example, by restoring encroached forests and placing more forest under protection, to enable unimpeded dispersal between forests along both sides of the river but also to link the Lower Kinabatangan area with other forests reverses. 94 New roads are currently proposed for many forested parts of Borneo including the Kinabatangan area (DGFC 2018; Alamgir et al. 2019; Sloan et al. 2019) to create a better infrastructure and improve connectivity and transportation in rural areas. This study did not (yet) reveal disruptive effects of the Kinabatangan road in any of the studied small mammal species and it is therefore possible that some species cross the road regularly. However, the extension of the road network might increase several threats for natural animal populations. First, crossing animals may get killed in road accidents that will increase with increasing number and length of roads (Laurance et al. 2009; Kang et al. 2016; Kleinschroth and Healey 2017; Monge-Nájera 2018). Second, fragmented forests become easily accessible, which might increase the risks of other anthropogenic threats, such as hunting or the introduction of invasive species and wildlife diseases. Illegal hunting is already known to be practiced in the protected Kinabatangan habitats (Evans et al. 2016; DGFC 2018; Pimid et al. 2020), although this may represent a rather minor problem for small mammals in this area. However, some species are likely to be hunted as pests when entering plantations (e.g. Puan et al. 2011; Azhar et al. 2013; Meijaard et al. 2018; Hood et al. 2019). The introduction of non-native or invasive species from housing areas and plantations (i.e. oil palm) into Kinabatangan habitats was documented in this study and represents a more immediate threat for the native species. The high adaptability of invasive species to habitat changes facilitates their expansion across the Kinabatangan landscape, in which they are likely to replace and outcompete less flexible native species. The presence of at least three invasive Rattus species in some forest fragments of the Kinabatangan area therefore raises serious conservation concerns. The implementation of suitable monitoring strategies and measures preventing a further spread of these invasive species should thus be considered a priority for the conservation of native small mammal communities. Finally, feral dogs or livestock can invade the remaining forests next to human habitations, fields or plantations, and disturb their integrity by killing wildlife, destroying vital habitat structures or spreading parasites and other diseases (e.g. Azhar et al. 2013; Tilman et al. 2017; Meijaard et al. 2018). This study detected genetic consequences of habitat fragmentation in some but not all studied species, suggesting that not all species are equally vulnerable to landscape modifications. It is rather likely that intrinsic traits (e.g. space use and locomotion, activity, mating system, dispersal patterns, resource requirements, microhabitat preferences, interspecific competition) may influence their susceptibility toward various habitat disturbances, forest fragmentation or the presence of invading species. As these intrinsic traits are still not well known for most Bornean small mammals, future research on the socioecology and habitat use of these taxa is urgently needed for the development of efficient conservation actions that prohibit extinctions in the future. Management plans were already developed for some larger mammals in the Kinabatangan region (e.g. Bornean elephant, orangutan, proboscis monkey, Sunda clouded leopard). Something similar could be attempted for

95 smaller mammals, and the integration of the results from this and future ecological studies is needed to conceptualise an integrative management plan that will ensure the long-term persistence of diverse small mammal communities in the Kinabatangan landscape. However, since research on small mammals and their conservation is given only minor priority, and since small mammals are still often regarded as pests by many humans despite their important ecological functions, such efforts should be accompanied by environmental education programs at local villages and schools aiming to raise awareness for the importance of small mammal diversity for ecosystem integrity and functioning. One component of such programs should be to deliver biological knowledge on the native small mammal fauna to which the species-specific information compiled in Appendix 4.5.3 may contribute.

4.4 References

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102 4.5 Appendix

Appendix 4.5.1 Number of individuals captured at each trapping location. Trapping location of the northern (NA-ND) and southern (SE-SI) riverside from up- to downriver as shown in Fig. 2.1 and Fig. 3.1 North South

Species NA1 NA2 NB1 NB2 NCa1 NCa2 NCa3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI Muridae Chiropodomys sp. ------1 ------1 - - Maxomys surifer ------2 16 - - 9 Maxomys whiteheadi 8 3 6 9 7 10 7 23 4 16 14 16 6 - 21 26 18 5 7 Niviventer cremoriventer 9 6 3 19 8 12 2 13 16 7 13 14 4 - - - 13 - 9 Rattus exulans - - 1 - 2 ------2 - 1 1 - 1 - - Rattus rattus - - - 2 1 1 - 1 - - - 2 1 2 1 - 1 - - Rattus tanezumi 1 - 2 1 1 - - - 1 - 1 2 - 10 - - - 1 - Rattus sp. - - 4 1 - - - 1 3 4 6 3 - - - 1 1 23 1 Sundamys muelleri 4 13 5 6 7 5 5 4 3 5 10 20 5 - 14 13 2 60 5 Sciuridae Callosciurus notatus 15 3 14 28 5 8 5 6 17 20 11 40 2 - 13 17 10 13 5 Callosciurus prevostii - - - 3 ------1 1 - - 1 - - - - Sundasciurus hippurus - - - - - 1 - - 2 - 2 ------Sundasciurus lowii - 3 ------7 10 2 6 1 - 13 17 6 - 3 Ptilocercidae Ptilocercus lowii ------1 1 - - - Tupaiidae Tupaia gracilis ------2 2 - 2 1 ------Tupaia longipes 7 4 7 13 2 3 - 3 6 4 8 6 6 - 21 28 4 - 4 Tupaia minor - - - - 3 - - 1 ------Tupaia tana 12 7 10 14 2 - - 11 9 11 4 10 8 - 13 15 12 1 7 Erinaceidae Echinosorex gymnura - - - 4 - - - - 1 1 2 1 1 - - 4 - - 2 Species 7 7 9 11 10 7 4 9 13 10 12 14 10 3 11 10 11 6 10 Size of forest fragments [km2] NA: 22.17, NB: 41.60, NC: 9.66, ND: 73.17, SE: 47.99, SF: 1.25, SG: 17.14, SH: 1.20, SI: 16.14

103 Appendix 4.5.2 Spatial distribution and frequencies of cytochrome b haplotypes for all analysed species. Sampling location of the northern (NA-ND) and southern (SE-SI) riverside from up- to downriver as shown in Fig. 2.1 and Fig. 3.1

North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI MURIDAE Chiropodomys sp. Chiro 1 ------1 ------Maxomys surifer Ms 1 ------13 - - - Ms 2 ------4 Ms 3 ------1 Ms 4 ------2 - - - 1 Ms 5 ------2 Ms 6 ------1 - - - Maxomys whiteheadi Mw 1 ------2 - - - 3 3 - 1 - - - - Mw 2 - - 2 2 - - - 2 1 ------Mw 3 ------2 - - - Mw 4 ------1 - - - - Mw 5 1 - 2 - 1 ------Mw 6 ------1 ------Mw 7 ------1 - Mw 8 ------3 ------Mw 9 ------1 - - - Mw 10 ------1 - - 2 Mw 11 ------1 - - - Mw 12 ------2 - 1 2 ------Mw 13 - - - 2 1 - - - - 2 - 2 - - 1 8 3 1 2 Mw 14 2 - 1 2 2 2 - 7 1 5 5 - 1 - 1 3 6 1 1 Mw 15 ------1 ------Mw 16 ------1 - - - - 1 - - Mw 17 ------1 1 - - Mw 18 ------4 - - - - Mw 19 ------2 ------1 1 Mw 20 ------1 ------Mw 21 1 ------Mw 22 1 ------1 ------

104 Appendix 4.5.2 continued North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI Maxomys whiteheadi Mw 23 1 - - 1 ------2 - 1 - - Mw 24 ------2 ------Mw 25 - - - 2 - - - - - 1 ------Mw 26 - 2 - - - - - 1 - - 2 ------Mw 27 ------1 - - 1 ------Mw 28 ------1 - - - Mw 29 ------1 - - 5 - - - - Mw 30 2 - - 1 - 2 3 4 1 1 4 3 - - 4 5 3 - 1 Mw 31 ------1 - - - - Mw 32 ------1 - - Mw 33 ------1 1 - Mw 34 ------1 ------1 1 - - Niviventer cremoriventer Nc 1 - - 1 ------Nc 2 1 - - 1 4 5 - - - - - 1 1 - - - 3 - - Nc 3 ------1 Nc 4 ------1 1 ------Nc 5 - - - - 1 ------Nc 6 ------2 Nc 7 - - - 1 ------Nc 8 - - - 1 - - - - 1 ------Nc 9 - - - - 2 ------Nc 10 ------1 ------Nc 11 - - - 1 ------Nc 12 - - - 1 - - - 1 - - 1 - 1 ------Nc 13 - 1 ------Rattus exulans Re 1 - - 1 - 1 ------Re 2 - - - - 1 ------2 - 1 1 - 1 - - Rattus rattus Rr 1 ------1 - - - - - Rr 2 - - - 2 1 1 - 1 - - - 2 1 1 1 - 1 - -

105 Appendix 4.5.2 continued North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI Rattus tanezumi Rt 1 1 ------1 ------Rt 2 - - 2 1 1 - - - 1 - - 2 - 10 - - - 1 - Rattus sp. Rattus 1 - - 1 1 - - - - 2 4 - 1 - - - - - 17 - Rattus 2 ------1 ------Rattus 3 - - 2 - - - - 1 1 - 5 - - - - - 1 1 1 Rattus 4 - - 1 ------1 1 - - - - - 5 - Rattus 5 ------1 - - - Sundamys muelleri Sm 1 - 2 ------2 - - - 1 - - - Sm 2 ------1 ------Sm 3 - - - - 1 ------6 - Sm 4 - - - 1 ------1 ------Sm 5 ------1 - - - - Sm 6 ------1 Sm 7 - - - - 1 - - - - 1 ------Sm 8 ------1 - Sm 9 - 1 1 - - 3 1 1 1 - 2 - - - - - 1 - 1 Sm 10 ------1 ------Sm 11 ------1 ------1 - Sm 12 1 ------Sm 13 ------2 - - 1 ------Sm 14 - - - - 1 - - 1 ------1 - - Sm 15 1 ------Sm 16 - - - 1 ------Sm 17 1 - - - 1 - 1 - - - - 6 - - - 1 - - 1

106 Appendix 4.5.2 continued North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI SCIURIDAE Callosciurus notatus Cn 1 - - - 1 ------1 ------Cn 2 - - - 1 - - - - 1 ------Cn 3 ------1 ------Cn 4 - - - - 1 1 - 1 ------Cn 5 ------1 ------Cn 6 ------1 ------Cn 7 ------2 ------Cn 8 ------1 Cn 9 ------1 - - Cn 10 ------1 - - - - Cn 11 ------1 Cn 12 ------2 - Cn 13 1 - 1 ------Cn 14 - 1 ------Cn 15 ------1 ------Cn 16 ------1 1 - - - Callosciurus prevostii Cp 1 ------1 - - - - Cp 2 - - - 2 ------Cp 3 ------1 ------Cp 4 ------1 ------Cp 5 - - - 1 ------Sundasciurus hippurus Sh 1 - - - - - 1 ------Sh 2 ------1 ------Sh 3 ------1 ------Sh 4 ------1 ------Sh 5 ------1 ------

107 Appendix 4.5.2 continued North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI Sundasciurus lowii Sl 1 - 3 ------Sl 2 ------2 ------Sl 3 ------3 1 ------Sl 4 ------1 ------Sl 5 ------1 - - Sl 6 ------1 - - Sl 7 ------2 Sl 8 ------1 - - - Sl 9 ------1 - - - Sl 10 ------2 - - - - Sl 11 ------1 ------Sl 12 ------1 ------PTILOCERCIDAE Ptilocercus lowii Pl 1 ------1 - - - Pl 2 ------1 - - - - TUPAIIDAE Tupaia gracilis Tg 1 ------1 ------Tg 2 ------2 1 ------Tg 3 ------2 1 ------Tupaia longipes see Appendix 3.6.5 Tupaia minor Tm 1 - - - - 1 ------Tm 2 - - - - 1 ------Tm 3 ------1 ------Tm 4 - - - - 1 ------

108 Appendix 4.5.2 continued North South

Haplotype NA1 NA2 NB1 NB2 NCa 1 NCa 2 NCa 3 NCb ND1 ND2 ND3 SE1 SE2 SP SF SG1 SG2 SH SI Tupaia tana Tt 1 1 1 1 ------Tt 2 - - - 1 ------Tt 3 - - - - 1 ------Tt 4 1 ------Tt 5 ------1 1 ------Tt 6 ------1 1 ------Tt 7 ------2 - - - Tt 8 ------1 ------Tt 9 ------1 - - Tt 10 ------2 - 1 - - - 1 Tt 11 ------1 - Tt 12 ------1 - - - - ERINACEIDAE Echinosorex gymnura Eg 1 - - - 1 ------Eg 2 ------1 ------Eg 3 - - - 1 - - - - - 1 1 1 - - - 2 - - - Eg 4 - - - 1 ------Eg 5 ------1 Eg 6 - - - 1 - - - - 1 - - - 1 - - 1 - - 1

109 Appendix 4.5.3 An introduction to the small mammals of the Lower Kinabatangan River In the following, the number of sampled individuals together with their spatial distribution along the river (red dots in map), phenotypic characteristics (taken from trapped individuals of this study as described in Chapter 2; averages of measurements are given for n > 5: median (range), n > 10: mean ± SD), details about the ecology (as described in the literature), the IUCN Red List classification (IUCN 2020), and implications drawn from this study are described for all small mammal species sampled along the Lower Kinabatangan River, during this study. All figures were taken or drawn by J. Brunke.

RODENTIA (Muridae) Chiropodomys sp (Pencil-tailed tree mice) Number of individuals and their distribution: 2, scarce, found only in two sites Phenotypic characteristics: Weight (W): 45 g, Head-Body length (HB): 116 mm, Tail length (T): 120 mm (1 specimen) Small mouse-like rodent with grey or grey brown upperparts. The underparts are cream white, the tail brown with a brush of hair at the tip. Taxonomy: Appropriate (cytochrome b) reference sequences for phylogenetic analyses were available from GenBank solely for the species C. gliroides. An unambiguous classification on species level was thus not possible based on morphometric characteristics and/or phylogenetic analyses but allowed a categorisation of the sampled individual as Chiropodomys sp. (Chapter 2). Four Chiropodomys species are known on Borneo (Chiropodomys major, C. muroides, C. pussillus, C. gliroides), of which three are rare endemics (C. major, C. muroides, C. pussillus). Their ecology is largely unknown, but all are nocturnal, arboreal and specialised seed predators (Phillipps and Phillipps 2016). A taxonomic affiliation to the following Chiropodomys species may be possible: C. major (Large pencil-tailed tree mouse) – Least Concern, unknown population W: 40 g, HB: 104 mm, T: 127 mm Generally scarce, but most common tree mouse on Borneo. A nomadic and canopy dwelling species, which is found in lowland and montane forests (Musser 1979; Wells et al. 2004; Phillipps and Phillipps 2016). Implications from this study: Due to the small sample size, effects of landscape features such as the Kinabatangan River or other fragmenting factors could not be assessed for this species. The small sample size might suggest an overall low prevalence of this species within Kinabatangan habitats.

110 Maxomys surifer (Red spiny rat) Least Concern, decreasing population Number of individuals and their distribution: 27, present in few forest sites on the southern riverbank Phenotypic characteristics: W: 135 ± 34 g, HB: 196 ± 17 mm, T: 196 ± 23 mm Medium sized rat, upperparts orange brown with short stiff spines and a distinct transition to the white underparts, colouration of upperparts extends around the neck (“collar”) and inner sides of the thighs. The tail is bicoloured with a dark grey colouration on top and lighter underneath, the skin of the tail easily gets discarded, possibly as an escape mechanism (if hold on). Ecology: Nocturnal and predominantly terrestrial rat that feeds on roots, fallen fruit, insects, and small vertebrates. Common in primary and secondary forests, adjacent gardens, grassland and rice fields, and in not too heavily disturbed areas (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: This species is widespread throughout Southeast Asia, Sumatra, Java and adjacent islands; on Borneo it is recorded from localities in lowlands and hills (Payne and Francis 2007; Phillipps and Phillipps 2016). Threats: It is a widespread species and currently not at risk, but population is declining in some areas due to deforestation (Francis 2008; IUCN 2020). Implications from this study: The distribution of this species on one riverside only identified the Kinabatangan River as effective dispersal barrier. Although, found in oil palm close to forest edge in other studies (Bernard et al. 2009), it remains unclear to what extent this species is able to tolerate or use oil palm plantations as habitat, and to what extent further forest reduction restricts genetic exchange and may cause genetic isolation in the already scattered Kinabatangan population. More studies are needed to evaluate this assumption.

111 Maxomys whiteheadi (Whitehead’s Maxomys) Vulnerable, decreasing population Number of individuals and their distribution: 206, present in all forest sites Phenotypic characteristics: W: 44 ± 10 g, HB: 125 ± 16 mm, T: 104 ± 8 mm The dark brown upperparts of this small rat have a reddish tinge and stiff spines, the underparts are cream to grey sometimes witha white spot on the belly. The tail is bicoloured with a dark grey colouration on top and lighter underneath, the skin of the tail easily gets discarded, possibly as an escape mechanism (if hold on). Ecology: This species is nocturnal and mainly terrestrial. The diet includes ants and other insects as well as plant matter. It occurs in tall and secondary forests, and was also reported from rice fields, plantations and disturbed areas, but only when adjacent to forest (Payne and Francis 2007; Francis 2008; Bernard et al. 2009; Phillipps and Phillipps 2016). Geographic distribution: Widespread throughout Peninsular Thailand, Malaysia, Sumatra and adjacent islands; and recorded throughout Borneo. On Borneo found in lowlands, hills and mountains (Payne and Francis 2007; Francis 2008; Achmadi et al. 2013; Phillipps and Phillipps 2016). Threats: Throughout its range M. whiteheadi is the most common Maxomys species but at risk with a population decline of more than 30% over the last 10 years because of loss of lowland forest in many areas (Francis 2008; Phillipps and Phillipps 2016; IUCN 2020). Implications from this study: Most abundant murid in Kinabatangan forests. Unimpeded dispersal maintains gene flow between sub-populations across and along the Kinabatangan River. While it could be verified in oil palm plantations in the past (Bernard et al. 2009; Mohd-Azlan et al. 2019; Chapman et al. 2019), it was absent from plantations in the present study. Nevertheless, was genetic connectivity slightly reduced by oil palm plantations, Possibly, M. whiteheadi forages mainly in oil palm from forest edges, although it was present in low numbers in plantation interiors elsewhere (Chapmann et al. 2019). To what extent this species is able to utilise oil palm plantations remains thus unclear. However, the active use of the river for migration may reduce fragmentation effects of barriers such as oil palm plantations along the river in this species.

112 Niviventer cremoriventer (Dark-tailed tree rat) Least Concern, unknown population Number of individuals and their distribution: 148, present in most forests of both riversides, but absent in smaller forests (< 2 ha). Phenotypic characteristics: W: 58 ± 14 g, HB: 139 ± 17 mm, T: 175 ± 19 mm A small rat with orange-brown upperparts containing long guard hairs and stiff spines. The underparts are cream-white to creamy orange, sometimes with orange spot on chest. The long tail is dark brown or grey brown with hairs at the tip. With long facial whiskers.

Ecology: Nocturnal and arboreal species that is predominantly active in small trees, lianas and shrubs to at least 5 m above the ground, and occasionally approaches the ground. It feeds on plant matter including fruits and seeds, but also on insects. N. cremoriventer is a forest-dependent species that prefers primary forest habitats but was found in secondary forests, forest edge and lightly wooded areas but never in oil palm plantations (Nakagawa et al. 2006; Wells et al 2007; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016; IUCN 2020). Geographic distribution: Widespread in Myanmar, Thailand, Malaysia, Sumatra, Java and adjacent islands. Found throughout Borneo in lowlands and hills (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not immediately at risk, but reduction of lowland forests may contribute to a substantial population decrease (Francis 2008; IUCN 2020). Implications from this study: The absence of this species in smaller forests suggests the need of large and structurally diverse habitats. Across the river genetic connectivity existed between sub- populations of both riversides, while gene flow along the river seems to be restricted. As this species depends on forest habitats, it is possible that oil palm represents an effective barrier to dispersal. However, the small sample size prevented an unequivocal identification of specific landscape barriers to dispersal, and further studies are needed to infer to what extent this species tolerates habitat modifications and degradation.

113 Rattus sp. (unknown rat) Number of individuals and their distribution: 48, present in most of the forests of both riversides, but absent in plantation Phenotypic characteristics: W: 79 ± 20 g, HB: 162 ± 12 mm, T: 137 ± 15 mm Medium sized-rat, with dark grey-brown upperparts, has long and dark-tipped guard hairs. The underparts are creamy grey, and the tail is dark grey-brown. Unlike any of the other sampled Rattus species, males of this species have very large testes.

Ecology: Unknown. Threats: Unknown. Taxonomy: Phenotypic and phylogenetic investigations allowed no taxonomic clarification for this cryptic species due to a lack of suitable reference sequences. However, its monophyletic position within phylogenetic trees suggests a distinct lineage but also membership in the Rattus genus. More detailed investigations on phenotypic, ecological and genetic characteristics are necessary to clarify the taxonomy of these Rattus sp. captured in the Kinabatangan area. An assignment to one of the following Rattus species known to occur on Borneo cannot be excluded: R. argentiventer (Ricefield rat) – Least Concern, stable population W: 85 – 240 g, HB: 140 – 210 mm, T: 130 – 205 mm A nocturnal rat which is most active on the ground, but climbs small trees. The diet includes plant matter (i.e. rice plants, grain, oil palm fruits) and some insects. Occurs predominantly in open areas like rice fields, plantations or grasslands. Invasive on Borneo, where it hasa scattered distribution (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). R. tiomanicus (Tioman rat) – Least Concern, increasing population W: 55 – 150 g, HB: 140 – 190 mm, T: 150 – 200 mm A nocturnal and terrestrial rat that also climbs small trees or bushes. The diet includes a wide range of plant and animal matter. Occurs in secondary and coastal forests, gardens, plantations, and in scrub and grasslands. Largely absent in dipterocarp forest. In some areas the most common rat in secondary forests and oil palm plantations, where it is often considered a pest. Widespread throughout the lowlands and hills on Borneo (Payne and Francis 2007; Francis 2008, Phillipps and Phillipps 2016).

Implications from this study: The absence of this species from the plantation site suggests an adaptation to forest habitats. Moreover, haplotype genealogies indicate a longer residence than for any of the other Rattus spp. verified in the Kinabatangan area. While the river did not act as a dispersal barrier for this species, small sample sizes prevent to investigate impacts of further

114 landscape structures. A possible dependency on forest habitats propose oil palm plantations as effective dispersal barrier and a susceptibility to fragmentation effects, but further investigations are necessary to clarify the taxonomic assignment of this species and its requirements in more details and to infer consequences of habitat disturbances.

Rattus exulans (Polynesian rat, Parcific rat) Least Concern, stable population Number of individuals and their distribution: 8, present in plantation and few forests of both riversides Phenotypic characteristics: W: 48 g (35 – 55 g), HB: 135 mm (110 – 149 mm), T: 126 mm (116 – 132 mm) Small rat with grizzled grey-brown and spiny upperparts. The underparts are grey-white. The tail is dark grey-brown.

Ecology: The smallest of the Rattus species is nocturnal and mainly terrestrial but climbs well. The omnivorous diet includes plant and animal matter. A commensal rat found in houses, gardens, plantations, rice fields, and secondary growth. In some areas it is considered as pest (Williams 1973; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016; GISD 2020a). Geographic distribution: Widespread throughout Southeast Asia, New Guinea, and the Pacific Islands. Found throughout Borneo. Threat: Not at risk (Francis 2008; IUCN 2020) Implications from this study: Probably introduced by human activity to the Kinabatangan area, but so far found only in low abundances within forest habitats. Heterospecific resource competition with other (larger) Rattus spp. and S. muelleri is likely and might restrict its expansion into forests. Nevertheless, population sizes of this species should be monitored and a further expansion of this invasive species into Kinabatangan habitats should be prevented.

115 Rattus rattus (Black rats) Spreading over six continents black rats (Rattus rattus) are globally the most widely distributed of all commensal animal species (Aplin et al 2011; Lack et al. 2012). The taxonomy of Rattus rattus is complex and not fully resolved. Initially, chromosomal variations suggested the existence of two distinct karyotypes of Rattus rattus, an Oceanic form (R. rattus, 2n = 38 – 40) and an Asian form (R. tanezumi, 2n = 42; Yosida 1981). Later phylogenetic analyses of mitochondrial DNA, however, revealed that Rattus rattus represents a species complex of at least six distinct species (Pagès et al 2010; Aplin et al 2011).

In the present study, phylogenetic analyses identified R. tanezumi in addition to R. rattus to occur in the habitats along the Lower Kinabatangan River (Chapter 2).

Rattus rattus (Black rat, Oceanic form) Least Concern, stable population Number of individuals and their distribution: 12, present in plantation and forests of both riversides Phenotypic characteristics: W: 109 ± 33 g, HB: 167 ± 26 mm, T: 169 ± 20 mm A medium-sized rat with dark grey upperparts and long guard hairs, the underparts are greyish white and the tail is dark grey-brown.

Ecology: This invasive species is nocturnal and sometimes diurnal. It is mainly terrestrial but climbs well. The diet includes a wide range of plant and animal matter. Commonly found close to human settlements and gardens near settlements, also found in a variety of natural and semi-natural habitats such as rice fields and oil palm plantations. It is often considered as a pest (Aplin et al. 2003, Payne and Francis 2007; Francis 2008; Wells et al. 2014; Phillipps and Phillipps 2016). Geographic distribution: Originally an Indomalayan species, introduced worldwide including Borneo. On Borneo initially thought to occur only in housing areas, but has repeatedly been found in primary and secondary forests (Wells et al. 2006, 2014; Payne and Francis 2007; Francis 2008; Aplin et al. 2011; Phillipps and Phillipps 2016; Loveridge et al. 2016). Threats: Not at risk (Francis 2008; IUCN 2020) Implications from this study: Probably introduced by human activity to the Kinabatangan area, but found so far only in low abundances within forest habitats. Heterospecific resource competition with other Rattus spp. and/or S. muelleri seems likely which might restrict its expansion into forests. Nevertheless, further invasion of this species into Kinabatangan habitats should be monitored and prevented.

116 Rattus tanezumi (Black rat, Asian form) Least Concern, increasing population Number of individuals and their distribution: 20, most abundant Rattus species in the plantation site and captured in various forest fragments along both riverbanks Phenotypic characteristics: W: 117 ± 47 g, HB: 167 ± 27 mm, T: 175 ± 26 mm A medium-sized rat with grizzled grey-brown upperparts and long guard hairs. The underparts are creamy grey, and the tail is dark grey-brown.

Ecology: Similar to R. rattus Geographic distribution: The origin of this rat is still not fully understood, but a native distribution from Eastern Asia is considered, from where it was possibly introduced to Japan, Southeast Asia, including the Sundashelf islands, the Philippines, New Guinea, South Africa and the United States (Aplin et al. 2011; Bastos et al. 2011; Lack et al. 2012; IUCN 2020). Threats: Not at risk (IUCN 2020) Implications from this study: The high prevalence of this species in the plantation suggests an expansion from human settlements into Kinabatangan forests. Although in low abundance, it occurs in almost all forest sites. Signals of unimpeded dispersal across and along the river indicates a high invasive potential of this species. Therefore, the population sizes of this invasive species should be monitored and a further expansion of this species within Kinabatangan habitats should be prevented in order to protect native small mammal assemblages.

117 Sundamys muelleri (Müller’s rat) Least Concern, decreasing population Number of individuals and their distribution: 186, present in all forest sites Phenotypic characteristics: W: 211 ± 45 g, HB: 216 ± 21 mm, T: 250 ± 21 mm The upperparts of this large rat are grizzled grey-brown with long dark guard hairs. The underparts are creamy white, the tail dark grey-brown. Ecology: A nocturnal rat which is mainly terrestrial but can climb. It feeds on various plant and animal matter primarily in moist habitats, along streams and riverbanks. It occurs in lowland primary and secondary forests, found at forest edges and lightly wooded areas near forests. Also present in anthropogenic vegetation such as fallow grassland and tree stands or housing areas with nearby vegetation (Lim 1979; Payne and Francis 2007; Francis 2008; Wells et al. 2014; Phillipps and Phillipps 2016). Geographic distribution: Widespread throughout Myanmar, Thailand, Malaysia, Sumatra, Palawan and adjacent islands. Also known from lowland and hills throughout Borneo (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not immediately at risk, but populations are declining considerably due to loss of primary and secondary forests throughout its range. Hunting further contributes to population declines (Francis 2008; IUCN 2020). Implications from this study: Highly abundant species in forests along the Lower Kinabatangan River. High abundances of S. muelleri have been suggested to negatively impact other species with overlapping ecology (Charles and Ang 2010), although this assumption needs to be investigated for the Kinabatangan region. The Kinabatangan River represents no effective barrier to dispersal for this species, but genetic patterns suggest constraints for dispersal through large oil palm plantations (Chapter 2). Active crossing of the river may reduce fragmentation effects of other barriers such as oil palm plantations. Being a generalist, S. muelleri might be able to utilise oil palm plantations that would explain its occurrence in plantations elsewhere (Bernard et al. 2009; Mohd-Azlan et al. 2019; Chapman et al. 2019). It was, however, absent from the plantation site in the present study that was 1 km away from forest edges. Therefore, S. muelleri may forage near forest edges, but may avoid further penetration into oil palm.

118 RODENTIA (Sciuridae) Callosciurus notatus (Plantain squirrel) Least Concern, increasing population Number of individuals and their distribution: 232, present in all forest sites Phenotypic characteristics: W: 207 ± 25 g, HB: 217 ± 30 mm, T: 193 ± 21 mm Medium-sized squirrel with brown upperparts and lateral white and black stripes. The underparts are orange, the tail bushy and grizzled brown. Ecology: This diurnal squirrel is most active in the early morning and late afternoon. Being arboreal it travels and feeds mainly in small trees. The omnivorous diet includes a wide variety of nuts, fruits, bark, leaves, and insects (mostly ants). It is the most abundant squirrel in gardens, plantations and secondary forests, and can live and breed entirely in monoculture plantations. In cocoa and oil palm plantation it is considered a serious pest. Less frequently found in the interior of undisturbed and tall forests but common in coastal and swamp forest (Hafidzi 1998; Abdullah et al. 2001; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Peninsular Thailand, Malaysia, Sumatra, Java and most intervening islands. On Borneo widespread in the lowlands and hills (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not at risk (Francis 2008; IUCN 2020). Implications from this study: The most abundant small mammal found along the Kinabatangan River, possibly because of its high ecological flexibility. Although C. notatus was not verified in oil palm in the present study, it was observed in plantations in other studies (Hafidzi 1998). An ability to utilise oil palm would represent another explanation for its high abundance in the Kinabatangan area. The high abundance of the generalist C. notatus within forests may limit populations of more specialised species. Genetic results suggest restrictions in dispersal across and also along the riversides. However, effects of other landscape characteristics could not be assessed in this study due to a small analysed sample size. More studies are clearly necessary to assess factors shaping the population genetic structures in this species.

119 Callosciurus prevostii (Prevost’s squirrel) Least Concern, decreasing population Number of individuals and their distribution: 6, scarce, present only in few forests Phenotypic characteristics: W: 373 g (315 – 405 g), HB: 275 mm (225 – 299 mm), T: 257 mm (230 – 269 mm) The upperparts of this large squirrel are entirely black, sometimes with faint lateral white stripe. The underparts are bright orange. The bushy tail is black. Various colour morphs exist throughout Borneo (Phillipps and Phillipps 2016; Lurz et al. 2017). Ecology: A diurnal squirrel, most active in the early morning and late afternoon. Usually arboreal on tree trunks and foliage of mid to high canopy. Descends to the ground only to cross gaps in the tree canopy. The diet includes fruits (especially those with a sweet or oily flesh), young leaves, bark and insects (i.e., ants, termites, and beetle larvae). A solitary and territorial species that is found in tall and secondary forest but enters gardens and plantations from adjacent forest to feed on fruits. Absent in open habitats and plantations occupied by plantain squirrels (Payne and Francis 2007; Francis 2008, Phillipps and Phillipps 2016; Lurz et al. 2017). Geographic distribution: Found in Peninsular Thailand, Malaysia, Sumatra, Sulawesi and smaller Indonesian islands. One of the most common forest squirrels on Borneo, where it is widespread in the lowlands and hills, but no records from South Kalimantan and lowlands of Central Kalimantan (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016; Lurz et al. 2017). Threats: Not currently at risk, but population is declining due to habitat loss and hunting (Francis 2008; IUCN 2020). Implications from this study: Although sampled in low numbers, this species could frequently be observed in the forests along the Kinabatangan, suggesting a higher prevalence than obtained by capturing in this study. Signals of restricted gene flow across and along the Kinabatangan. However, impacts on population structure by oil palm and other landscape characteristics could not be assessed in this study due to small sample size. Possibly, further forest reductions restrict genetic exchange and cause genetic isolation in the scattered Kinabatangan populations, but this needs further investigations in the future.

120 Sundasciurus hippurus (Horse-tailed squirrel) Near Threatened, decreasing population Number of individuals and their distribution: 5, scarce, present only in three forest sites on the northern riverbank Phenotypic characteristics: W: 342 g (280 – 360 g), HB: 253 mm (249 – 265 mm), T: 230mm (195 – 262 mm) A large squirrel with reddish brown colouration of the upperparts, shoulders and head are distinctively grey. The underparts are white, and the bushy tail is grizzled grey. Ecology: This diurnal and arboreal squirrel moves frequently between canopy of mid storey trees and the ground to retrieve nuts from canopy and to store them at ground level. Specialised on hard nuts of sub- canopy trees, but diet includes also seeds, earthworms and insects. It has large home ranges and is nomadic within them. Inhabits primary and old secondary forests mainly in lowlands or lower hills (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Distributed throughout Peninsular Thailand, Malaysia, and Sumatra. Rare squirrel on Borneo with sometimes patchy distribution throughout its range, but locally common in lowlands of Sabah and Sarawak (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Threatened by habitat loss throughout its range (Francis 2008; IUCN 2020). Implications from this study: Due to the small sample size, effects of landscape characteristics on population structures could not be assessed for this species. However, the small sample size might suggest an overall low abundance and rarity of this species within Kinabatangan habitats. Its specialised diet might prevent it from using monocultures such as oil palm. Furthermore, further encroachment of forests in the Kinabatangan area may decrease the availability of necessary resources and thus represents a threat to this rather specialised species.

121 Sundasciurus lowii (Low’s squirrel) Least Concern, stable population Number of individuals and their distribution: 68, higher abundance in forests on southern riverside, in the north found only in few forest sites Phenotypic characteristics: W: 91 ± 15 g, HB: 158 ± 16 mm, T: 91 ± 19 mm A small squirrel with dark brown upperparts and a creamy ring around the eyes. The underparts are creamy white, the bushy tail is rather short and grizzled brown. Ecology: This squirrel is diurnal and most active in the early morning and late afternoon. It is scansorial, foraging from the ground to the sub-canopy (< 5 m), mainly travels and feeds in small standing trees, in fallen trees and on the ground. Feeds mainly on surface bark, but the diet also includes fruits, insects and fungi. A common squirrel of primary and secondary habitats including shrubs near forests (Whitten and Whitten 1987; Abdullah et al. 2001; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Found in Peninsular Thailand and Malaysia, Sumatra and smaller Indonesian islands. On Borneo it is recorded throughout the lowlands and lower hills (Payne and Francis 2007, Francis 2008, Phillipps and Phillipps 2016). Threats: Not at risk (Francis 2008, IUCN 2020). Implications from this study: This species showed a strong genetic fragmentation across and along both riverbanks. However, besides the Kinabatangan River no other effective gene flow barrier could be identified in this study due to small genetic sample sizes. It is possible that resource distribution or certain space use and ranging patterns impacted population genetic patterns in this species. Overall low migratory propensity, and further habitat fragmentation may isolate populations. However, a larger sample size is needed to evaluate this hypothesis.

122 Further squirrels sighted but not captured in forests along the Kinabatangan River:

Exilisciurus exilis (Bornean Pygmy Squirrel) Data Deficient, unknown population Number of individuals and their distribution: unknown Phenotypic characteristics: W: 12 – 26 g, HB: 62 – 82 mm, T: 42 – 62 mm Very small squirrel with olive-brown upperparts and buff underparts, the tail is grizzled brown. Ecology: The ecology of this squirrel is largely unknown. It is diurnal and active mainly on tree trunks, on which it ranges in all strata up to the canopy to feed on bark and ants. It occurs in tall and logged dipterocarp forests (Payne and Francis 2007; Phillipps and Phillipps 2016). Geographic distribution: Endemic to Borneo, very common in the lowlands and hills (Payne and Francis 2007; Phillipps and Phillipps 2016) Threats: Threatened by the substantial forest loss on Borneo (IUCN 2020).

Ratufa affinis (Giant squirrel) Near Threatened, decreasing population Number of individuals and their distribution: unknown Phenotypic characteristics: W: 875 – 1500 g, HB: 320 – 380 mm, T: 370 – 444 mm Largest Bornean squirrel, the colouration is variable but usually with darker upper- and pale underparts, has a long tail. Ecology: A diurnal squirrel, which is mainly active in tall trees. It descends to the ground only to cross gaps in the canopy. The diet consists of seeds, figs, leaves and bark. Found in tall dipterocarp and lower montane forests, and occurs also in secondary forests, but rarely enters plantations (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Distributed in Peninsular Thailand and Malaysia, Sumatra and smaller Indonesian islands. On Borneo it is widespread throughout lowlands and hills (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Threatened because widespread habitat loss and hunting through much of its range contribute to considerable population declines. It is sensitive even to selective logging (Francis 2008; IUCN 2020).

123 SCANDENTIA (Ptilocercidae) Ptilocercus lowii (Pen-tailed tree shrew) Least Concern, decreasing population, listed on CITES Appendix II Number of individuals and their distribution: 2, scarce, found only in two forest sites on the southern riverbank Phenotypic characteristics: W: 45 g, HB: 142 – 144 mm, T: 178 – 195 mm The outer appearance of the Pen-tailed tree shrew is unlike that of all other tree shrews, with grey-brown upperparts and cream underparts, it is rather mouse-like. The long grey-brown tail is naked with a feather-like flag of black to white hairs towards the tip.

Ecology: The only nocturnal tree shrew. Unlike most nocturnal mammals, it has bright silver rather than orange eye shine in reflecting light. It is usually arboreal, but sometimes travels on ground. Often travelling on vines, and forages on vertical tree trunks from the base to the canopy. The diet consists mainly of insects and other arthropods, but it was also observed to feed on alcoholic palm nectar. Occurs in forests but was also recorded in gardens and disturbed habitats including secondary forest and plantations (Emmons 2000, Wiens 2008; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: The distribution spreads throughout Peninsular Thailand and Malaysia, Sumatra and adjacent islands. It is scarce on Borneo with a relict distribution in the lowlands and hills (Emmons 2000, Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not currently at risk but undergoes substantial population declines due to forest loss throughout its range (Francis 2008; IUCN 2020). Implications from this study: The small sample size prevented to assess effects of landscape features such as the Kinabatangan River and habitat fragmentation on population structure. However, this species could frequently be observed in further forests along the Kinabatangan, suggesting a higher abundance than that obtained by this study. Possibly, its rather insectivorous diet prevents high trappability with fruit baits. To what degree forest fragmentation may restrict genetic exchange and cause genetic isolation between demes, needs to be investigated in future studies.

124 SCANDENTIA (Tupaiidae) Tupaia gracilis (Slender tree shrew) Least Concern, decreasing population, listed on CITES Appendix II Number of individuals and their distribution: 7, scarce, on both riversides present in few forest sites Phenotypic characteristics: W: 60 g (45 – 75 g), HB: 151 mm (138 – 174 mm), T: 176 mm (117 – 190 mm) Small and slender tree shrew with olive-brown upperparts and white shoulder stripes. Underparts are cream-white, the rather slender tail is grey-brown. Ecology: This scarce tree shrew is diurnal. Displaying a rather scansorial activity, it climbs up to 1.5 m above ground. The diet includes insects and fruits. It has very large home ranges. A resident of primary and logged forests (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Confined to Borneo and adjacent small islands. Recorded from the lowlands and lower hills, but absent in south-eastern Kalimantan. In Sabah, the least common tree shrew (Emmons 2000; Wells et al. 2007; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not currently at risk but substantial loss of forests on Borneo represent a major threat to the declining population (IUCN 2020). Implications from this study: While the Kinabatangan River seems to represent an effective dispersal barrier for this species, impacts of other landscape features could not be assessed for this species due to a small sample size. However, the small number of captured individuals might suggest an overall low abundance within Kinabatangan habitats of this overall rare species, and further forest loss is likely to restrict genetic exchange and cause genetic isolation further in the scattered Kinabatangan populations.

125 Tupaia longipes (Bornean tree shrew) Least Concern, decreasing population, listed on CITES Appendix II Number of individuals and their distribution: 126, present in most forests of both riverbanks Phenotypic characteristics: W: 180 ± 28 g, HB: 219 ± 24 mm, T: 197 ± 23 mm The upperparts of this medium-sized tree shrew are dark brown with orange shoulder stripes, the underparts are creamy-orange, and the slender tail has a uniform brown colouration. Ecology: This diurnal tree shrew has a rather terrestrial activity pattern. It is most often seen around fallen trees and branches, in low woody vegetation, and rarely off the ground. The diet consists mainly of insects (ants, termites) and other arthropods, but it feeds also on sweet and oily fruits. It occurs in virgin and logged forests as well as in gardens and plantations with dense understory but is absent from habitats with sparse understory structure (Chapter 3, Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). This very “nervous” and “agile” tree shrew has large territories with sizes up to 16 ha (Chapter 3; Figure left bottom), which might prevent males to monopolise more than one (sometimes two) females. Both sexes display high territoriality toward the same sex (Emmons 2000). Large territories and intense intrasexual competition may induce dispersal in both sexes alike. Geographic distribution: Endemic to Borneo, common throughout the area Male (M-Tl: 15.6 ha) and female (F-Tl: 13.4 ha) (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps home ranges 2016). Threats: Not currently at risk but deforestation impacts populations especially in the lowlands (IUCN 2020). Implications from this study: Genetic exchange between the two sub-populations existing north and south of the Kinabatangan is restricted by the river, but infrequent accidental crossings maintain some degree of gene flow between riversides. Recent habitat fragmentation along the southern riverbank has already reduced gene flow, while habitat connectivity along the northern riverbank maintains genetic exchange between populations. This species is largely absent in habitats with poor understory structures, such as oil palm plantations. Therefore, further expansion of oil palm plantations within the Kinabatangan area is likely to isolate the fragmented T. longipes population even further.

126 Tupaia minor (Lesser tree shrew) Least Concern, decreasing population, listed on CITES Appendix II Number of individuals and their distribution: 4, scarce, found only in few sites on the northern riverside Phenotypic characteristics: W: 40 – 70 g, HB: 105 – 138 mm, T: 145 – 154 mm Smallest of all Bornean tree shrews with olive-brown upperparts and white shoulder stripes. The underparts are cream-white, the tail grey-brown. Ecology: Besides P. lowii, this is the most arboreal of the Bornean tree shrews, travelling along lianas or leaf bunches in the mid canopy stratum, but occasionally approaches the ground. Most active during the day, it forages for insects (often in association with Yellow-bellied Bulbuls and Racket-tailed Drongos; Emmons 2000) and fruits. It is a common resident of lowland primary, logged (when vine covered) and peat-swamp forests (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Geographic distribution: Occurs in Peninsular Thailand, Malaysia, and on Sumatra. On Borneo it is found throughout the lowlands (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not currently at risk but deforestation contributes to population declines in some areas (Francis 2008; IUCN 2020). Implications from this study: The small sample size prevents to assess effects of landscape features such as the Kinabatangan River and other landscape characteristics on population structures. However, the presence of this species on one riverside only proposes the Kinabatangan River as effective dispersal barrier. The inability to utilise monocultures as habitat may restrict genetic exchange between demes isolated by oil palm and highlights the need for more comprehensive studies.

127 Tupaia tana (Large tree shrew) Least Concern, decreasing population, listed on CITES Appendix II Number of individuals and their distribution: 146, most abundant tree shrew species, present in all forest sites Phenotypic characteristics: W: 234 ± 36 g, HB: 228 ± 21 mm, T: 189 ± 12 mm The largest of all tree shrews. Has a very long muzzle. Body upperparts are reddish-brown with a distinct black stripe on the back originating from the head. Shoulders and head are grey-brown. The bushy tail is often brightly red to fiery orange (ventral). Ecology: This common tree shrew is diurnal and the most terrestrial one. Being an ecological generalist, it feeds on a wide variety of arthropods, earthworms, and some fruits. It is rarely found outside of lowland primary forests or dense, shaded areas in logged and secondary forests (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Compared to other tree shrews, it has relatively small home ranges (HR: 4 – 11 ha) that overlap between males and females (Brunke et al. unpublished data, Figure left-hand side). Although territories are mainly defended against the same sex in this pair-living species, territoriality is less strong than in T. longipes and extra-pair paternity has been shown to occur (Munshi-South 2007). A triangulation study in the forest of the LKWS revealed strong territory overlaps of

Male (M1-Tt: 5.3 ha; M2-Tt: females (Brunke et al. unpublished data, Figure left). It is known that 11.1 ha) and female (F1-Tt: females disperse further than males in this species, and females seem to 4.0 ha; F2-Tt: 4.8 ha) home perform active mate choice (Emmons 2000; Munshi-South 2007; Munshi- ranges South et al. 2007; Munshi-South 2008). Geographic distribution: Occurs on Sumatra and adjacent small islands, and on Borneo, where it is widely distributed, throughout the lowlands. In Sabah this is the most common tree shrew (Emmons 2000; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not currently at risk but deforestation contributes to population declines in some areas (IUCN 2020). Implications from this study: Sub-populations of this species showed high genetic differentiation between riverbanks. As stronger signals of restricted gene flow within the coherent northern sub- population were present, it is possible that resource distribution or socio-ecological factors shaped population genetic patterns in this species. Effects of other landscape characteristics could not be assessed for T. tana in this study due to a small sample size. Like T. longipes, this species is largely

128 absent from habitats with poor understory structures, and further expansion of oil palm plantations within the Kinabatangan area might easily isolate populations. Compared to studies conducted in large and continuous forests elsewhere in Borneo (Emmons 2000; Munshi-South 2008), no sex- biased dispersal could be observed in the Kinabatangan population. Habitat fragmentation may already have caused alterations in sex-specific dispersal patterns. However, a larger sample size is needs to evaluate this hypothesis.

ERINACEIDAE (Gymnures) Echinosorex gymnura (Moonrat) Least Concern, unknown population Number of individuals and their distribution: 16, present in forests of the northern and southern riverside Phenotypic characteristics: W: ~1000 g, HB: 401 ± 21 mm, T: 258 ± 16 mm Large rat-like animal. The whole body is white with some scattered black hairs. The tail is naked with pink colouration. Has an intense odour that is unlike that of any other animal.

Ecology: The moonrat is terrestrial and nocturnal and stays in burrows during daytime. Feeds on terrestrial and aquatic small animals, mostly arthropods, earthworms, crustaceans or frogs. On Borneo it occurs mainly in forests but enters gardens and plantations elswhere. Enters water and prefers muddy streams and damp areas in lowland primary and logged forest with high-quality habitats. It can tolerate a certain degree of habitat modification (Liat 1967; Gould 1978; Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016; Brozovic et al. 2018). Geographic distribution: Based on phenotypic characteristics two subspecies are recognized (E. gymnura gymnura and E. gymnura alba; Corbet 1988). On Borneo only E. gymnura alba is known from many sites in the lowlands (Payne and Francis 2007; Francis 2008; Phillipps and Phillipps 2016). Threats: Not currently at risk but habitat loss probably imposes threat to populations (Francis 2008; IUCN 2020). Implications from this study: Larger rivers such as the Kinabatangan represent no barrier to dispersal in E. gymnura, most likely due to its affinity to aquatic habitats for feeding. Whether other geographic landscape characteristics, such as forest fragmentation, influence or restrict dispersal between demes, could not be tested in the present study due to the small sample size. However, E. gymnura highly depends on dense, high-quality forests and seems to avoid areas close to oil palm plantations (Brozovic et al. 2018). An impact of oil palm plantations on dispersal is therefore possible,

129 but further genetic and behavioral studies are clearly necessary to infer to what extent this species is affected by oil palm and forest modifications.

Refernces

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131 132 Acknowledgements

During the long process of accomplishing this thesis I received great help, support, guidance and patient of many people and institutions. With this I would like to provide my thanks to all of them.

First and foremost I thank Prof. Dr. Ute Radespiel who provided me with the opportunity to work on this subject. I thank her for her supervision, which has guided me through all the phases of this work. She invested an enormous amount of time for me and her inspiring comments and discussions during that time were a great help.

Posthumously I thank Prof. Dr. Elke Zimmermann who has encouraged me to this work and has given me the opportunity to accomplish this thesis in the Institute of Zoology. She accompanied this work over the years and has been supportive.

I am very grateful to Dr. Benoit Goossens who gave me the opportunity to get in touch with the wonderful Bornean wildlife. For the incredible time I had in DG and for the great support provided to realise this project in many ways. Among others, to enable fieldwork and make the lab work at Cardiff University possible.

I thank Prof. Dr. Michael W. Bruford who admitted the possibility to work in his lab at Cardiff University. Who provided his help during lab work and analyses and whose expertise, suggestions and very constructive criticism contributed to the improvement of previous manuscripts.

I would also thank Isa-Rita Russo who introduced me to the world of lab work and genetics and willingly provided help during data analysis and manuscripts writing. For help provided during data analysis I also thank Pablo Orozco-terWengel. I also thank him for his interest and ideas in further projects resulting from this work.

Many people participated in the field and lab work and I am grateful to each of them. A special thanks to the staff and students of DG who provided help, comfort and laugh. Among them I thank Samsir, Doyo, Anto, Budin, Mike and Josie for making small mammal trapping and habitat analysis a great delight. I thank Jumrafiah Abd. Shukor for supporting this work as Malaysian counterpart.

The members of the Institute of Zoology I would like to thank for all the help and for the moments shared together. It was a pleasure to work with you all.

My last and deepest thanks go to my family and friends who helped me in so many ways during these long years. None of this would have been possible without their patient, support and care. They helped me overcome setbacks and stay focused on my study.

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