Habitat Creation for Animals by Teredinid Bivalves in Indonesian

Home , Ariadna bicolor, Bankia (bivalve), Bruguiera gymnorhiza, Dartfish, Isognomonidae, Mangrove rivulus

Habitat creation for animals by teredinid bivalves in Indonesian

mangrove ecosystems

A thesis submitted in partial fulfilment of the requirements for the award of the

degree of Doctor of Philosophy of the University of Portsmouth

Ian Wyndom Hendy

Institute of Marine Sciences, School of Biological Sciences

University of Portsmouth

PO4 9LY

October 2012

Abstract

A better understanding of the fundamental role large woody debris (LWD) plays within mangrove ecosystems may provide further insights into important ecological processes, such as wood degradation and biodiversity maintenance within mangrove forests.Though the volume of fallen wood in mangrove forests can be huge, little is known of the breakdown pathways and biodiversity maintenance of

LWD in mangrove ecosystems. The degree of mangrove fauna dependent upon

LWD and the need for such substratum in mangrove ecosystems may provide further insights in to the important role of woody biomass in these otherwise globally shrinking habitats due to forest harvesting.

The breakdown, recycling and flux of nutrients from LWD within mangrove forests is maintained by biodegrading organisms in areas from terrestrial to marine habitats. The tidal inundation sets limits on the wood degrading communities within the mangrove forests of Sulawesi. This study presents details of the environmental and biological association of biodegrading organisms within the forests in the

Wakatobi Marine Park (WMP), Sulawesi. Wood boring animals belonging to the family Teredinidae are the dominant biodegraders of LWD in the mid- to low- intertidal areas of the mangrove forests. Teredinid attack greatly reduces the volume of LWD in the mid- to low intertidal areas of the forests. Within the forests, emersion time was the greatest influence of the distribution of the biodegrading organisms spanning from the supra-tidal down to the low intertidal.

The response of Rhizophora stylosa prop-roots to physical damage and the activity of teredinids upon damaged prop-roots were investigated. With severe levels of root damage, the level of teredinid activity increases, resulting in root death and iii

detrital input. However, when the roots were exposed to a superficial and moderate level of damage, an over-compensation of tissue re-growth was observed.

LWD in the intertidal zone is often tunnelled by teredinids. The tunnels are blind-ending cylinders that taper to a small opening at the wood surface. However, larger openings appear when wood is heavily tunnelled and the surface is broken open. Teredinid death then leaves niches for cryptofauna. The greater the number of teredinid tunnels within LWD, the more diversity was found. Animals of particular interest were the dartfish, Parioglossus interruptus and the intertidal spider, Desis martensi found in the vacant teredinid tunnels.

Desid spiders were abundant within the LWD and dartfish collected from within teredinid-attacked LWD were smaller than dartfish populations not within

LWD. Desids and dartfish residing within the wood may benefit from the significantly lower temperatures within teredinid-attacked detritus compared to external air temperatures. Desis martensi has a life-history strategy centred on strong parental care, with lots of energy invested in to its young. Vulnerable stages of dartfish exploit the vacant teredinid tunnels. If it were not for the tunnels created by the teredinids the unusual behaviour adopted by dartfish and spiders would not be possible. Thus, many animals in mangrove forests of the WMP rely on LWD as a predation refuge enhanced by the teredinid tunnels within the LWD. A variety of different species were found inside teredinid attacked LWD, and the cryptic behaviour of the fauna ranged from breeding to predator avoidance. These findings indicate that in forests where wood is harvested, reduced availability of LWD will result in reduced biodiversity.

Acknowledgements

First and foremost, I wish express my sincere and total gratitude to Simon

Cragg, for seeing something in me which I believed I was not capable of. I thank

Simon for his tireless work ethic and professionalism. Thank you for your support, encouragement and guidance.

I am very grateful towards Operation Wallacea for providing research funding and providing the resources to make this project possible. I also want to give my sincerest thanks to Amat and Kungdan – my wonderful mangrove posse who worked with me with enthusiasm and passion, in long hard days when in the field. I miss those guys. In addition, I also wish to give an extended very large Ian bear-hug to several Hoga stalwarts – like Simon, you have been a role-model and inspiration to me, Richard Barnes and Chris Todd.

To the dynamic and cheerful IMS team – it has been a pleasure to have been a part of this institution for a number of years. The enthusiasm, passion and drive of each and every one of you are infectious. In the early years, the support network provided by Maureen and Ray and during the completion of my study Lucy, Graham,

Amaia, and Matt has been a constant source of encouragement, thank you.

I would also like to give a big thank you to Scott Armbruster, for his expert opinions and rapid feed-back with various drafts of this thesis. My gratitude and appreciation also goes out to David Jones (The Natural History Museum of London) for identification of termites. J. Lawrence for his identification of coleopteran larvae, and G. Jones and L. Ryvarden for their help towards identification of the v

basidiomycetes fruiting bodies. A big thank you goes out to Trevor Willis, for his expert advice on using the PRIMER statistical package.

Furthermore, I would like to thank my very hard working American colleagues who I have worked with on Hoga Island. Wayne, John and Theresa, each of you have been a pleasure to work with. I have learnt so much with other scientific disciplines – all of which is thanks to you guys. The seemingly hard endless hours you all put in to your lab-work was a lesson in itself.

A special mention must go to Peter Chivers – another role model of mine.

Although you are gone and missed, you have never been forgotten - your encouragement and belief in me during my A- levels and GCSEs when attending

Highbury College was such a boost. Also, very special thanks go out to Tricia

Farnham, my A- level biology teacher.

Combined with my collaborators and colleagues, to my family and friends – your support and constant belief in me throughout has been amazing and a real source of energy. Clare, thank you for putting up with me over the last year, you have been a ray of sunlight. Mum, Matt thank you both for supporting me all the way. To my boys at the gym, you must be the only bodybuilders who know what teredinid bivalves are! Thank you all, for being so understanding and patient.

Grandpa, although you are no longer with us – this PhD is dedicated to you, I miss you…

List of contents

Chapter 1

General introduction

1.1 Overview 1

1.1.1 Indo-Pacific mangrove systems 2

1.1.2 Mangrove biomass and partitioning 6

1.1.3 Detrital input and breakdown pathways 8

1.1.4 Role of fallen mangrove wood 10

1.2 The Teredinidae (Mollusca: Bivalvia) 13

1.3 Study site, the Wakatobi Marine Park 15

1.4 Research aims and objectives 18

Chapter 2

MANGROVE FORESTS OF THE WMP, EAST SULAWESI, INDONESIA

2.1 Introduction 20

2.2 Materials and methods

2.2.1 Belt transects, tree identification, and measurements of emersion time

and salinity 22

2.2.2 Mangrove forest descriptions 23

2.2.3 Tree densities and species basal area 29

2.3 Data analysis 31

2.4 Results 32 vii

2.4.1 Mangrove area and botanical distributions 32

2.4.2 Environmental factors 34

2.4.3 Canopy-forming trees and forest characteristics 36

2.4.4 Tree species dominance – basal areas, biomass and tree densities

2.4.5 Site similarity and tree species diversity 55

2.5 Discussion 58

Chapter 3

INVESTIGATION OF WOOD BIODEGRADATION IN MANGROVE

ECOSYSTEMS

3.1 Introduction 66

3.2 Materials and methods 71

3.2.1 Measurement of wood-loss in panels 71

3.2.2 The measurement and collection of dead wood and arbitrary scale classification and wood-loss from teredinid attack 72 3.2.3 Wood degrading organisms and environmental factors in the Galua and

Laulua mangrove forests 75

3.3 Data analysis 76

3.4 Results 77

3.4.1 Wood degradation within the landward, Sombano and seaward, Langira

forests 77

3.4.2 Sombano, Langira and Kal 1 panel analysis – wood loss 77

3.4.3 The biodegrading organisms found on LWD 78

3.4.4 The volume of LWD in mangrove forests 87

3.4.5 Wood-degrading organism distribution and environmental variables –

emersion, distance from the land and salinity 89 viii

3.4.6 The change between terrestrial and marine wood-degrading organisms

3.4.7 Teredinid panel-fragmentation 96

3.5 Discussion 97

Chapter 4

THE EFFECT OF TEREDINID TUNNELLING ON MANGROVE PROP-

ROOTS

4.1 Introduction 110

4.2 Materials and methods 115

4.2.1 Root scar experiment 115

4.2.2 Root biscuit analysis 116

4.3 Data analysis 118

4.4 Results 118

4.4.1 Root scar analysis upon live roots 118

4.4.2 Root biscuit analysis 123

4.4.3 Atypical growth of lignified tissues 125

4.5 Discussion 127

Chapter 5

HABITAT CREATION FOR ANIMALS BY TEREDINIDS

5.1 Introduction 132

5.2 Materials and methods 134 ix

5.2.1 Assessing the sediment macro-infauna, root epifauna and LWD cryptic

fauna 134

5.2.2 Assessing the effect of teredinid tunnelling on LWD in mangrove

forests 135

5.2.3 Recruitment and succession of animals to drilled wooden blocks

mimicking teredinid-attacked LWD 136

5.3 Data analysis 137

5.4 Results 138

5.4.1 Comparisons of animal diversity in different substrata 138

5.4.2 Biodiversity in teredinid tunnelled wood 142

5.4.3 Recruitment and succession of animals within drilled wooden blocks

mimicking teredinid tunnels 158

5.5 Discussion 166

Chapter 6

BIODIVERSITY MAINTENANCE BY VACANT TEREDINID TUNNELS

6.1 Introduction 172

6.2 Materials and methods 175

6.2.1 Field studies 175

6.2.2 Laboratory studies 177

6.3 Data analysis 180

6.3.1 Field studies 180

6.3.2 Laboratory studies 181

6.4 Results 182 x

6.4.1 Temperature analysis 182

6.4.2 Parioglossus interruptus autecology analysis 185

6.4.3 Desis martensi autecology analysis 190

6.5 Discussion 193

Chapter 7

GENERAL DISCUSSION 202

Reference list 215

Contribution of collaborators

Section Contributions of collaborators of work reported in this thesis Institution Texas A&M University – Corpus 3.1 Jason Williams prepared all arc G.I.S mangrove forest maps (Figure 3.1) Christi

6.2.2 John Eme performed all thermal tolerance experiments for Parioglossus interruptus University of North Texas

6.2.2 Wayne Bennett and Theresa Dabruzzi designed the wind tunnel used for the wood cooling experiment University of West Florida

xii0 xii Declaration

While registered as a candidate for a doctorate degree, I have not been registered for any other research reward. The results and conclusions embodied in this thesis are the work of the named candidate and have not been submitted for any other academic post-graduate degree.

Ian Wyndom Hendy

Chapter 1

GENERAL INTRODUCTION

1.1. Overview

Destruction and declines of ecologically important habitats continue; from the global reduction of rain-forests (Foley et al., 2007) to the vast coral-bleaching events of

Caribbean reefs (McWilliams et al., 2005; Donner et al., 2007). Such ecosystems are now reported as vulnerable and are a major cause of concern (Hughes, 1994; Laurence et al., 2002). Moreover, the animals that depend upon the resources within those habitats are also under threat (Laurence et al., 2002; Pandolfi et al., 2003). The knock-on effect of the continued destruction will also magnify problems for ecosystems adjacent to the impacted areas (Clausen and York, 2008). Eutrophication and increasing sedimentation rates in freshwater systems and coastal-marine environments have been linked to forest harvesting and farming (Nisbet, 2001; Brush, 2009). Thus, anthropogenic modifications within natural habitats have been negatively linked to many ecological issues – from declines of ecosystem resources to the reduction of ecosystem biodiversity. Indeed, those issues also relate to the intertidal forest systems of tropical coastlines – mangrove forests – which are also under threat (Duke et al., 2007).

This study is an investigation in to the ecology of mangrove forests of Sulawesi,

Indonesia. The study begins with detailed assessments and descriptions of different mangrove forests, defined by their geomorphological setting such as land-locked and forests connected with seagrass beds and coral reefs. Plant communities, their densities and basal areas were recorded within each forest. Woody biomass was also studied, by investigating the breakdown pathways of detritus from large trees to reveal how fallen 2 wood is recycled and processed in areas of each mangrove forest that have different biodegradation processes by organisms found in terrestrial to marine areas of the forests, such as fungi, coleopteran larvae, termites and teredinids. In greater detail, wood-boring molluscs were investigated. Teredinids are the primary focus of this study upon fallen wood in the mid- to low- intertidal areas of mangrove forests. Fallen wood was split apart and all fauna were identified to their lowest achievable taxonomic classification.

This study demonstrates that teredinid-attacked fallen wood in mangrove ecosystems enhances the biodiversity of fauna. However, this study also demonstrates that sea level rise and decreasing pH due to global warming may impact the degradation of LWD in the mangrove forests of Sulawesi.

1.1.1. Indo-Pacific mangrove systems

This study was conducted in eight unusual mangrove forests from the Wakatobi

Marine Park (WMP), Tukang Besi Archipelago, Sulawesi, Indonesia. More than 17,000

Indonesian islands are carbonate in origin (Tomascik et al., 1997). Eastern Indonesia is situated at the junction of three major tectonic plate regions. There are five major fault zones, sutures that separate the tectonic plates. These sutures run along Molucca,

Sorong, Borneo, Sulawesi and the Banda Sea (Hall and Wilson, 2000). It is these major fault regions that may give rise to the numerous fissured rocky islands that have coral- rock intertidal zones. Most Indonesian mangrove forests are associated with muddy fluvial deposits - sediments derived from rivers and streams. Nonetheless, some mangrove forests have developed on calcareous coral-plateaus (Tomascik et al., 1997).

Mangrove forests develop best where low wave energy and shelter enhance the deposition of fine particles, which enables these salt-tolerant trees to establish roots and grow (Alongi 2002). The retention of sediments within mangrove environments reduces the turbidity of the surrounding ocean waters (Engelhart et al., 2007), which enhances the photosynthetic ability of the zooxanthellae within coral colonies as turbid water reduces sunlight penetration (Kleypas 1996). Mangroves are the only trees living at the interface where the terrestrial environment meets the marine environment along tropical coastal regions (Figure 1.1) (Sherman et al., 2003, Middleton and Mckee 2001).

Mangrove forests consist of groups of trees that form distinct zones occupying the eulittoral and also dominating the supralittoral fringes (Woodroffe, 1982; Morton, 1983;

Tomascik et al., 1997). In Indonesia, zones of mangrove trees often include genera such as Rhizophora and Sonneratia, commonly found at the low intertidal zone (Tomascik et al., 1997; Giesen et al., 2006) and trees such as Avicennia, Xylocarpus, Bruguiera and

Excoecaria commonly found further in to the upper to high intertidal zone (Tomascik et al., 1997; Giesen et al., 2006). Zonations of mangrove trees are controlled by environmental factors such as hydrology, salinity, soil chemistry, nutrient availability, predation, competition and also immersion time (Tomascik et al., 1997; McKee and

Faulkner, 2000; Pi et al., 2009). Mangrove zonation is also dependent upon substratum type. Sonneratia alba frequently dominates sandy substrata whereas Avicennia marina are commonly found on muddy areas (Giesen et al., 2006).

Figure 1.1, A Rhizophora-dominated Indonesian mangrove forest at low tide. When flooded, seawater will immerse the exposed Rhizophora stylosa prop-roots that extend

up to 1.5 meters above the sediments.

Indonesia has the largest area of mangrove forests of any country, covering 45,

421 km2 (Spalding et al., 1997); almost one quarter of the World’s mangrove forests.

Such forests are important along Indonesian coastal areas as they play a vital role in the ecological, economic and social development of coastal Indonesian communities

(Tomascik et al., 1997; Alongi, 2007). Indonesian mangrove forests also provide habitat for many transient animals between adjacent ecosystems such as seagrass beds and coral reefs (Sheaves, 2005). Many juvenile and vulnerable fish species enter the mangrove forests at hightide – seeking the refuge of mangrove roots to avoid predation

(Laegdsgaard and Johnson, 2001). Those fish will then follow the ebbing tide to seek refuge with adjacent reefs (Sheaves, 2005). Mangrove forests also support biodiversity within adjacent coral reefs as they deliver organic nutrients to sessile reef-inhabitating 5 invertebrates (Granek et al., 2009). Thus, the connectivity between mangrove ecosystems and coral reefs enhances the survival of the fauna from both habitats. Such connectivity has rarely been studied in Indonesia even though the Indo-Pacific bioregion is considered to be a biodiversity hotspot, containing the world’s richest diversity of marine life (Roberts et al., 2002; Allen, 2008).

Indonesia is in the top ten percent of countries with ecosystems containing a high level of marine animal diversity coupled with high animal endemism (Roberts et al.,

2002; Allen, 2008). When considering mangrove trees, Indonesia has 52 of the world’s

70 mangrove species (Giesen et al., 2006). Thus, mangrove forests have their center of biodiversity richness located in the Indo-West-Pacific, where these forests have important biological and physical functions (Tomascik et al., 1997; Kathiresan and

Bingham, 2001).

Environmental factors such as daily changes in water flow, anoxic sediments and salinity make mangrove environments an unusual habitat for woody plants to become established (Tomascik et al., 1997; Taylor et al., 2005). To combat daily tidal inundation, anoxic sediments and salinity, mangrove trees have morphological (Figure

1.2) and physiological characteristics to enable those plants to thrive. These adaptations include knee and prop roots, pneumatophores, viviparous seedlings, and xeromorphic features of the leaves including thick cuticles, waxy coatings, and sunken stomata

(Clough, 1984; Tomascik et al., 1997; Alongi, 2002). Some of the features that distinguish mangrove trees from terrestrial trees enhance their ability to thrive in hyper- saline soils (Clough, 1984; Alongi, 2002).

Figure 1.2, Adaptations of Indo-Pacific mangrove trees. a, A complex tangle of Rhizophora

stylosa prop-roots, b, densely packed knee-roots belonging to the tree Bruguiera gymnorhiza, c,

large, corky lenticels - allowing for gas exchange on the trunk surface of B. cylindrica, d,

pneumatophores (breathing roots covered in small lenticels) belonging to the Xylocarpus

granatum tree and e, the fruiting structure (flower and propagule) of B. gymnorhiza.

1.1.2. Mangrove biomass and partitioning

Indonesian mangrove forests are very productive. The biomass of these habitats can rival or even surpass that of other marine ecosystems such as coral reefs (Alongi,

2007; Alongi, 2009). Mangrove forest biomass can vary greatly with latitude

(Komiyama et al., 2008), and within sites (Donato et al., 2011). Investigations into whole-ecosystem carbon-storage and sequesteration within mangrove ecosystems found that the majority of carbon is retained within the sediments (Figure 1.3) (Donato et al.,

2011). 7

Figure 1.3, The total carbon storage of Indonesian mangrove forests. The partitioning of

mangrove-derived carbon from the trees, fallen wood, roots and sediments. Data obtained from

Donato et al., (2011).

The partitioning of biomass for individual mangrove trees also varies (Ong et al.,

1995; Ong et al., 2004). The majority of biomass of mangrove trees remains in the trunk, with estimates ranging from 60 – 74 % (Figure 1.4) (Ong et al., 1995; Ong et al., 2004).

A key component of litter in mangrove forests is wood, due to the large proportion of trunk-biomass (Robertson, 1990). Biomass partitioning of mangrove trees among conspecifics varies with tree size. Generally, more biomass is partitioned into the trunks of trees with more than 25 cm circumference, than in smaller trees of the same species:

60 – 70 % compared with 15 – 20 % respectively (Ong et al., 2004). Thus, just a few large trees can significantly contribute to the total above-ground biomass within a mangrove forest (Ong et al., 1995). 8

Figure 1.4, The biomass partitioning of Rhizophora apiculata. From a 20-year old Rhizophora-dominated mangrove forest in Malaysia, 74% of the total biomass is

allocated to the trunks of Rhizophora apiculata. The remaining 26 % of the biomass

is partitioned between the canopy and roots. Figure adapted from Ong et al., (1995).

1.1.3. Detrital Input and Breakdown Pathways

Mangrove forests are a sink for carbon (Bouillon et al., 2008; Donato et al.,

2011), and supports food webs (Figure 1.5) (Robertson and Daniel, 1989a; Feller, 2002;

Maria and Sridhar, 2003). The recycling of wood is much less well understood than that of leaf litter (Kristensen et al., 2008), even though the majority of mangrove biomass is wood (Robertson, 1990; Ong et al., 1995). However, wood has been shown to have a similar carbon flux as leaf litter (Robertson and Daniel, 1989a). The flux of carbon from wood in a Rhizophora-dominated Australian mangrove forest was estimated to be 9 similar to that of leaves with 44g C m-2 y-1 compared with 62g C m-2 y-1 from the consumption of leaves by crabs (Robertson and Daniel, 1989a).

Sesarmid crabs process mangrove leaf litter (Robertson and Daniel, 1989a;

Nordhaus and Wolff, 2007) and shrimps consume much of the mangrove-derived leaf litter detritus (Ronnback et al., 2002; Kenyon et al., 2004). Fallen wood is processed by fungi (Anandar and Sridhar, 2004), coleopteran larvae (Feller, 2002) and termites (Korb and Lenz, 2004). Basidiomycetes are responsible for the initial breakdown of the refractory components of wood in the high intertidal to supratidal areas within mangrove forests (Anandar and Sridhar, 2004). Coleopteran larvae and termites break down wood due to their mechanical processing and digestion within the upper to mid-intertidal zones

(Feller, 2002). However, in the seaward edges of mangrove forests the breakdown of wood is due to the Teredinidae (Robertson and Daniel, 1989a; Robertson, 1990), and limnoriid isopods, although limnoriids are not often found in Indonesian mangrove forests and they only excavate superficial areas of the wood surface (Cragg and Hendy,

2010). The biodegradation processes of large woody debris (LWD) within mangrove ecosystems are essential for nutrient sequestration within the sediments (Donato et al.,

2011). The retention of organic material within the sediments enhances mangrove productivity (Sherman et al., 2003), and the recycling of wood in mangrove forests provide nutrients for adjacent habitats such as seagrass beds and reefs (Granek et al.,

2009). 10

Figure 1.5, Woody detrital input and wood degrading organisms. a, a large fallen log in a mangrove forest – serving as the primary base of an elaborate food web b, the fruiting

body of the fungus Pycnoporus sanguineus, a basidiomycete initiating saprophytic decay of

a large log in the high intertidal c, a cerambid beetle larvae found inside a fallen Excoecaria

agallocha tree and d, termites inside a log also in the high intertidal area of the mangrove

forest.

1.1.4. Role of fallen mangrove wood

Mangrove forests provide an arena where mobile terrestrial animals exploit the forest floor at low tide, arboreal animals thrive within the canopy, benthic animals and pelagic animals utilise the forest at high tide (Cragg and Hendy, 2010). These animals 11 range from a variety of invertebrates such as zooplankton, sponges, ascidians (Perry,

1988), crustaceans (Robertson, 1990), insects and molluscs (Coomans, 1969; Plaziat,

1984; Feller and Mathis, 1997), vertebrates including fish, amphibians, reptiles and birds

(Kathiresan and Bingham, 2001; Giesen et al., 2006). Similar to the zonation patterns of the trees, distinct patterns of distribution can be observed with the mangrove fauna. Root epifaunal distributions are influenced by various abiotic and biotic variables such as tidal flow, salinity, food availability, temperature, competition and predation (Alongi, 2002).

The diversity of animals within the Indo-Pacific mangrove forests is similar to that commonly found in tropical terrestrial forests and marine ecosystems such as the giant centipede, Scolopendra, hunting spiders (Lycosidae) and moray eels Gymnothorax richardsoni and octopus. Such animals are dependent upon fallen logs within the mangrove environment which offer protection for particularly vulnerable animals including juveniles (Figure 1.6). Large woody debris serves as a predation refuge for many aquatic communities (Everett and Ruiz, 1993), therefore fallen wood is an ecologically important resource for increasing biodiversity within mangrove habitats.

Mangrove forests also serve as a nursery refuge for many fish and crustaceans

(Kathiresan and Bingham, 2001; Nagelkerken et al., 2008). Juvenile fish depend upon the complex root systems to avoid predation (MacDonald et al., 2009). The majority of biodiversity is centred on the available hard substrata, as the sediments are generally anoxic and therefore unsuitable for many animals (Prasad and Ramanathan, 2008).

Figure 1.6, Wood-dwelling cryptic fauna removed from fallen logs in Indo-Pacific

mangrove forests. a, a large lycosid spider waits in camouflage on a mangrove log b, an

octopus with eggs found inside a fallen log c, a giant Scolopendra also found inside a

mangrove log and d, one of many early juvenile octopus found inside a mangrove log.

Mangrove ecosystems have a complex topography (Sasekumar and Chong,

1998), and the mangrove trees enhance the forest complexity that increases the level of biodiversity (Gratwicke and Speight, 2005).

The nursery function of mangrove forests has been well documented, especially in Caribbean forests. Many juvenile reef fishes use Caribbean forests as feeding grounds

(Mumby et al., 2003; Nagelkerken and Velde, 2004; Dahlgren et al., 2006). The larger tidal range than the Caribbean within the Indo-Pacific mangrove forests may reduce the activity of feeding reef fish (Nagelkerken and Velde, 2004). Research into the 13 mangrove-nursery function is focused upon juvenile fishes and invertebrates exploiting a variety of connected marginal habitats such as seagrass beds and reefs (Sheaves, 2005;

Unsworth et al., 2008; Luo et al., 2009). Connectivity of adjacent ecosystems, mangrove-seagrass-reef linkages within Indonesian marine habitats are not common

(Tomascik et al., 1997), unlike than in the Caribbean. The movement of different fish species between habitats in the Caribbean are either facultative or obligate meaning that some animals are not bound to any particular habitat and some are (Mumby et al., 2003).

However, some species are dependent exclusively upon the resources available within that habitat. For example, the herbivorous fish, Scarus guacamaia is specifically bound to life in the mangrove forests (Mumby et al., 2003). If it were not for the resources available within the forests the abundance and distribution of Scarus will suffer local extintions (Mumby et al., 2003). The same fate may also be true for mangrove- dependent animals in the Indo-Pacific.

1.2. The Teredinidae (Mollusca: Bivalvia)

The wood-boring family Teredinidae (Bivalvia) consists of fifteen genera and seventy-one species, and they are sedentary molluscs (Turner, 1966). Teredinids create an extensive labyrinth of tunnels within fallen wood through their boring-action (Figure

1.7) (Manyak, 1982; Robertson and Daniel, 1989a). The tunnels they create remain as their home for the duration of their adult life. Teredinids are widely distributed, from tropical to temperate seas; where they are the dominant consumers of fallen wood along coastlines (Distel, 2003). They may be dispersed either as adults in floating wooden debris or as planktonic larvae drifting near the sea surface (Scheltema, 1971). 14

100µm

Figure 1.7, Teredinid shell morphology and teredinid wood processing. a, A teredinid shell, b, a

close-up of the teredinid shell showing the denticulate teeth that help create the tunnels within fallen

wood, c, a cross section sample of fallen wood attacked by the Teredinidae highlighting many

openings of teredinid tunnels and d, a heavily fragmented piece of wood exposing tunnels created

by extensive teredinid boring-activity.

Teredinids have long been considered as economic pests (Laverty, 2002;

Borojevic et al., 2010). They damage man-made wooden structures in marine environments such as boats, piers and pilings (Laverty, 2002; Godwin, 2003; Borojevic et al., 2010). It has been estimated that the cost of damage to marine structures, ranging from ships, docks, wharfs, piers and fishing equipment was one billion dollars U.S in

1986 (Distel, 2003). However, there has been no attempt to investigate the ecological 15 benefit of teredinids within natural habitats such as mangrove forests. Teredinids process large volumes of dead wood within mangrove habitats due to their tunnelling and wood digestion (Distel, 2003), and by doing so they enhance nutrient recycling (Robertson and

Daniel, 1989a).

Investigations of mechanical wood degradation have mainly focused upon coleopteran larvae in Caribbean mangrove forests (Feller and Mathis, 1997; Feller,

2002; Sousa et al., 2003). However, research investigating the decomposition of trees from the activity of teredinids remains limited (Robertson, 1990; Filho et al., 2008). It has been estimated that teredinids process up to 50% of dead mangrove wood

(Robertson and Daniel, 1989a). Without teredinid activity, there would be a major increase of LWD build-up in Indonesian mangrove forests.

1.3. Study Site, The Wakatobi Marine Park

This investigation was conducted between 2006 and 2011 within the WMP, which is the second largest marine park in Indonesia covering an area of 1.39 million hectares (05° 12’ - 06° 10’ S, 123° 20’ - 124° 39’ E). The park is situated within the middle of the Wallacea region – the centre of marine biodiversity (Kwakkel-Hol, 2006).

The marine park consists of four islands, the largest and most northern being Wangi-

Wangi. Southeast to Wangi-Wangi is Kaledupa, and then followed by Tomia and the

Binongko Islands (Figure 1.8). The first two letters of each island make up the name of the marine park, Wakatobi (Kwakkel-Hol, 2006). There are about 10,000 people living in the WMP (Clifton and Unsworth, 2010), and the majority of them are either farmers or fishermen (Kwakkel-Hol, 2006). Over recent years declines in fish stocks due to artisanal fishing practices (Cullen et al., 2007; Cullen, 2010) and increases of mangrove 16 forest harvesting have been reported (Cragg and Hendy, 2010) - to the detriment of the

WMP (Pilgrim et al., 2007).

Figure 1.8, The location of the Wakatobi Marine Park, SE Sulawesi, Indonesia. a, The four

Islands that make the Wakatobi, b, the mangrove sites used in this study. On Kaledupa Island are the

Sombano, Langira, Galua and Laulua mangrove forests, and on the Derawa Island are the mangrove

forests Kaluku, Loho, One Onitu and Gili.

Despite extensive research of the Indo-Pacific mangrove forests (Tomascik et al.,

1997), there has to date been only one study detailing the ecology of the Wakatobi mangrove forests. The study found that of the 70 known mangrove tree species 17

(Polidoro et al., 2010), 24 species are found in the forests on the Kaledupa and Derawa

Islands within the WMP (Figure 1.9). There have been various studies from the WMP where research has concentrated on very detailed investigations of single species

(Barnes, 2003; Jimenez and Bennett, 2005; Taylor et al., 2005). However, no investigation has attempted to study whole mangrove ecosystems in various locations within the WMP.

Many large mangrove forests within Indonesia grow on muddy fluvial sediments

(Tomascik et al., 1997) though there are forests associated with coarse, sandy sediments such as the mangrove forests located in Jakarta Bay and Berau Islands (Tomascik et al.,

1997). Sulawesi has a much smaller area of mangrove forests due to the mountainous and rocky topography and unsuitable substratum (Yulianto et al., 2004). However, there are islands within Sulawesi where mangrove forests thrive, such as the islands of

Kaledupa and Derawa within the WMP. Much of the coastlines from these islands have calcareous mud or sandy sediments that cover areas of fissured raised fossil coral. The forests from these islands are small and some have a close proximity with large areas of seagrass beds and fringing reefs. Such connectivity has rarely been studied along the coasts of Indonesia (Tomascik et al., 1997).

WMP

Figure 1.9, The distributions of the world’s mangrove forests according to the number of different tree species. The world’s mangrove forests illustrated as different colours according to the number of different tree species common to those areas. The forests of the WMP are located in the region with the greatest 1.4. number of different trees, between 36 – 46 species. Figure adapted from Polidoro et al., (2010). 1.5.

1.4. Research Aims and Objectives

The main aims of this investigation were to determine whether the trees within

each mangrove forest of the WMP provides a source of nutrients to food webs via LWD

(Chapter 2), and to determine if the Teredinidae are ecosystem engineers (chapters 3, 4,

5 and 6). This study also aims to determine if LWD from mangrove trees provides a

refuge from predation for mangrove fauna (Chapters 5 and 6). The diversity of animals

removed from teredinid-attacked LWD was studied – to establish if any relationships

occur between a greater number of tunnels and a greater number of animals (Chapter 5).

In addition, drilled wooden blocks were used in the forests to investigate the recruitment

and succession of animals in tunnelled wood. This work investigated details of the

animal biodiversity that vacant teredinid-tunnels may promote. Of particular interest was

a fish, Parioglossus interruptus (Ptereleotridae) found in teredinid tunnels when

emersed, and a spider, Desis martensi (Desidae) found in teredinid tunnels when

immersed (Chapter 6). This study also aims to reveal an unusual ecosystem service – the 19 promotion of biodiversity created by teredinids by understanding how and why animals exploit vacant teredinid tunnels. Finally, this study draws upon the results and then reviews the data within a wider context by discussing implications of the effect of increasing sea-levels and decreasing pH upon mangrove tree distributions, the biodegradation processes and the Teredinidae (Chapter 7).

Using allometric equations, basal areas of mangrove trees were used to calculate biomass. Regression methods were not used, as the felling of trees is required (Kairo et al., 2009). The volume of fallen wood was measured within mangrove forests from the strandline to the seaward edge to determine the amount of LWD available to organisms.

All organisms found within teredinid-attacked LWD were identified to their lowest achievable taxonomic classification. Salinity, emersion time, and temperature were measured to determine if the abundance and distribution of all biodegrading organisms and mangrove fauna within vacant teredinid tunnels were influenced by those variables.

This study tested the hypotheses that mangrove forests in the WMP will have 1. site-specific biodegrading organisms upon woody detritus, and 2. mangrove forests in the WMP will have specific environmental factors that influence the distribution of the biodegrading organisms upon LWD. Furthermore, the Teredinidae will, 3. significantly breakdown LWD through their tunnelling, and 4. enhance ecosystem-level biodiversity within the mid- to low- intertidal areas of the forests through their tunnelling. Finally, this study predicts that through teredinid modification of the environment and habitat creation for other animals, the Teredinidae should be considered as ecosystem engineers.

Chapter 2

MANGROVE FORESTS OF THE WAKATOBI MARINE PARK, EAST SULAWESI, INDONESIA

2.1. Introduction

Mangrove trees are a source of leaf litter and carbon via large woody detritus

(LWD) (Donato et al., 2011), and fallen trees support food webs (Robertson and

Daniel, 1989a, Maria and Sridhar, 2003). However, the contribution of fallen wood is

much less understood than that of leaf litter with respect to nutrient recycling

(Kristensen et al., 2008), even though the majority of mangrove biomass is wood (Ong

et al., 1995). The standing stock of mangrove-derived carbon can rival and surpass that

of other marine systems such as coral reefs (Alongi, 2007; Alongi, 2009). One very

detailed investigation of Indo-pacific mangrove forests has estimated whole ecosystem

carbon storage, which demonstrated that these forests are effective carbon sinks, and

they are the most carbon-rich ecosystems in the Indo-Pacific (Donato et al., 2011).

Mangrove forest biomass can vary greatly with latitude and within sites (Komiyama et

al., 2008; Alongi, 2009; Donato et al., 2011).

Mangrove forest above-ground biomass varies from as low as 7.9 t ha-1 for a

Rhizophora mangle forest in Florida (Lugo and Snedaker, 1974) to as high as 436.4 t

ha-1 for a Bruguiera-dominated forest in eastern Indonesia and 460 t ha-1 in a

Rhizophora apiculata dominated forest in Malaysia (Komiyama et al., 2008).

Patchiness in above-ground biomass has also been reported within individual

mangrove forests. The differences of biomass can change with the tidal heights of 21

mangrove trees. In an Indonesian mangrove forest, Sonneratia sp. found at the

outermost seaward edge had the lowest total biomass, 169.1 t ha-1, in the mid-intertidal

estimates for Rhizophora sp. totaled 356.8 t ha-1 and for the Bruguiera sp. recorded at

the strandline a total above-ground biomass of 436.4 t ha-1 was recorded (Komiyama et

al., 2008), with much of the carbon sequestered into mangrove sediments (Donato et

al., 2011).

Pristine, well-developed mangrove forests contain trees with a greater basal area, tree height, density and species diversity (Roth, 1992) - such as the mangrove forests of the large deltaic systems in Papua (Tomascik et al., 1997; Alongi, 2009).

There are many ways to define the structure and composition of a mangrove forest. Forest structure and composition are determined by the tree species, density, dominance and the distribution of trees (McElhinny et al., 2005). The diversity of tree taxa within a forest can be used as a measure of mangrove forest complexity (Bosire et al., 2006). Mangrove tree basal-area can be used as a measure of woody biomass - studies have demonstrated that tree girths alone can be used to calculate biomass using various allometric equations (Komiyama et al., 2008). However, tree girths and the distributions of trees are determined by environmental variables such as the degree of tidal immersion (Jimenez and Sauter, 1991; Pi et al., 2009), soil chemistry and salinity

(Joshi and Ghose, 2003; Krauss et al., 2008).

This study aims to determine whether the mangrove trees within each forest are a source of LWD that may serve mangrove food webs. Also, do the mangrove trees have particularly large basal areas compared with other forests? Tree species basal areas within each forest were used to calculate mangrove tree biomass, using allometric equations (Komiyama et al., 2005; Komiyama et al., 2008) as this method is less 22 intrusive and less destructive than regression methods – as the felling of trees is required to determine their dry weights (Kairo et al., 2009). This study also aims to reveal the environmental variables that determine tree-species dominance, abundance and their distributions within different mangrove forests within the WMP.

2.2. Materials and methods

2.2.1 Belt transects, tree identification, and measurements of emersion time and salinity

Forest structure surveys were conducted during five 10-week expeditions to the

WMP, Banda Sea, Southeast Sulawesi from June to September on Kaledupa Island (05°

27.53’’S, 123° 46.33’’E) and the Derawa Islands (5° 33'9"S, 123° 52'3"E). The areas of each mangrove forest - Sombano, Langira, Kaluku, Loho, One Onitu and Gili (see chapter 1) - were calculated using a Garmin 76Csx GPS, by way-pointing the perimeters. Continuous belt-transects (2 metres either side of the tape) were completed within the forests beginning at the strandline and ending at the outer edge of each forest.

Depending on the size of the forest investigated, between three and five transects were established by calculating the area of each forest. A number of measurements were made along every transect, each mangrove tree within two metres either side of the transect tape was recorded and identified. Tree heights were calculated by using measured distances from a Nikon Laser Rangefinder 550 or by estimating tree height using a two metre wooden pole. The substratum of the entire transect was also recorded. Ground water salinities were measured using a Bellingham and Stanley E-Line Aquatic hand- held refractometer. To determine the tidal profile within the forest, high tide level was 23 marked on mangrove trees using high visibility string. At low tide, the distances from the substratum to the mark on each tree were measured and then subtracted from the mark of seawater at high tide to establish the height above sea level. Emersion times were estimated by relating the tidal heights to chart datum in regional tide tables.

2.2.2. Mangrove forest descriptions

Sombano

The Sombano mangrove is a large land-locked forest with mature Excoecaria - and Bruguiera trees, where tidal flooding occurs during only spring tides or other exceptionally high tides. Tidal flooding in Sombano is peculiar; seawater percolates from under a raised fossil coral plateau, at the seaward end of the forest.

Sombano has several species of mangroves. The forest has large trees with girths of B. gymnorhiza measuring up to more than two metres. The forest also has a modest mixed sapling bank from the strandline to the seaward edge. Extensive stands of twenty to twenty-five metre high Excoecaria agallocha and Bruguiera gymnorhiza trees dominate, with emergents of twenty metre Rhizophora apiculata and Xylocarpus spp.

The most dominant species is E. agallocha, which was found in high numbers. Many of the larger B. gymnorhiza trees have been cut down, therefore reducing the mean tree girths for B. gymnorhiza in Sombano. Very large mature R. apiculata trees are also harvested in Sombano (Figure 2.1 a). In periods of heavy rainfall, freshwater run-off enters the Sombano forest from the landward edge in the form of small streams (Figure

2.1 b).

Langira

The Langira strandline is situated at the base of the fossil-coral plateau that 24 marks the end of the land-locked Sombano forest. The forest is established on a fossil- coral plateau that extends down to mean sea level. The sediments over-laying the plateau are a mix of mud at the strandline changing to a mixture of sand and shells at the outermost seaward edge.

a b

c d

e f

Figure 2. 1 Images of mangrove sites on Kaledupa and Derawa. a, A mature felled Rhizophora apiculata tree in the inland Sombano forest. Note the straight-edged planks cut from a mobile saw mill; evidence of forest harvesting b, Freshwater run-off streaming in to the Sombano forest strandline c, The fringing Langira forest. Small stands of R. stylosa trees grow in pockets of soft sediments in fissured coral holes within the coral-plateau. A sapling Bruguiera gymnorhiza tree can be seen in the foreground d, Dense R. stylosa stands in the Kaluku forest on Derawa e, The landward zone of Loho. The fossil-coral shore line of Loho can clearly be seen in the background f, The inland Gili forest on Derawa. The fossil-coral strandline can be seen in background behind the large mature B. gymnorhiza trees in the foreground and f, One Onitu at low tide, exposing the dense prop-roots of the many R. stylosa stands.

The substratum is interspersed with many crowns of fissured coral-rocks protruding through. In places, the fossil-coral becomes exposed, with numerous coral holes consisting of calcareous mud. Langira had many large open channels also consisting of calcareous mud with dense seagrass beds over laying these channelled areas. Distinct zonation patterns of trees was observed, with Rhizophora apiculata found at the strandline, and R. stylosa populations extending out into the eulittoral zones. The

Rhizophora trees in Langira were very shrubby, and were established in coral sink-holes

(Figure 2. 1 c). Tree canopies are very patchy, and maximum tree heights reach no more than twelve metres, with trees averaging four meters.

The Rhizophora and Bruguiera trees attained the largest basal area within

Langira. Thus, those trees were important contributors to the total forest biomass. High densities of R. stylosa were found in Langira. Bruguiera gymnorhiza had the greatest tree-girths with some trees measuring almost two metres in girth as did a few

Xylocarpus granatum trees. Xylocarpus spp. were found only at the strandline.

Kaluku

The Kaluku forest (Figure 2. 1 d) is constrained at the landward edge by a fossil- coral wall. The forest is inundated by all tides, and no salinity gradient was found.

Kaluku had almost mono-specific stands of R. stylosa, with very few S. alba trees.

Located at the landward edge of Kaluku outside of the chosen plots, Bruguiera gymnorhiza, Xylocarpus granatum and Pemphis acidula were present. R. stylosa trees were densely populated in Kaluku, the girth of these trees were comparatively small and therefore the total basal area was also small compared to the very large biomass of the few S. alba trees.

Loho

The Loho forest was the largest of the mangrove forests surveyed on Derawa.

The forest is established within a sheltered bay and contains a large abundance of R. stylosa trees. Lining the Loho strandline was a large fissured, over-hanging fossil-coral embankment (Figure 2. 1 e). Within the forest from the strandline and out to the seaward edge, R. stylosa was the dominant species in relation to both abundance and basal area.

B. gymnorhiza stands were also present, this species was only located at the landward zones within the plots with trees up to sixteen meters in height located within the belt transect.

Gili

The Gili forest on Derawa was separated with the One Onitu forest by a fossil- coral plateau, similar to that of the Sombano and Langira mangrove forests. The Gili forest is an inland forest surrounded by a fossil-coral wall (Figure 2. 1 f).The forest is the smallest of all the sites studied. Seawater enters the forest through a tunnel under the fossil-coral plateau that separates Gili and One Onitu. The Gili forest is inundated by all tides, except neap tides less than 1.4 metres above sea level. No difference of tidal height was found from the strandline to the seaward edge. The middle region of the forest had the lowest recorded emersion times when compared to the landward and the outermost seaward edges suggesting that the middle region of the forest was ‘sunken’ compared to the outer edges of the forest - similar to the topography of the Sombano forest. The composition of the substratum changed with distance from the land. Within the landward regions, the substratum consisted of a thick organic-rich mud - in places the mud was more than forty centimeters deep. The middle area of the forest had a calcareous mud substratum, although the area where seawater enters into the forest the 28 substrata was predominantly sand. The changing substratum was a determining factor for the tree distributions within Gili. S. alba were found only in sand-dominated substrata, whereas R. apiculata dominated in areas of thick organic-rich mud. The most abundant tree species within Gili was Bruguiera gymnorhiza and the forest also had a young bank of Rhizophoraceae saplings. Large, mature Sonneratia alba trees were present at the seaward zone, and Rhizophora apiculata were in the landward regions.

One Onitu

On the exposed seaward edge of the fossil-coral plateau in front of the inland

Gili forest is One Onitu (Figure 2. 1 g). Like the Langira forest, One Onitu also has connectivity to seagrass and reefs. The entire One Onitu forest is established in the high intertidal and no difference was found with emersion time across the forest. B. gymnorhiza trees were found in areas with higher emersion times. Furthermore, the substratum in the forest consisted of calcareous mud. R. stylosa trees were the most common species in One Onitu with very few B. gymnorhiza and S. alba trees.

In the Derawa forests, anthropogenic impact was slight and removal of mangrove material was minimal. However, the Kaledupan mangrove forests, Sombano and Langira are exposed to a high level of wood harvesting. The constant sound of chain-saws echoed in Sombano. Frequent evidence of felled mature R. apiculata and B. gymnorhiza trees are observed with many clearings with fresh milled debris consisting of saw dust and straight-edged cut planks being found within Sombano. Large bundles of newly felled young B. gymnorhiza trees were also found and large amounts of B. gymnorhiza trees were observed in storage beneath the houses of local residence living on Kaledupa.

Furthermore, the abundant R. stylosa trees in Langira were cut on a daily basis. Unlike the case in Sombano, whole trees in Langira were not often cut down; however branches 29 were harvested for fish fences. The level of cutting in Langira is high, with almost every tree within the forest experiencing some degree of harvesting.

2.2.3 Tree densities and species basal area

Along each of the transects within the Sombano and Langira mangrove forests, replicate 20 x 20 m plots were selected at random from a total of between five and 25 plots depending upon the length of the transect. Sombano had a total of 22 plots and

Langira had a total of 19 plots. For the forests in Derawa two 10 x 10 m plots were selected at random from a total of between three and 12 plots depending upon the length of each transect. In the Kaluku forest a total of 12 plots in six transects were established.

In the Loho forest, 10 plots in five transects were established. In the One Onitu forest, 15 plots in five transects were established. In the Gili forest six plots in three transects were established (Table 2.1). To determine the basal areas of each tree species within the plots between the sites, diameter at breast height (DBH), or the circumference at breast height

(CBH) of each tree were recorded using the equation r2 (where = 3.142 x radius2).

The radius was calculated by dividing the circumference by x 2, or by dividing the diameter by 2.

30 Table 2.1 Plots labelled 1-5 used to measure mangrove tree basal areas and tree

species density within each site and transect. Numbers within each site are the

distance from the landward edge – the strandline (metres).

Plots Site Transect 1 2 3 4 5 Sombano 1 100 140 240 260 340 2 100 140 240 260 340 3 100 140 240 260 340 4 100 140 240 260 5 100 140 240 Langira 1 0 160 320 440 460 2 0 160 320 440 460 3 0 160 320 4 0 160 320 5 0 160 320 Kaluku 1 20 30 2 0 10 3 0 20 4 0 20 5 10 20 6 0 50 Loho 1 10 40 2 0 20 40 3 10 50 4 80 150 5 20 One Onitu 1 0 20 30 60 2 10 40 50 70 3 0 40 60 4 40 70 5 30 70 Gili 1 30 70 2 0 10 3 10 30

2.3 Data analysis

The univariate and non-parametric multivariate techniques, of multidimensional scaling plots (MDS) distance-based linear modelling (DISTLM) and DIVERSE analyses contained in the PRIMER 6 (Plymouth Routines in Multivariate Ecological Research) package were used to explore forest characteristics. Similarities of botanical communities between the sites were displayed using non-metric MDS plots. To investigate spatial patterns of similarity, each mangrove forest was grouped by overlaying data from a Bray-Curtis Cluster analysis. DISTLM was employed to verify relationships between emersion time, distance from the land (DFL), salinity and substratum type with the mangrove trees. DISTLM produces a sequential test, which assesses the variation each predictor (environmental variable) has on its own, and a conditional test, which explains the variation of all the variables (McArdle and

Anderson, 2001). The most parsimonious model was identified using the Akaike information criterion (AIC). AIC ranks models from all possible combinations of the environmental variables. Distance-based redundancy analyses (dbRDA) were used for visualizing the results as an ordination, constrained to linear combinations of the predictor variables. The DISTLM was based on abundance data with 4999 permutations.

To measure variation in botanical diversity between forests the following measurements were used, total species = S, total individuals = N, Pielou’s evenness = J’ and Shannon-Weiner diversity index = H’. The Shannon-Weiner index is calculated by,

H’ = - pi ln pi, where pi is the proportion of individuals that taxon i contributes to the total sampled population. Pielou’s evenness was expressed by; J¢ = H¢/H¢max, where 32

H¢max is the maximum value of diversity for the number of taxa present (see Magurran,

2004 for details).

2.4 Results

2.4.1. Mangrove area and botanical distributions

The Sombano and Langira forests cover the largest area of all the sites studied, 50 and 60 ha respectively. The mangrove forests on Derawa - Kaluku, Loho, One Onitu and

Gili are very small, with areas of 0.5 ha, 3.1 ha, 1, and 0.9 respectively (Table 2.2).

Table 2. 2, Areas of mangrove forests investigated and mean distance from the strandline to

the seaward edge. The area of the sites are in hectares and the transect lengths are in metres

(mean ± SE).

Site Area (ha) Transect length (mean ± SE)

Sombano 50.0 444.0 ± 42.6

Langira 60.0 388.0 ± 16.2

Kaluku 0.5 37.8 ± 5.9

Loho 3.1 103.6 ± 18.0

One Onitu 1.0 91.8 ± 6.4

Gili 0.9 71.3 ± 16.7

Tree distributions within the WMP (Table 2.3) such as Rhizophora stylosa were found in the mid- to low- areas of each mangrove forest. The only mangrove forest 33

investigated without R. stylosa is the land-locked Sombano forest. Rhizophora apiculata

was found in every forest accept Loho and One Onitu. Bruguiera gymnorhiza occurred

at each forest and Sonneratia alba was found in every site except Sombano.

Table 2. 3, Mangrove plant species found in the WMP. Sites: S = Sombano, L = Langira, K =

Kaluku, Lo = Loho, O = One Onitu & G = Gili. Taxonomic authorities were obtained from the

International Plant Names Index.

Species Family Sites Acanthus ebracteatus Vahl. Acanthaceae S Acrostichum speciosum Willd. Pteridaceae S Aegiceras floridum Roem. & Schult. Myrsinaceae L Avicennia marina Vierh. Avicenniaceae L Bruguiera cylindrica Blume. Rhizophoraceae S Bruguiera gymnorhiza (L.) Sav. Rhizophoraceae S, L, K, Lo, O, G Casuarina equisetifolia J.R. Forst. & G. Forst. Casuarinaceae L Cordia subcordata Lam. Boraginaceae L Crinum asiaticum L. Amaryllidaceae L Derris trifoliata Lour. Leguminosae S Drynaria quercifolia (L.) J. Sm. Polypodiaceae S Dolichandrone spathacea K. Schum. Bignoniaceae S Excoecaria agallocha L. Euphorbiaceae S Ficus sp. Moraceae S Heritiera littoralis (Dryand.) Sterculiaceae S Hibiscus tiliaceus L. Malvaceae L Nypa fruticans Wurmb. Arecaceae S Osbornia octodonta F. Muell. Myrtaceae L Pemphis acidula Forst. Lythraceae L, K Rhizophora apiculata Blume. Rhizophoraceae S, L, K, G Rhizophora stylosa Griff. Rhizophoraceae L, K, Lo, O, G Sonneratia alba Griff. Sonneratiaceae L, K, Lo, O, G Xylocarpus granatum Koen. Meliaceae S, L, K Xylocarpus moluccensis M. Roem. Meliaceae S, L

2.4.2. Environmental factors

Salinity

The lowest salinity levels of this study were recorded from Sombano, which had a brackish water environment (Table 2.4). In Sombano a gradient of salinity from the strandline was recorded, beginning at zero PSU and increasing to twenty-eight PSU at the seaward edge (refer to Figure 2.2). Conversely, no salinity gradients were observed in Langira, Kaluku, Loho, One Onitu and Gili - as salinities were almost fully marine ranging from 31.5 – 34.8 PSU.

Table 2.4, Mangrove forest variation of salinity levels. Salinity

measurements defined as PSU were recorded along each transect at

five metre intervals within every forest studied (mean ± SE).

Site Salinity

Sombano 15.3 ± 0.7

Langira 33.3 ± 0.1

Kaluku 34.4 ± 0.2

Loho 31.5 ± 0.2

One Onitu 32.2 ± 0.2

Gili 34.8 ± 0.2

Emersion

Emersion times (hours) within the high, middle and low intertidal zones differed within and between the mangrove forests (Table 2.5). The Sombano and Gili forests had 35 the greatest emersion times in each intertidal zone when compared with the other sites.

Sombano and Gili are basin-type mangrove forests, which explain the long emersion times recorded from the low-intertidal zones.Langira, Kaluku, Loho and One Onitu each had similar emersion times when compared with each other within each intertidal zone.

Table 2.5, Variation in emersion time between different mangrove forests. The

intertidal zones of each forest include, the strandline designated as the high intertidal,

the centre of the mangrove is considered as the mid intertidal and the seaward fringing

areas have been classified as the low intertidal zones. Emersion is the time each

mangrove forest zone is not immersed in seawater over a 25 hour spring tidal cycle

(hours, mean ± SE).

Mangrove forest area

Site Strandline Centre Seaward

Sombano 18.5 ± 0.4 18.4 ± 0.3 20.8 ± 0.4

Langira 9.6 ± 1.1 6.5 ± 0.1 5.4 ± 0.1

Kaluku 8.8 ± 0.6 7.0 ± 0.2 5.3 ± 0.5

Loho 9.4 ± 1.1 6.4 ± 0.1 5.7 ± 0.2

One Onitu 8.6 ± 0.6 7.4 ± 0.7 7.3 ± 0.2

Gili 12.4 ± 0.1 11.9 ± 0.2 13.2 ± 0.4

*The Sombano and Gili forests both have a sump-like shore height profile. The centre

regions of both forests have the lowest recorded height above sea levels compared

with the strandline and seaward edges. 36

2.4.3. Canopy-forming trees and forest characteristics

The investigation into the canopy forming trees and forest characteristics of the sites (Figure 2.2) revealed that the tallest trees were found in the Sombano forest with heights of trees greater than 23 metres. The Gili forest also had very large trees, ranging up to 22 metres. However, the forest canopies within Langira, Kaluku, Loho and One

Onitu were notably lower. The tallest trees from those forests ranged in heights between

8 and 16 metres.

The main substratum in Sombano and Gili was organic-rich mud. However, beyond 70 metres from the strandline within the Gili forest, the substratum changed from organic-rich mud to sand. Langira also had organic-rich mud, but only between 0-

200 metres from the strandline. Extending from 200 metres from the strandline and out to the seaward edge in Langira the substratum consists of calcareous mud and sand with some seagrass beds. Kaluku, Loho and One Onitu each have a calcareous mud and sand substratum from the strandline to the seaward edges. In addition, each mangrove forest had protrusions of raised fossil-coral.

Figure 2.2 Mangrove canopy-forming trees and site-specific environmental variables. (a), the distribution of canopy-forming trees, (b) hydrographic conditions and (c), substratum type along transects at the six mangrove forests investigated - continues from page 36 to 41. 37 Sombano

Bruguiera gymnorhiza Excoecaria agallocha Tree stumps Heritiera littoralis Nypa fruticans Rhizophora apiculata Xylocarpus granatum Xylocarpus moluccensis a 25

Tree Tree height (m) 5

0 0 50 100 150 200 250 300 350 400 450

Emersion Salinity b 25 40 35 20 30 15 25 20

10 15 Salinity Emersion Emersion (h) 10 5 5 0 0 0 50 100 150 200 250 300 350 400 450 Distance from strandline (m) c = Raised fossil coral = Organic-rich mud

Substratum type 38 Langira

a Rhizophora apiculata Rhizophora stylosa Sonneratia alba Xylocarpus granatum Bruguiera gymnorhiza 25

Tree Tree height (m) 5

0 0 50 100 150 200 250 300 350 400 b 25 Emersion Salinity 40 35 20

15 25 20

10 15 Salinity Emersion Emersion (h) 10 5 5 0 0 0 50 100 150 200 250 300 350 400 Distance from strandline (m) c = Raised fossil coral = Organic-rich mud = Calcareous mud = Seagrass beds

Substratum type 39

Kaluku a 25 Rhizophora stylosa

10 Tree Tree height (m) 5

0 0 5 10 15 20 25 30 35

Emersion Salinity 25 40 35 20

15 25

20 Salinity

10 15 Emersion Emersion (h) 10 5 5 0 0 0 5 10 15 20 25 30 35 Distance from strandline (m) c = Raised fossil coral = Sand = Calcareous mud

Substratum type 40

Loho a 25 Bruguiera gymnorhiza Rhizophora stylosa

10 Tree Tree height (m)

0 0 20 40 60 80 100 120 b 25 Emersion Salinity 40 35 20

15 25 20

10 Salinity Emersion Emersion (h) 15 10 5 5 0 0 0 20 40 60 80 100 120

Distance from strandline (m) c = Raised fossil coral = Calcareous mud

Substratum type 41 One Onitu a Bruguiera gymnorhiza Rhizophora stylosa Sonneratia alba 25

Tree Tree height (m) 5

0 0 20 40 60 80 100

b 25 Emersion Salinity 40 35 20

15 25 20 10 Salinity Emersion Emersion (h) 15 10 5 5 0 0 0 20 40 60 80 100

Distance from strandline (m)

c = Sand = Calcareous mud

Substratum type 42

Gili a Bruguiera gymnorhiza Rhizophora apiculata Rhizophora stylosa Sonneratia alba 25

Tree Tree height (m) 5

0 0 10 20 30 40 50 60 70 80 90 100 b Emersion Salinity 25 40 35 20

15 25 20

10 15 Salinity Emersion Emersion (h) 10 5 5 0 0 0 10 20 30 40 50 60 70 80 90 100

Distance from strandline (m) c = Organic-rich mud = Calcareous mud = Sand = Raised fossil coral

Substratum type 43

2.4.4. Tree species dominance - basal areas, biomass and tree densities

Basal areas of the species found varied considerably (Table. 2.6). Within

Sombano, E. agallocha and B. gymnorhiza were the most dominant species. The Langira forest had lower basal areas in comparison to Sombano. However, the most dominant species in Langira were B. gymnorhiza and R. stylosa. In Derawa, R. stylosa was dominant within all forests except Gili. In Kaluku, the most dominant species was S. alba, although only one species was recorded within the plots. Bruguiera gymnorhiza was dominant in One Onitu and Gili, and S. alba had a large basal area within the Gili forest – however, only one species was also recorded within the plots.

Basal area measurements of the E. agallocha trees in Sombano (Figure 2.3) were dominant in each of the 22 plots investigated. Bruguiera gymnorhiza was also very dominant, however in the landward regions this species was less dominant compared with E. agallocha. Rhizophora apiculata was dominant in areas of the forest – but this species was patchy. Heritiera littoralis was dominant at the strandline; however the basal area measurements of this species decreased with increasing distance from the strandline.

2 -1 Table 2.6, The variation of tree-species basal area measurements from each of the mangrove forests investigated. All basal area measurements (m ha ) were recorded from trees within designated plots in each mangrove forest (mean ± SE).

Tree Species Site Sombano Langira Kaluku Loho One Onitu Gili Bruguiera cylindrica 0.02 ± 0.08 Bruguiera gymnorhiza 9.98 ± 1.93 1.63 ± 0.44 1.7 ± 1.14 21.08 ± 7.87 16.31 ± 3.56 Excoecaria agallocha 11.04 ± 1.2 Heritiera littoralis 2.74 ± 1.97 Rhizophora apiculata 3.24 ± 1.25 0.32 ± 0.16 7.1 Rhizophora stylosa 1.37 ± 0.25 10.75 ± 1.4 14.75 ± 2.1 7.36 ± 2.06 Rhizophoracae saplings 0.14 ± 0.1 Sonneratia alba 0.35 ± 0.26 12.16 ± 12.16 13.25 ± 7.5 33.8 Xylocarpus granatum 0.96 ± 0.47 0.66 ± 0.66 Xylocarpus moluccensis 0.03 ± 0.03

Bruguiera cylindrica 25 Bruguiera gymnorhiza 0.4 20 0.3 15 0.2 10

5 0.1

0 0 100 140 240 260 340 100 140 240 260 340

Excoecaria agallocha

25 25 Rhizophora apiculata

)

1 -

20 20

2 15 15

10 10

5 5 Basalarea(m 0 0 100 140 240 260 340 100 140 240 260 340

Heritiera littoralis 25 Xylocarpus granatum 25

20 20

15 15

10 10

5 5

0 0 100 140 240 260 340 100 140 240 260 340 Distance from strandline (m) Distance from strandline (m)

Figure 2.3, Basal areas of mangrove species in Sombano. Differences between mangrove

tree basal areas (mean ± SE) recorded from within plots (n = 22) at different distances from

the strandline. *The y-axis of the figure belonging to B. cylindrica is a different scale to all

other figures, due to the very low basal area. 46

Within the Langira forest, Rhizophora stylosa was dominant in each of the 19 plots, the basal area measurements increased with increasing distance from the strandline (Figure 2.4). Bruguiera gymnorhiza was also dominant within the plots from Langira. R. apiculata was found only in regions at the high-intertidal.

Sonneratia alba had a low abundance and was very patchy, but the basal area was large as these species have a large circumference. Xylocarpus granatum also had a very low abundance in Langira and this species was found only at the strandline, with some very large individuals.

In the plots from Kaluku only two tree species were recorded (Figure 2.5).

Rhizophora stylosa was very dominant and found in 11 of the 12 plots. Although

Sonneratia alba had a very large basal area measurement, this species was found only in 1 plot located at the strandline.

Within the Loho forest (Figure 2.6), Rhizophora stylosa was very dominant.

This species was found in 9 of the 10 plots. However, Bruguiera gymnorhiza had a comparatively low basal area and was found close to the strandline.

Bruguiera gymnorhiza was dominant within the One Onitu forest from the mid-intertidal and below (Figure 2.7). Rhizophora stylosa had a similar distribution to B. gymnorhiza however this species was less dominant. Sonneratia alba was very patchy, and found only from the strandline and down to the mid-intertidal - S. alba was very dominant with very large basal area measurements.

The most dominant tree species in the Gili forest was Bruguiera gymnorhiza

(Figure 2.8). However, there were emergent Rhizophora apiculata trees found within the mid-intertidal areas and Sonneratia alba trees found only at the seaward edges. 47

Only one S. alba tree was recorded within the plots, however this specimen had a very large basal area.

Rhizophora stylosa 10 10 Rhizophora apiculata

8 8

6 6

4 4

2 2

0 0 0 160 320 440 460 0 160 320 440 460

Sonneratia alba

10 Bruguiera gymnorhiza 10

)

1 - 8

ha 8

2 6 6

4 4

2 2 Basalarea(m

0 0 0 160 320 440 460 0 160 320 440 460

10 Xylocarpus granatum 0.8 Xylocarpus moluccensis 0.7 8 0.6 6 0.5 0.4 4 0.3 0.2 2 0.1 0 0 0 160 320 440 460 0 160 320 440 460 Distance from strandline (m) Distance from strandline (m)

Figure 2.4, Basal areas of mangrove species in Langira. Differences between mangrove tree

basal areas (mean ± SE) recorded from within plots (n = 19) at different distances from the

strandline. * The y-axis of the figure belonging to X. moluccensis is a different scale to all other

figures, due to the very low basal area.

140 Rhizophora stylosa

120

) 1

- 100

2 m 80

Basalarea( 40

0 0 10 20 30 40 50 140 Sonneratia alba

120

)

1 -

ha 100

2 80

40 Basal area (m area Basal

0 0 10 20 30 40 50 Distance from strandline (m)

Figure 2.5, Basal areas of mangrove species in Kaluku. Differences between mangrove tree basal areas (mean ± SE)

recorded from within plots (n = 12) at different distances from

the strandline.

35 Rhizophora stylosa

) 1

- 25

2 20

Basal area (m area Basal 5

0 0 10 20 30 40 50 60 70 80 90 100 110 120 130 140 150

35 Bruguiera gymnorhiza

) 1

- 25

2 20

Basal area (m area Basal 5

0 0 10 20 30 40 50 60 70 80 90 100 110 120 130 140 150 Distance from strandline (m)

Figure 2.6, Basal areas of mangrove species in Loho. Differences

between mangrove tree basal areas (mean ± SE) recorded from

within plots (n = 10) at different distances from the strandline.

90 Brugueira gymnorhiza 80 70 60 50 40 30 20 10 0 0 10 20 30 40 50 60 70

Rhizophora stylosa 90

) 70

1 -

ha 60

2 50 40 30 20 Basal area (m area Basal 10 0 0 10 20 30 40 50 60 70

90 Sonneratia alba 80 70 60 50 40 30 20 10 0 0 10 20 30 40 50 60 70 Distance from strandline (m)

Figure 2.7, Basal areas of mangrove species in One Onitu. Differences between mangrove tree basal areas (mean ± SE) recorded within plots (n = 15) at different distances from the strandline.

Bruguiera gymnorhiza Rhizophoracea saplings

) 0.7

1 -

30 0.6

2 25 0.5 20 0.4 15 0.3 10 0.2

5 0.1 Basalarea(m 0 0 0 10 20 30 40 50 60 70 0 10 20 30 40 50 60 70

35 Sonneratia alba

) Rhizophora apiculata 35

1 -

30 30

2 25 25 20 20 15 15 10 10

5 5 Basalarea(m 0 0 0 10 20 30 40 50 60 70 0 10 20 30 40 50 60 70 Distance from strandline (m) Distance from strandline (m)

Figure 2.8, Total basal areas of mangrove species in Gili. Differences of the total

basal areas (mean ± SE) recorded for each individual mangrove tree within plots (n = 6)

at different areas from the strandline. * The y-axis of the figure belonging to

Rhizophoracea saplings is a different scale to all other figures, due to their low basal

areas.

Biomass estimates from the basal area measurements were calculated using allometric equations (Komiyama et al., 2005; Komiyama et al., 2008) (Table 2.7).

Sombano had the greatest estimates of biomass, followed by Langira. The mangrove forests in Derawa, Kaluku, Loho, One Onitu and Gili had comparatively low estimates of biomass.

Table 2.7, Biomass estimates of mangrove species from each of the sites studied. Differences of biomass estimates (mean ± SE) of mangrove

basal areas recorded from plots within each forest. Allometric equations were used to calculate biomass. Wood density constants for B. gymnorhiza, B. cylindrica, B. saplings, S. alba = 0.186*DBH+2.31, R. saplings, R. stylosa = 0.128*DBH+2.6, X. spp. = 0.0823*DBH+2.59, E.

agallocha, H. littoralis = 0.168*DBH+2.47 and R. apiculata = 0.235*DBH+2.42.

Species Site (t ha) Sombano Langira Kaluku Loho One Onitu Gili Bruguiera cylindrica 0.4 ± 0.1 Bruguiera gymnorhiza 328.6 ± 0.1 50.3 ± 0.3 1.3 ± 0.1 32.3 ± 0.6 7.8 ± 0.1 Excoecaria agallocha 605.8 ± 0.1 Heritiera littoralis 223.6 ± 1.4 Rhizophora apiculata 236.8 ± 1.0 13.3 ± 0.1 1 Rhizophora stylosa 83.0 ± 0.03 18.5 ± 0.03 28.3 ± 0.04 21.6 ± 0.05 Rhizophoracea saplings 0.1 ± 0.003 Sonneratia alba 12.2 ± 1.5 3.4 21.4 ± 1.4 3.4 Xylocarpus sp. 66.4 ± 0.3 111.2 ± 26.7

*The breakdown of the allometric biomass equations for e.g. Rhizophora apiculata 0.235*DBH+2.42; where 0.235 = wood density + 2.42 (constant) x log10 of

the DBH (diameter at breast height). To obtain the biomass in kgs, the log10 mass is exponentialy raised to the power of 10 and then divided by 1000 to calculate

the biomass weight in tons. 53

The densities of trees between the sites varied considerably (Table 2.8). There were more than 750 Bruguiera gymnorhiza trees per hectare in Sombano, however less than 65 B. gymnorhiza per hectare were estimated to be in Langira. Differences of numbers of

Rhizophora stylosa were also found. In the Langira forest there were less than 260 R. stylosa trees per hectare. However, there were more than 900 R. stylosa per hectare in the Loho mangrove forest. The least abundant mangrove trees were B. cylindrica in Sombano, and

Xylocarpus moluccensis found in the high intertidal areas within Langira.

Table 2.8, Densities of different tree species in mangrove forests within the study sites. Tree densities per hectare were recorded from plots within each of the mangrove sites investigated (mean ± SE).

Tree Species Site Sombano Langira Kaluku Loho One Onitu Gili

Bruguiera cylindrica 2 ± 1.4

Bruguiera gymnorhiza 766 ± 231.42 61 ± 22.9 100 ± 32.38 333.3 ± 88.18

Excoecaria agallocha 563 ± 105.4

Heritiera littoralis 51 ± 18.1

Rhizophora apiculata 77 ± 26.18 42 ± 20.8 16 ± 16.66

Rhizophora stylosa 255 ± 42 781 ± 64.4 930 ± 98.95 420 ± 115.5

Rhizophoracea saplings 216.7 ± 79.24

Sonneratia alba 6 ± 5.06 0.1 ± 0.1 40 ± 23.5 16 ± 16.66

Xylocarpus granatum 67 ± 38.5 3.0 ± 2.2

Xylocarpus moluccensis 1.0 ± 1.0

2.4.5. Site similarity and tree species diversity

A Multidimensional Scaling plot (MDS) of mangrove forest similarities of the composition and abundance of trees was used to compare patterns of tree diversity between each mangrove forest (Figure 2.9). The inland mangrove forests,

Sombano and Gili were more similar to each other compared to the seaward forests with 50 % similarity to each other. The Kaluku and Loho forests are established in sheltered bays and they share an 80 % similarity. The MDS plot also grouped One

Onitu and Langira with 50 % similarity – both forests are established in exposed coastal areas.

Transform: Fourth root Resemblance: S17 Bray Curtis similarity (+d) 2D Stress: 0.1 Sites Sombano Langira Kaluku Loho Onitu Gili

Figure 2.9, An MDS plot illustrating similarities between mangrove tree community

structure and abundance of trees within each of the mangrove forests investigated. The data

were square root transformed and a Bray Curtis similarity matrix was used to calculate inter-

sample similarities within each of the forests.

Analyses of the diversity of mangrove trees (Table 2.9) determined that

Sombano has the highest tree diversity, and abundance of species. The Langira forest had a relatively high diversity and evenness. The analysis for both Loho and Kaluku revealed they had the lowest tree diversity. This was also mirrored with low evenness values from the forests – also due to the dominance of R. stylosa. Although One

Onitu had a low density and abundance of species, the Pielou’s evenness value was higher than any of the other mangrove forests.

Table 2.9, Comparison between the various indices of tree species between

the mangrove forests investigated. S = number of species, N = total individuals,

J’ = Pielou’s evenness, H’ = Shannon-Weiner diversity index.

Site S N J’ H’ Sombano 6 1526 0.627 1.124 Langira 6 368 0.515 0.922 Loho 2 980 0.291 0.202 Onitu 3 560 0.648 0.712 Kaluku 2 782 0.002 0.001 Gili 4 583 0.643 0.891

Distance-based redundancy analysis (dbRDA) was used for visualizing the correlations of the trees with emersion time, substrate and salinity (environmental variables) measured from each forest and plot as an ordination, constrained to linear combinations of the predictor variables. The environmental variables influencing the dbRDA ordination axis 1 (Figure 2.10 a) showed that 94.7 % of the environmental data explains 45.2 % of the total variation and for axis 2, 5 % of the environmental 57

data explains 2.4 % of the total variation. The distributions of trees within each

mangrove forest except Sombano were strongly affected by emersion time and the

substratum type (Figure 2.10 b) and the trees within Sombano were strongly affected

by salinity (Figure 2.10 b). However, a few trees from Langira were also grouped

within Sombano. The trees, R. stylosa, R. apiculata and B. gymnorhiza were affected

by emersion time and substratum type (Figure 2.10 c). However, E. agallocha, X.

granatum and H. littoralis were strongly correlated with lower salinity (Figure 2.10

c). Two distinct groups were found, as the Sombano forest was different from the

other forests.

b c

Figure 2.10, dbRDA ordinations. a, the distribution of trees identified from plots within each mangrove forest. Different coloured symbols signify different mangrove forests, b, the environmental vectors that influence the trees within each mangrove site and, c, the tree distribution within the mangrove forests explained as vectors. Ea = Excoecaria agallocha, Hl = Heritiera littoralis, Xg = Xylocarpus granatum, Ra

= Rhizophora apiculata, Bg = Bruguiera gymnorhiza and Rs = Rhizophora stylosa. PSU = practical salinity units. 58

A conditional test revealed that emersion time explains 94.7 % of the total

variation, followed by salinity and mud with 5% and 0.3% of the total variation

respectively. All three variables were retained, therefore no correlation. In the

distance based linear modelling (DISTLM) test, the sequential test revealed that the

trees from the mangrove forests were correlated with emersion time, salinity and

substrate type (Table 2.10). The most parsimonious model (AIC = 596. 18, R-Sq =

0.48) for the mangrove forests explains 47.7% of the variation, with emersion time

explaining 40 % followed salinity explaining 4%, and then mud explaining 3%.

Table 2.10 Distance-based linear modelling (DISTLM). Sequential test results relating the

mangrove trees within each site (abundance data) with emersion time, salinity and mud recorded

from each of the mangrove forests.

P- Predictor variables AIC SS (trace) Pseudo-F value Proportion Accumulative df

Emersion 603.3 66543 56.33 < 0.001 0.4 0.40429 83

Salinity 598.81 7200.9 6.4997 < 0.01 0.04 0.44804 82

Mud 596.18 4824.2 4.5426 < 0.05 0.03 0.47735 81

2.5 Discussion

Inland forests, such as the Sombano and Gili forests and the coastline-

dominating Langira and One Onitu forests which are linked to other ecosystems have

rarely been studied within Indonesia (Tomascik et al., 1997). This type of

topography is not common to Indonesia; however the forests do contain mangrove-

tree species combined with zonation patterns often reported for Southeast Asia

(Tomascik et al., 1997; Giesen et al., 2006; Cragg and Hendy, 2010). 59

Zonation patterns of trees were found in Sombano, with Heritiera littoralis found only at the strandline and more salt-tolerant species, Bruguiera and

Excoecaria found further towards the sea. The mix of Bruguiera gymnorhiza and

Excoecaria agallocha trees may be due to the intolerances of environmental variables. Both species are commonly found in the same areas within mangrove forests (Duke, 1992). Very subtle changes of tidal emersion were found in Sombano

(Figure 2.2), therefore waterlogging of mangrove root-tissues did not create distinct zonation patterns between the Excoecaria and Bruguiera trees. As differences of immersion time; the waterlogging of mangrove roots may contribute to species- specific zonation patterns. Bruguiera gymnorhiza is more tolerant to immersion than

Excoecaria agallocha, which is more tolerant than the Heritiera littoralis trees. The differences in tolerance to waterlogging between different mangrove trees are related to their anatomical features of cortex root-tissues, epidermis and hypodermis (Pi et al., 2009). The Sombano substratum consisted of organic-rich mud and the forest was established in the high intertidal zone. Those factors did not affect the zonation patterns of plants within Sombano. The distribution of trees within Sombano is controlled by salinity, which increased in concentration with distance from the landward edge.

The bank of mixed saplings within Sombano is an indication of young stands.

Older trees have greater girth measurements and are less densely distributed (Alongi,

2002). The sparser, but large Xylocarpus spp. in Sombano may contribute to increases of the total above-ground biomass with increasing age. A similar pattern was found in mangrove forests in Vietnam - where stands also become less dense, due to increases in tree girths (Alongi, 2002).

The low salinities and long emersion times recorded at the strandline within 60

Sombano explains the limited distribution of Nypa fruticans and Heritiera littoralis.

The presence of N. fruticans is an indicator of freshwater environments, and is often found adjacent to H. littoralis stands (Tomascik et al., 1997) and these palms do not tolerate salinities above fifteen PSU (Blasco et al., 1996; Alongi, 2007).

Furthermore, N. fruticans does not tolerate more than three hours of inundation per day (Blasco et al., 1996), which the same thing is seen from this study as N. fruticans were found in areas from Sombano with less than three hours of immersion and salinities less than fifteen PSU.

Adjacent and seaward to the Sombano forest is the Langira mangrove forest.

Emersion time was found to influence the zonation patterns within that forest.

Rhizophora apiculata was established in areas of the forest with increased emersion times. However, R. stylosa and S. alba were more common in areas less frequently inundated. The Langira forest is established on a large coral-rock plateau. Hard coralline substrata are often dominated by R. stylosa (Tomascik et al., 1997).

However, hard substrata may prevent Rhizophora propagules from embedding and therefore establishing. Rhizophora propagules need to be anchored in to the sediment; this is an important factor for their survival (Kathiresan and Bingham,

2001). Thus, the fossil-coral substratum in Langira may impede and reduce the successful establishment of propagules, which may explain the patchy canopy within

Langira. Sonneratia alba were found only at the outermost seaward-edges of the forest, which had the highest salinity levels. S. alba commonly grows on the outermost seaward margins, especially in sandy and rocky areas that are fully marine

(Tomascik et al., 1997; Giesen et al., 2006).

Within Langira, Xylocarpus spp. were found only in areas with the longest emersion time and lowest recorded salinities. These environmental factors determine 61 the distributions of Xylocarpus spp. Within other southeast Asian mangrove forests,

Xylocarpus spp. were also reported in areas only inundated by high-tides, in the landward regions and in areas of lower salinities (Tomascik et al., 1997; Giesen et al., 2006). Although individual Xylocarpus spp. within the Langira forest are very large, those trees were not dominant due to their very low abundance within the forest.

Higher mangrove tree densities have reduced basal areas (Alongi, 2002). The same thing was seen with this study; the Sombano forest had the largest density of B. gymnorhiza trees, but those trees had small girths. Langira and Loho had the lowest densities of B. gymnorhiza, but those trees were larger in comparison with Sombano

- which highlights the importance of these larger mature trees. Less distributed, larger trees are not only an indication of age, but they also significantly contribute to an increase of biomass for the forest (Along, 2002). The mean recorded circumference for B. gymnorhiza in Langira and Loho was greater than those reported from Sombano. However, many of the larger mature B. gymnorhiza trees from Sombano had been felled. Thus, much of the Sombano tree biomass had been removed.

Langira had the lowest recorded densities of R. stylosa, but a lower basal area for this species. The lower measurements of mean girth of the R. stylosa stands recorded from Langira may be due to the result of the heavy harvesting pressures imposed upon the forest (Walters, 2005a; Walters et al., 2008). The reduced structural complexity of the R. stylosa stands, low densities and low canopy height within Langira, may also be attributed to forest age (Alongi, 2002), as tree densities decline, with age. However, some trees within Langira were not harvested – such as the large Xylocarpus spp. 62

Within each of the mangrove forests the basal areas of the trees differed. The differences of tree girths may be attributed to species densities, changes of environmental conditions and harvesting pressures (Alongi, 2002; Giesen et al.,

2006; Walters et al., 2008). The differences of tree-girths recorded between the forests suggest that the total dominance and living biomass of the tree species within each forest also differs.

Using basal-area measurements as an indication of tree-species dominance and of the quantity of living biomass within the forests of the WMP, comparisons of woody biomass can be made with other forests, using allometric equations

(Komiyama et al., 2008). The numerous large trees from Sombano provided estimates of forest biomass that were similar to the extensive mangrove forests reported from other areas in Indonesia, Malaysia and Papua New Guinea, which had biomass values ranging from 430 to 600 tonnes per hectare (Komiyama et al., 2008;

Alongi, 2009). Values reported from Langira were lower than Sombano, but higher than estimates from the mangrove forests on Derawa. The biomass of trees within

Langira is similar to those reported from other Indonesian mangrove forests that range up to 170 tonnes per hectare for Rhizophora-dominated forests (Komiyama et al., 2008). The largest contributors of living biomass within Langira are the

Xylocarpus spp. demonstrating that even a very few trees can significantly contribute to the total above-ground forest biomass (Ong et al., 1995; Alongi, 2002). The lowest estimates of forests biomass were recorded from the mangrove forests on Derawa.

The Kaluku, Loho, One Onitu and the Gili forests are similar to the ranges of mangrove forest biomass reported from Florida and Kenya, where biomass estimates range between 8 – 20 tonnes per hectare (Lugo and Snedaker, 1974; Kairo et al.,

2009). 63

Mangrove forests develop best in areas associated with high rainfall and freshwater input (Lira-Medeiros et al., 2010), such as Papua New Guinea with a mean annual rainfall of 3,485 mm year-1 (Spalding et al., 1997) where impressive thirty to forty metre tall Bruguiera and Rhizophora trees develop (Alongi, 2007).

However, the mean annual rainfall for Sulawesi is much lower - 1,525 mm year-1

(Dechert et al., 2005). Within the WMP the mean annual rainfall varies between 850

– 3300 mm year-1 (Crabbe and Smith, 2005). This may explain the lower tree heights and shrubby appearance of the mangrove trees within the seaward mangrove forests of the WMP.

Mangrove productivity, biomass and growth, increases with decreasing salinities combined with increases of nutrient input (Lugo, 1980; Clough, 1984;

Sherman et al., 2003) - such as Sombano and the landward regions of Langira. The

Sombano forest has a thick covering of organic-rich mud and the lowest salinities of the forests in this study. Thus, these environmental factors may explain the relatively high dominance, diversity and evenness of the trees within Sombano when compared with all the other forests studied in this investigation. It is known that mangrove forests attain their greatest biomass, diversities, canopy heights and productivities in areas where salinities are between ten and twenty-five PSU (Robertson, 1990;

Tomascik et al., 1997; Joshi and Ghose, 2003; Sherman et al., 2003). The Sombano forest contains trees that reach their greatest heights compared to the other mangrove forests researched in the WMP.

Many Rhizophora and Bruguiera trees in Langira have been heavily cut for fuel wood. Removal of wood from a mangrove forest will reduce the habitat heterogeneity (Gratwicke and Speight, 2005). In general, there are many environmental implications from habitat loss (Granek and Ruttenburg, 2008); 64 mangrove forests protect coastlines from storm damage (Walters, 2005b; Walters et al., 2008) – and they also serve as vital nurseries and feeding grounds for many invertebrates and fish species (Kathiresan and Bingham, 2001). Although the plant diversity is high, wood-harvesting is common practice within Sombano and in the adjacent seaward Langira forest. Unsustainable harvesting of wood has long-term detrimental effects (Bosire et al., 2006), and the intensity and frequency of disturbance can have permanent effects on forest structure and tree composition

(Hauff et al. 2006).Harvesting will lead to a change in ecosystem function due to increasing canopy gaps. Gaps within the forest canopy lead to abiotic changes of light penetration and temperature, combined with increases of cyanobacteria, which lead to a reduction of grazing animals (Granek and Ruttenburg, 2008) - such as

Terebralia sulcata, which were often found in the forests.

The Sombano and Gili forests have a high level of plant-diversity and evenness - these forests are senescent, a type of forest that has rarely been studied

(Duke, 2001) and both are established inland and separated from direct tidal inundation. Senescent mangrove forests may promote higher tree diversity, when canopy gaps form due to falls of old trees the disturbance creates new opportunity for other trees to establish (Duke, 2001).

Sombano and Gili demonstrated evidence of the larger emergent trees becoming senescent. Up-rooted fallen trunks of large trees, and canopy gaps were common in both forests. Determining the age of mangrove trees is challenging due to the lack of growth rings. Nonetheless, these large trees have been estimated to be very old (Cragg and Hendy, 2010). Another indicator of senescent forests is pedogenesis analyses. The older the forest, the greater the organic content is sequestered within the sediments; as the particulate organic carbon derived from 65 large woody detritus is partially retained. Organic-rich mangrove forest mud contains ligno-cellulosic debris derived from decaying higher plants, and has ten-times the organic content of sediments of less mature forests (Ashton and Macintosh, 2002;

Marchand et al., 2003). Sombano and Gili have organic-rich mud substrata, and both forests contain the tallest trees. The sediment composition observed within all the forests accept Sombano and Gili consisted of a higher proportion of calcareous material, such as sand and shell fragments. Mangrove forests lose significant amounts of dissolved organic carbon to coastal waters (Kristensen et al., 2008).

Thus, the less organic debris within the sediments may provide reasons for the smaller trees recorded within the seaward forests of this study.

The estimates of living-biomass of mangrove trees reported from this study are comparable to those from other mangrove forests and these trees are a source of litter for the mangrove forests in the WMP, but forest harvesting and the removal of forests-products may impact the faunal communities of the WMP (Cragg and Hendy,

2010). Many animals; fish and invertebrates rely upon large woody debris to avoid predation, as the complexity of fallen wood offers protection for animals (Wright and

Flecker, 2004). The reduction of the mangrove forest complexity due to the harvesting pressures imposed upon these already impacted mangrove forests - and the loss of old trees within the WMP is accelerating. These impacts may have further consequences as the available dead wood within these forests may be important habitat for many animals (Cragg and Hendy, 2010).

Chapter 3

INVESTIGATION OF WOOD BIODEGRADATION IN MANGROVE ECOSYSTEMS OF THE WMP

3.1 Introduction Detrital breakdown and nutrient recycling is a fundamental ecological process for every natural ecosystem. This process occurs in terrestrial and aquatic environments – whether coral reefs (Hata et al., 2002), freshwater streams (Karlsson et al., 2005; Greenwood et al., 2007), or tropical rain forests (Cerri et al., 2007).

Over fifty percent of the World’s organic carbon is sequestered in to terrestrial ecosystems (Malhi et al., 2002), with the remaining carbon taken up by marine systems (Muller-Karger et al., 2005). Within marine intertidal areas, mangrove forests sequester much of the carbon and are considered to be carbon-sinks (Donato et al., 2011). However, nutrient recycling in mangrove forests is complex as they are sites where terrestrial and marine ecosystems directly interact.

In mangrove ecosystems, carbon is fixed via photosynthesis to form various plant tissues such as large woody trunks, roots and leaves (Ong et al., 1995). The mangrove organic material is then processed in the direct transfer of energy to higher trophic levels. The biodegradation of organic material is processed in different ways via saprophytic decay (Sarma and Vittal, 2001; Ananda and Sridhar, 2004) and mechanical breakdown of plant material (Robertson and Daniel, 1989a; Filho et al.,

2008). Thus, litter within mangrove forest environments, broken down in different ways serves as the base for intricate food webs (Benner and Hodson, 1985). The principal component of woody plants and dead plant material is lignocellulose, which is the most abundant biomass on earth (Eggert et al., 1996; Kalmis et al., 67

2008). Thus, the biodegradation of large woody debris (LWD) in mangrove forests is an essential contribution to the global carbon cycle (Weedon et al., 2009; Donato et al., 2011). The flux of carbon in an Australian mangrove forest from the gradual degradation of LWD is reported to be similar in quantity to that of leaf litter consumption by crabs (Robertson and Daniel, 1989a). Much of the mangrove- derived carbon is assimilated within nearby marine ecosystems, which are typically oligotrophic. A study of mangrove-derived organic matter outwelling in Panama demonstrated that connectivity is important as many sessile reef invertebrates rely upon mangrove organic matter (Granek et al., 2009).

Wood constitutes between 20-50% of the net primary production of mangrove systems (Bunt et al., 1979, Robertson, 1990), therefore fallen wood provides a large proportion of organic carbon to a mangrove ecosystem (Donato et al., 2011). LWD when broke down serves to enhance the longterm flow of nutrients in the mangrove forest ecosystem (Benner and Hodson, 1985; Krauss et al., 2005). A study within a tropical Australian mangrove forest revealed that LWD will remain within the ecosystem for more than fifteen years (Robertson and Daniel, 1989a).

LWD also promotes forest productivity by enhancing soil pedogenesis and providing habitat for germinating seeds (Allen et al., 2000; Krauss et al., 2005) and animals

(Cragg and Hendy, 2010), and may even contribute to the mangrove forest nitrogen budget when processed (Robertson and Daniel, 1989a). Thus, LWD is a key resource within mangrove ecosystems.

Previous studies of mangrove forest litter degradation focused on leaves and twigs (Lambert et al., 1980; Robertson, 1985). Litter input from leaves, reproductive parts, small branches and twigs, and litter processing has been studied from both the

Caribbean (Feller and Mathis, 1997; Brooks and Bell, 2002) and Australian 68 mangroves (Robertson, 1985; Robertson and Daniel, 1989a). The animals that process such material include coleopterans (Feller, 2002; Sousa et al., 2003), marine intertidal crustaceans such as sphaeromatid isopods (Brooks and Bell, 2002;

Svavarsson et al., 2002), sesarmid crabs (Robertson, 1985; Schories et al., 2003) and pulmonate gastropods (Proffitt and Devlin, 2005). However, there is a paucity of knowledge regarding the processes affecting the breakdown of LWD with only a few studies that provide rates of mangrove wood decay. The loss of LWD dry mass in an

Australian mangrove forest ranged from 10 – 97 grams dry weight per year, per metre squared (Robertson and Daniel, 1989a). However, leaf litter is much more labile than LWD and tissue loss is greater. In Caribbean mangrove forests, 27% of leaf litter tissues remained after eight months, and 50% of LWD tissues still remained after eighteen months (Middleton and McKee, 2001). However, the rates of wood degradation vary over different tidal zones within mangrove ecosystems, due to the variety of different wood degrading organisms found at different tidal zones

(Kohlmeyer et al., 1995).

The biodegrading organisms which process LWD in mangrove forests are the direct link that enhances nutrient recycling (Kohlmeyer et al., 1995; Babu et al.,

2010). This consortium of organisms forms a diverse group from the terrestrial to the marine regions of the forests - such as the basidiomycetes (Fryar et al., 2004;

Lonsdale et al., 2008; Ravikumar et al., 2009), coleopterans (Feller and Mathis,

1997; Feller and Mckee, 1999; Mckeon and Feller, 2004), termites (Leponce et al.,

1995) and teredinid bivalves (Robertson and Daniel, 1989a; Filho et al., 2008).

Saprophytic degradation of LWD is slow (Benner and Hodson, 1995), and is caused by basidiomycetes which occur above the hightide mark in mangrove forests

(Kohlmeyer et al., 1995). There are three types of decay recognised in fungi; 69 ascomycete fungi create soft rot and decay the cellulose and hemicellulose (Bucher et al., 2004). Soft rot is usually restricted to the surface of wood and results in minimal lignin degradation (Bucher et al., 2004). The second form of decay is white rot, which facilitate a rapid decay of all the components of wood. This type of decay is usually caused by basidiomycetes, which affect even the most recalcitrant component of wood, lignin. Thus, the basidiomycetes are essential for the decay of

LWD in the supratidal areas in the mangrove forests. The third type of decay is brown rot, which causes the rapid decay of cellulose and hemicellulose. This type of decay is not found in wet or waterlogged wood. Thus, brown rot is not found in aquatic environments (Bucher et al., 2004).

However, in the mid- to low- intertidal regions of mangrove forests, biodegradation is accelerated when LWD is exposed to wood-borers, in particular animals belonging to the family Teredinidae (Robertson and Daniel, 1989a;

Middleton and Mckee, 2001; Filho et al., 2008). Thus, LWD will break down at different rates, by a variety of different pathways. This variation will depend on the plant species composition within the mangrove forests and the distribution of wood- decaying and wood-boring animals (Robertson, 1990; Middleton and Mckee, 2001).

The Teredinid boring activity is particularly important in forests where there is a high level of woody biomass consisting of fallen trunks (Robertson, 1990;

Kohlmeyer et al., 1995).Teredinids process fallen logs into fine fragments of faecal material that contribute to nutrient recycling and reducing the build-up of LWD in mangrove ecosystems (Filho et al., 2008). Teredinids cannot tolerate regular prolonged emersion; therefore the distributions of LWD along a tidal gradient in mangrove forests will influence the activity of the Teredinidae (Robertson, 1990). In the mid- to low- intertidal zones of a Rhizophora-dominated Australian mangrove 70 forest, approximately fifty percent of the original woody mass of large fallen logs had been processed by the Teredinidae over two years, whereas, in the high- intertidal only five percent of the original mass of fallen logs had been lost over two years, as the teredinids were absent (Robertson, 1990).

This study describes the biodegradation processes from the terrestrial regions of the mangrove forest to the marine regions of the forest by linking the degradation processes to the environmental gradients within the forests. Biodegrading organisms associated with the degradation and recycling of LWD each have a specific role within mangrove ecosystems. The aims of this study were to identify the organisms that break down woody detritus within mangrove forests and to determine the environmental factors that affect their distribution within those forests. This study also aimed at comparing the rate of LWD fragmentation caused by teredinid tunneling in intertidal mangrove forests and at nearby coral reef sites.

LWD biomass in the mangrove forests was measured in a very precise way to determine the volume of fallen wood available to the biodegrading organisms. All organisms that break down LWD within each mangrove forest were identified.

Salinity and emersion were measured to determine if the distribution of wood- degrading organisms were influenced by those variables. Two distinct forests that had terrestrial and marine biodegrading organisms were chosen to investigate the biodegradation boundary of LWD determined by the sharp change between terrestrial and marine wood-degrading organisms. Panels were also used to determine both the environmental variables and the level of wood degradation by teredinids.

This work tested the hypothesis that the biodegrading communities in the mangrove forests, 1. change with increasing distance from the land and, 2. are influenced by the environmental factors of salinity and emersion time. 71

3.2 Materials and methods

3.2.1 Measurement of wood-loss in panels

A total of sixty-five wooden panels, cut from Bruguiera gymnorhiza planks were placed within the Sombano and Langira forests and Kal 1 reef site to determine the dry-weight losses of wood from teredinid activity (Figure 3.1 c). Panels had a volume of 275 cm3 and measured. All panels were exposed for twelve months. The dry-weight loss of wood was determined by oven drying the panels for 48 hours at

60°C. The percentage of the sectional area of the panel containing teredinid tunnels was calculated using the digital analysis package ImageTool Version 3.00 (The

University of Texas Health Science Centre at San Antonio).

Panels (n = 5) were placed 0, 20, 100, 240, 380 and 500 metres from the landward edge within the Sombano forest. The panels within the Langira forest were placed at 720, 820, 940, 980 and 1070 metres from the Sombano landward edge.

Panels on the Kal 1 reef site were placed at 10 metres depth, 1500 metres from the

Sombano landward edge (Figure 3.1 c).

To determine wood-loss from teredinids, 25 panels within Langira, 60 panels within Galua and 20 panels each in Kaluku, Loho and One Onitu (Figure 3. 1 a) were placed. Twenty panels were positioned at a depth of 10 metres from three sub-tidal reef sites; in the Kal 1, Pak Kasim’s and the Ridge 1 reefs (Figure 3.1 c) in July 2010 and removed in July 2011. The percentage of wood fragmentation attributed by the activity of the Teredinidae was estimated from the panels from the mangrove forests and each reef site. All panels were cut longitudinally to determine the level of wood fragmentation attributed by teredinid activity. The percentage of the sectional area of each panel represented by tunnels and the area of panel remaining was estimated. 72

The panels were positioned in areas of each mangrove forest where natural

LWD had been attacked by teredinids. Panels within the Derawa forests, Kaluku,

Loho and One Onitu were placed in areas inundated by all high tides. All teredinid pallets were removed after 12 months exposure by carefully splitting each panel open. Each teredinid pallet was used to identify the different teredinid species using the teredinid identification key of Turner (1971).

3.2.2 The measurement and collection of dead wood and arbitrary scale classification and wood-loss from teredinid attack

Five transects extending from the strandline and ending at the seaward fringe were established in the Sombano and Langira forests on Kaledupa Island (see chapter

2 methodologies). In Sombano a total of 22, 20 x 20 m plots were used to measure the biomass of LWD and in Langira 19, 20 x 20 m plots were used to measure the biomass of LWD. All pieces of LWD, defined as being more than 2 cm in diameter, were measured to estimate their volume. 73

a b

Figure 3.1 Sites within Kaledupa and Derawa Island used for sampling the variety of

biodegrading organisms and to reveal the environmental factors that determine the

biodegradation of LWD. a, map of Kaledupa demonstrating the geographical distribution of

each site; b, mangrove forests have been colour-coded by their site-specific biodegradation

process (terrestrial = brown, marine = blue) and height above sea level (mean ± SE). The

boundary between the terrestrial and marine biodegradation is indicated by the red line within

the Galua and Laulua forests; and c, the three reef sites used in this study. The red dotted line

is the transect used to determine wood-loss.

The height above sea level of each individual piece of LWD was measured.

The high tide marks were recorded on mangrove trees using high visibility string. At low tide, the distance from the substratum to the mark on the trees were measured and then subtracted from the marks on the trees to establish height above sea level. 74

Emersion times for the LWD were estimated by relating their tidal heights to data in regional tide tables.

A total of 168 LWD samples were collected at low tide from the five transects within the Sombano and Langira forests. Sombano is 0 – 500 metres from the landward edge and Langira is 700 – 1100 metres from the landward edge of the

Sombano strandline. These forests are separated by a 200 metre wide fossil-coral plateau. Levels on the scale of wood biodegradation detailed in Table 3.1 were assigned to each LWD item to define the levels of biodegradation in the Sombano and Langira forests. Wood-degrading organisms processing each LWD item were identified to their lowest achievable taxonomic classification.

Table 3.1 A scale used to describe level of wood degradation. The level of degradation

imposed upon the LWD in the Sombano and Langira mangrove forests.

Level Condition and appearance of specimen

0 No visible signs of deterioration or degradation from both inside and out of the woody debris.

1 The tough outer covering of bark reveals signs of flakiness. Some initial epiphytic fungal fruiting bodies are visible in areas.

2 Areas of bark can be pulled off by hand and is soft in areas. Fungal fruiting bodies increase. Some exposure of sapwood, with initial softening.

3 Bark is very flaky, and in most areas can be pulled off with relative ease. Sapwood is easily penetrated to a depth of 2 – 4 cm using a knife. Sapwood begins to deteriorate revealing holes – sometimes with fungal growths.

4 Bark, if present is very soft/flaky. The sapwood has numerous holes, which is often soft to the heartwood.

5 The bark, when visible, crumbles easily. The woody debris can appear to be ‘collapsed’. Both sapwood and heartwood are very soft. Often, under story plants grow through the woody debris. In addition, basidiomycete activity upon the detrital material is common. 1 *IWH developed the decay-scale from field observations of fallen trees within the mangroves from

this study.

3.2.3 Wood degrading organisms and environmental factors in the Galua and Laulua mangrove forests Two mangrove forests, Galua and Laulua with no disruption of environmental gradients by a fossil-coral plateau were chosen to investigate the environmental factors that create the boundary change between terrestrial and marine wood-degrading organisms (Figure 3.1 b). Eight, 10 m wide belt transects were used to study the degradation boundary in the Galua forest and three, 30 m wide belt transects were used to study the degradation boundary for the Laulua forest to determine the biodegradation processes. In both mangrove forests the transects began in areas dominated by terrestrial wood-degrading organisms, well above the high intertidal zones and ended in areas where marine wood-degrading organisms dominated at the seaward edges. A total of 583 and 332 pieces of LWD were examined in the Galua and Laulua forests respectively. The volume of LWD items were calculated for every piece investigated, and all wood-degrading organisms from the LWD samples were identified. Salinities were measured using a Bellingham and

Stanley E-Line Aquatic hand-held Refractometer. Emersion times for the LWD items in each transect was calculated using the results obtained from the shore heights above sea level of from each transect and the local tide tables for Kaledupa.

Identification of marine and terrestrial woody detrital decomposers

Coleopteran larvae, worker and soldier termites, and pallets (the diagnostic calcareous hard parts of teredinids) were stored in absolute ethanol and identified in the laboratory. Digital images of all basidiomycete fruiting bodies and termites, focussing on morphology and colour characteristics were used for identification 76 within the Galua, Laulua and Sombano forests using an Olympus E-510 SLR with a

50mm Macro-lens.

3.3 Data analysis

Mean values of the volumes of LWD between the sites, and at different tidal heights were tested for statistical significance using an ANOVA, with sites and tidal height as factors. A 1-way ANOVA was used to test for differences between the dry weights of panels and distance from the strandline. A regression analysis was used to determine differences between the percentage of the sectional area represented by teredinid tunnels and the level of fragmentation imposed to each panel in the forests and reefs. The non-parametric Kruskal-Wallis test was used to test for differences of wood-decay. All percent data were normalized using Arcsin transformations. A

Tukey’s pairwise comparison test separated values into statistically distinct subsets in all ANOVAs using MINITAB (MINITAB Inc, version 13.20).

Distance-based linear modeling (DISTLM) was employed to verify relationships between different wood-degrading organisms and emersion time, distance from the land (DFL) and salinity. DISTLM produces a sequential test, which assess the variation each predictor (environmental variable) has on its own, and a conditional test, which explains the variation of all the variables (McArdle and

Anderson, 2001). The most parsimonious model for predicting the distribution of biodegrading organisms affected by emersion time, distance from the land, and salinity was identified using the Akaike information criterion (AIC). AIC ranks models from all possible combinations of the environmental variables. Distance- based redundancy analyses (dbRDA) were used for visualizing the results as an ordination, constrained to linear combinations of the predictor variables. The

DISTLM was based on presence/absence data with 4999 permutations and a 77

Sorensons transformation. DISTLM was performed using PRIMER+PERMANOVA version 6.1.11.

3.4 Results

3.4.1 Wood degradation within the landward, Sombano and seaward, Langira forests

The levels of decay estimated from 168 LWD samples within the Sombano forest (Figure 3. 2 a) had the greatest percentage of the highest categories of degradation - levels 4 and 5, and the lowest percentage of non-degraded LWD - level zero. A decline in degradation at 400-500 metres was determined by the reduced percentage frequency of the level 5 categories and increases of non-decayed level zero LWD. Furthermore, in the Langira forest beyond the fossil-coral plateau at 700 metres from the landward edge, the decay levels of the LWD were very low, evidenced by the high percentage of the level zero ratings combined with no level 5 ratings. The level of degradation is significantly lower with further distances from the land (Kruskal-Wallis test, degradation rating vs. distance from land, P = < 0.001).

3.4.2 Sombano, Langira & Kal 1 panel analysis – wood loss

The percentage dry-weight loss of wood from the panels within the Sombano and Langira forests and the Kal 1 reef site showed significant differences with increasing distance from land (Figure 3.2 b). More wood was lost with increasing distances from the land. No weight losses were detected from panels exposed within

Sombano. With increasing distance from the strandline and decreasing emersion time, panels lost > 50 % of their weight.

3.4.3 The biodegrading organisms found on LWD

The basidiomycetes

Basidiomycetes were found only in the Sombano and Galua forests (Table

3.2). Both forests had distinct basidiomycete mycoflora, with the exception only of

Ganoderma sp. and Pycnoporus sanguineus, which were found in both forests

(Figure 3.3). The basidiomycetes mycoflora consisted of eight families and 13 species. Auricularia cornea, Favolus spp., Oudimansiella canari, Panaelus spp. and

Schizophyllum commune were found only in Galua even though their host substrata, fallen Xylocarpus granatum trees, were common to both forests. Similarly, Lenzites acuta were found only in Sombano even though its host, Bruguiera gymnorhiza is found in both forests.

79 a

100 Level 5 90 80 Level 4 70 Level 3 60 50 Level 2 40 30 Level 1 20 FOSSIL CORAL Level 0 Frequency of LWD decaycategory(%) 10 PLATEAU 0 b Sombano Langira Emersion time (hours) 120 25 a a a a a a bcd 100

80 15 60 d b cd cd 10 40 d

% % Dry weight(grams) FOSSIL 5 Emersion time (hours) 20 CORAL PLATEAU 0 0 0 20 100 240 380 500 720 820 940 980 1070 1500

Distance from strandline (m)

Figure 3.2 Wood decay levels and wood loss in areas of the mangrove forests with different emersion times with increasing distance from the land a, The frequency (%)

of decay levels assigned to the LWD (n = 168) within Sombano and Langira (refer to

Table 3.1); b, The percentage dry-weight loss of wooden panels (mean ± SE, 1-way

ANOVA, distance from land vs. percentage dry weight; F11, 48 = 178.75, p = < 0.001)

exposed for 12 months deployed in different areas of the Sombano and Langira forests

exposed to different emersion times (dashed line).

Table 3.2 The biodegrading organisms recorded from the forests within the Wakatobi

Marine Park (WMP). Wood on which organisms found: Xg = Xylocarpus granatum, Ea =

Excoecaria agallocha, Bg = Bruguiera gymnorhiza and Ct = Ceriops tagal. Sites: G = Galua, S

= Sombano, Lau = Laulua, L = Langira, On = One Onitu, Lo = Loho and Ka = Kaluku.

Common name Family Species Site Substratum Bracket fungi (basidiomycetes) Auriculariaceae Auricularia cornea Ehrenb.:Fr. G Xg Entolomataceae Entoloma sp (Fr.) P. Kumm. S Ea

Ganodermataceae Ganoderma sp P. Karst S,G Xg

Panaeoleae Panaeolus sp (Fries) Quel. G Xg

Physalacriaceae Oudimansiella canari (Jungh.) Hohnel G Xg

Polyporaceae Favolus sp P. Beauv. G Xg

Lenzites acuta Berk. S. Bg

Microporus xanthopus P. Beauv S Ea

Pycnoporus sanguineus (L.) Murrill S,G Xg

Trametes elegans (Fr.) Fr. S Ea, Bg

Trametes versicolour (L.:Fr) Pilat G Ct

Schizophyllaceae Schizophyllum commune Fr. G Xg

Stereaceae Stereum ostrea (Blume and T. Nees) Fr. G Bg

Beetle (coleopteran) larvae Cerambycidae S, G

Elateridae S

Oedemeridae S, G, Lau, L

Tenebrionidae S, G

Termites Kalotermitidae Cryptotermes Ga, Lau In decaying wood Rhinotermitidae Coptotermes Ga, Lau Arboreal

Prorhinotermes Ga Arboreal

Termitidae Microcerotermes Ga, Lau In decaying wood

Nasutitermes Ga, Lau Arboreal

Limnoriids Limnoriidae Limnoria insulae Menzies, 1957 L, Lo, O,

L. pfefferi Stebbing, 1904 L, Lo, O,

L. sellifera Cookson et al., 2012 L, Lo, O,

L. unicornis Menzies, 1957 Lo,

Shipworms (teredinids) Teredinidae Bactronothorus thoracites Gould, 1856 La, Ga

Dicyathifer caroli Iredale, 1936 Ga

Lyrodus massa Lamy, 1923 La, On, Lo, Ka

Spathoteredo obtusa Sivickis, 1928 La, On, Ka, Lo

Teredo fulleri Clapp, 1923 La, Lo, On

T. furcifera von Martens, 1894 La, On, Ka, Lo

T. mindanensis Bartsch, 1923 La, Ga, On, Lo, Ka

Ganoderma sp. and Pycnoporus sanguineus were recorded from both forests, with both basidiomycetes degrading fallen Xylocarpus granatum trees. All recorded basidiomycetes within Sombano and Galua had host-specific substrata.

Schizophyllum commune was ubiquitous in the landward regions in the Galua forest, on dead Xylocarpus granatum trees only. However, one exception was observed with

Trametes elegans which was found only in the Sombano forest, on Excoecaria agallocha and Bruguiera gymnorhiza trees.

a b

c d

Figure 3.3 Basidiomycetes fruiting bodies decaying LWD. a, Pycnoporus sanguineus on a fallen Xylocarpus sp. tree in the Galua forest; b, the fruiting bodies of Microporus xanthopus on an Excoecaria agallocha snag in Sombano; c, Lenzites acuta decaying a fallen Bruguiera gymnorhiza tree in Sombano; and d, Trametes elegans on a chopped E. agallocha tree in

Sombano. 82

Coleopteran larvae Larvae from four coleopteran families were found; Cerambycidae, Elateridae,

Oedemeridae and Tenebrionidae (Figure 3.4 and Table 3.2). The Sombano forest contains all four coleopteran families. The Galua forest had three coleopteran families. The Langira and Laulua forests had one coleopteran family, the

Oedemeridae. In addition, the Oedemeridae were found within all sites except

Sombano.

Within Sombano, many Excoecaria agallocha trees had 3 – 4 cm diameter circular blisters created by Aeolesthes holosericeus belonging to the family

Cerambycidae. A. holosericeus caused damage within many live and dead E. agallocha trees in the Sombano forest.

a b

c d

Figure 3.4 Dorsal views of coleopteran larvae removed from LWD. a, a beetle larva

in the family Cerambycidae from the Galua forest; b, an Elatrid larva from Sombano; c,

a beetle larva in the family Oedemeridae from the Laulua forest; and d, a beetle larva

from to the family Tenebrionidae from the Sombano forest. 83

Termites There were three termite families consisting of five genera in the Galua and

Laulua forests (Figure 3.5 and Table 3. 2). Of the five genera, Cryptotermes and

Microcerotermes were found in decaying wood from both Galua and Laulua.

Prorhinotermes, Coptotermes and Nasutitermes were found in arboreal nests above spring high tide.

a b c

d e

Figure 3.5 Dorsal views of soldier termites removed from LWD. a, frequently found

in decaying mangrove wood, Cryptotermes from Galua and Laulua; b, Prorhinotermes

found well above the hightide mark in Galua; c, Microcerotermes found in Galua and

Laulua decaying wood; d, Coptotermes common to Galua and Laulua (dorsal and ventral view); and e, Nasutitermes found in arboreal nest in the landward/and supratidal areas of

Galua and Laulua.

Limnoridae

Superficial biodegradation of the LWD identified as small ridges on the wood surface were created by limnoriid isopods, including a new species identified from the forests of Derawa - Limnoria sellifera (Cookson et al., 2012). In total, four limnoriid species were found, Limnoria insulae, L. pfefferi, L. unicornis, and L. sellifera (Figure 3.6). No limnoriids were recorded from the landward forests,

Sombano and Gili. In addition, L. insulae, L. pfefferi and L. sellifera were found in the Langira, Loho and One Onitu forest sites (Table 3.2). L. unicornis was found only in the Loho forest.

a b

c d

Figure 3.6 SEM images of limnoriid isopods removed from LWD. a, lateral view of L.

Teredinidaeunicornis; b, a new described species of limnoriid, L. sellifera with its unusual saddle- shaped pleotelson exposed; c, the pleotelson of L. insulae; and d, a ventral view of L.

pfefferi.

Five species belonging to the family Teredinidae were found in experimental

panels (Figure 3.7): Teredo furcifera, T. fulleri, T. mindanensis, Lyrodus massa and

Spathoteredo obtusa. Teredinids were recorded from all forest sites except Sombano

(Table 3.2). Bactronophorus thoracites and Dicyathifer manni were only found in

LWD.

a b

c d

Figure 3.7 Teredinid pallets - diagnostic features of teredinids - removed from panels in the mangrove forests of the WMP. Pallets belonging to a, Teredo mindanesis; b, Lyrodus massa; c, T. fulleri ; d, T. furcifera; and e, Spathoteredo obtusa.

Polypore fungi, lignocellulose-degrading basidiomycetes are ubiquitous in

Sombano. All basidiomycetes were found on decaying fallen logs above the high tide water mark. Other organisms including termites, Cryptotermes and Nasutitermes are also common within Sombano, with extensive termite galleries frequently found inside the LWD. Coleopteran larvae are also very common within decaying wood in

Sombano.

Beyond the fossil-coral plateau that separates Sombano from Langira, in the seaward region of the forest distinct changes of the biodegradation of LWD were found. These changes were evident by the rapid reduction of saprophytic decay with increasing distance from the landward edge. The terrestrial biodegrading organisms; basidiomycetes, termites and coleopteran larvae are not the main decomposers of wood in the mid- to low- areas of the forests.

The Galua and Laulua forests contained the same wood-degrading organisms found in LWD from Sombano and Langira. The coleopteran and termite communities in the Laulua forest had similar distributions of those recorded from

Galua. Neither coleopteran larvae nor termites were found beyond 170 metres from the land in the Galua forest. Combinations of termites and coleopteran larvae, termites and basidiomycetes, and coleopteran larvae and basidiomycetes were found together processing twenty-two items of fallen wood in the high intertidal of Galua.

One item of LWD in the Laulua forest contained termites and coleopteran larvae in the high intertidal. Furthermore, only two items of LWD from the Galua and Laulua forests contained teredinids and coleopteran larvae together in the low intertidal. All other wood-biodegrading organisms identified from the LWD in Galua and Laulua forests were found exclusively on their own.

3.4.4 The volume of LWD in mangrove forests

The total volumes of LWD varied between the sites (ANOVA, F3, 66 = 18.31, p = < 0.001). Sombano had the greatest volume of LWD, when compared to the

Langira, Galua and Laulua forests (Figure 3.8). The volume of LWD also varied at different tidal heights between the sites (ANOVA, F3, 66 = 9.58, p = < 0.001). The volume of LWD pieces from Sombano was 0.16 ± 0.025 m3 and the largest log recorded was 9.6 m3. The volume of LWD pieces from Langira was 0.03 ± 0.004 m3 and the largest log was 0.4 m3. The height above sea level recorded for Sombano was

2 ± 0.11 metres above sea level and the mean height above sea level recorded for

Langira was 1 ± 0.35 metres above sea level. In the Galua forest, the total volume of dead wood was significantly greater at the strandline, more than 2m above sea level, when compared to areas further out to the seaward edge, which are between 1.99 -

1.90 and less than 1.90m above sea level. The largest log located in Galua at more than 2.0 metres above sea level was 1.83 m3. The largest logs at 1.99 – 1.90 and less than 1.90 metres above sea level were 0.48 m3 and 0.32 m3 respectively. The volume of dead wood for the Laulua forest was lower than that in the Galua forest. The largest log located at Laulua was also found at more than 2.0 metres above sea level was 0.5 m3, and 0.08 m3 and 0.05 m3 at 1.99 – 1.90 and less than 1.90 metres above sea level respectively.

a 140 Sombano

120

) 1

- 100 ha 3 3 80 60

LWD LWD (m 40 20 0 b > 2 1.99-1.9 < 1.9 10 a Langira ab

) 1

- 6

ha 3 3

4 b

LWD LWD (m 2

0 > 2 1.99 - 1.9 < 1.9 c 30 a Galua

) 1

- 20

3 15 10

LWD LWD (m b 5 b 0 > 2 1.99 - 1.9 < 1.9 d 1.4 a Laulua

1.2

) ab

1 1

- ha

3 3 0.8 b 0.6

0.4 LWD LWD (m 0.2 0 > 2 1.99 - 1.9 < 1.9 Height above sea level

3 -1 Figure 3.8 The volume of LWD (m ha ) found in areas of different tidal heights in different mangrove forests. The volume of decaying wood recorded from: a, Sombano; b, Langira; c, Galua; and d, the Laulua mangrove forest (mean ± SE). Letters above the bars = Tukeys Pairwise comparisons. Note: y-axes are different scales.

3.4.5 Wood-degrading organism distribution and environmental variables – emersion, distance from the land and salinity

In Galua, basidiomycetes had the greatest emersion time (Figure 3.9 a, 1-way

ANOVA, emersion time vs. Galua biodegradation processes; F3, 442 = 318.52, p = <

0.001). The termites and coleopteran larvae had no difference of emersion time, although both animals had a lower recorded emersion time than basidiomycetes. The

Teredinidae had the lowest emersion time compared with the emersion times of all the terrestrial wood-degrading organisms recorded in this study.

Similar observations were found from Laulua (Figure 3.9 b 1-way ANOVA, emersion time vs. Laulua biodegradation processes; F2, 103 = 211.77, p = < 0.001).

The differences of the emersion times calculated for each of the biodegradation processes, clearly demonstrate that the Teredinidae had the least emersion time compared with the terrestrial wood-degrading organisms.

Differences of salinity levels from the strandline out to the seaward edge were found in Laulua. The soil salinity levels recorded in areas where the coleopteran larvae, termites and the Teredinidae dominated were 42.4 ± 1.0, 44.2 ±

1.1 and 33.4 ± 0.2 respectively (1-way ANOVA: F2, 103 = 103.15, p = < 0.001). The mean distances from land for the coleopteran larvae, termites and the Teredinidae were 59.14 ± 9.07, 59.48 ± 10.43 and 201.37 ± 2.12 metres respectively (1-way

ANOVA: F2, 103 = 207.91, p < 0.001).

a a 25 Galua b

b 20

15 c

10 Mean emersion Meanemersion (h)

0 Basidiomycetes Coleopteran Termites Teredinids larvae b 25 a a Laulua

15 b

10 Mean emersion Meanemersion (h) 5 Absent in forest 0 Basidiomycetes Coleopteran Termites Teredinids larvae

Wood-degrading organism

Figure 3.9 Emersion times (hours) of wood-degrading organisms removed from

LWD.a a, wood-degrading organisms from Galua; and b, Laulua (mean ± SE).

In all areas of each forest where terrestrial wood-degradation occurred the heights above mean sea level were higher than in areas of the forest exposed to marine biodegradation only - refer to map (Figure 3.1 b, 1-way ANOVA, height above sea level vs. intertidal zone, F5,644 = 236.84, p = < 0.001).

As emersion time decreases, the biodegrading organisms processing LWD change from terrestrial to marine organisms (Figure 3. 10 a and b). 91

a Basidiomycetes Coleopteran larvae Termites Teredinids 100

20 Items LWD of

0 25 20 15 10 5 0 Emersion (h) b Coleopteran larvae Termites Teredinids 30

5 Items LWD of

0 25 20 15 10 5 0 Emersion (h)

Figure 3.10 The biodegrading organisms found at different tidal heights within

mangrove forests upon LWD. a, The Galua; and b, Laulua mangrove forests. Different

symbols signify different wood-biodegrading organisms.

The environmental variables influencing the dbRDA ordination axes (Figure

3.11) showed that the distance from the land and emersion time are best correlated with axis 1 for both Galua (Figure 3.11 a), and Laulua (Figure 3.11 b). In the Galua forest, 93.4 % of the environmental data explains 42.2 % of the total variation. In the

Laulua forest, 97.2 % of the environmental data explains 52.9 % of the total variation. In the Galua forest, axis 2 of the dbRDA ordination revealed that 5.2 % of the environmental data explains 2.3% of the total variation. In the Laulua forest, axis

2 of the dbRDA ordination revealed that 2.8% of the environmental data explains

1.5% of the total variation. For both forests, distance from shore increases, and emersion time decreases for the Teredinidae. 92

Figure 3.11 dbRDA ordinations. a, in the Galua forest the vectors of presence/absence of wood-degrading organisms in LWD with bubble-plots showing environmental variables that have the strongest influence on the wood- degrading organisms. The dbRDA axes explain the percentage variation of the most parsimonious fit; and b, in the

Laulua forest. Inset boxes - show each environmental variable: DFL = distance from land (m), and emersion time (h). 93

A conditional test for Laulua revealed that DFL explains 97.2 % of the total variation, followed by emersion with 2.8 %. In the DISTLM test, for Laulua the sequential test revealed that the wood-degrading organisms in LWD were correlated with distance from land (DFL), emersion time and salinity (Table 3.3). The most parsimonious model (AIC = 629.35, R-Sq = 0.54) for the wood-degrading organisms explains 54% of the variation, with DFL explaining 50% followed emersion explaining 2 %, and then salinity explaining 2%.

A conditional test for Galua revealed that emersion explains 93.4 % of the total variation, followed by DFL and salinity with 5.2 % and 1.4 % respectively. For

Galua the sequential test revealed that the wood-degrading organisms in LWD were correlated with emersion, DFL and salinity (Table 3.3). The most parsimonious model (AIC = 2731.6, R-Sq = 0.45) for the wood-degrading organisms explains 45% of the variation, with emersion explaining 39% followed by DFL explaining 5 %, and then salinity explaining 2 %.

Table 3 .3 Distance-based linear modelling (DISTLM). Sequential test results relating

wood-degrading organisms (presence/absence transformed with a dummy variable) with

emersion time distance from land (DFL) and salinity in areas where LWD was found in the

Galua and Laulua forests.

Predictor Site variables AIC SS (trace) Pseudo-F P-value Proportion Accumulative df

Laulua DFL 634.72 40971 104.73 < 0.001 0.5 0.5 104

Emersion 631.4 1990.6 5.2989 < 0.05 0.02 0.53 103

Salinity 629.35 1451.3 3.9748 < 0.05 0.02 0.54 102

Galua Emersion 2776.4 141960 282.2 < 0.001 0.39 0.39 444

DFL 2742.2 17413 37.454 < 0.001 0.05 0.44 443

Salinity 2731.6 5768.8 12.737 < 0.001 0.02 0.45 442

3.4.6 The change between terrestrial and marine wood-degrading organisms

A distinct change between wood-degrading organisms was found in LWD from Galua and Laulua. A biodegradation-boundary, determined by the sharp change between terrestrial and marine wood-degrading organisms in LWD from Galua and

Laulua was recorded within a shore height of 0.07 metres of each other (Figure 3.12).

The change between basidiomycetes, teremites, coleopteran larvae and teredinids in

LWD from Galua and Laulua occurred at 1.89 ± 0.01 (mean ± SE) metres above sea level and 1.82 ± 0.02 metres above sea level respectively.

95 Galua a 1000 Basidiomycetes Coleopteran larvae Teredinidae Termites Boundary 900

800

700

600

500

400

Distance from Distance from shoreline (m) 300

200

100

0 0 50 100 150 200 250 300 b Distance along strandline (m) Laulua 250

200

150

100 Distance from Distance from shoreline (m)

0 0 20 40 60 80 100 120 140 160 180 200 Distance along strandline (m)

Figure 3.12 The distribution of wood-degrading organisms in LWD. a, wood-degrading

organisms identified as different symbols in the Galua; and b, Laulua forests. Dotted line =

boundary between terrestrial and marine wood-degraders. All LWD items with no visible

degradation were not plotted, and the area between 220 – 600 metres in Galua was not sampled. 96

3.4.7 Teredinid panel-fragmentation

Analysis of panel fragmentation demonstrated that with increasing teredinid

attack intensity an increasing percentage of the panel broke-up (Figure 3. 13,

Regression analysis, R-Sq (adj) = 0.41, p = < 0.001). Within mangrove forests and in

reef sites, the percentage of the sectional area represented by tunnels vs. the

fragmentation of panels remaining found that teredinid activity and panel

fragmentation increased in the reef sites compared with panels exposed in the forests.

Across the longitudinal section of each panel from the mangrove sites, the percentage

of the sectional area represented by teredinid tunnels was 35.1 ± 4.3 % (mean ± SE, n

= 30) and the percentage surface area of mangrove panel-loss (fragmentation) was

1.8 ± 0.6 %. Within the reef sites across the longitudinal section of each panel, the

percentage of the sectional area represented by teredinid tunnels was 92.4 ± 1.4 %

(mean ± SE, n = 30), and the percentage surface area of the reef panel-loss was 29.1

± 4.3 %.

80 Intertidal Subtidal

60 50 40 30

% of total loss total ofpanel% 20 10 0 0 20 40 60 80 100 % of total area represented by teredinid tunnels

Figure 3.13 Wood loss through teredinid activity within panels in mangrove forests and coral reefs. The percentage of total panel loss estimated from panels exposed to different levels of teredinid attack within intertidal (grey dots) and subtidal (black dots) environments (regression analysis, teredinid attack intensity vs. panel fragmentation; F1,58 = 41.32, p = < 0.001).

The mangrove forests in the Tukang Besi Archipelago have a wide range of

LWD volume, comparable with other mangrove forests from different countries

(Table 3.4). The Sombano forest had an LWD volume similar to the upper ranges of

LWD volume reported from Micronesia (Allen et al., 2000) and Florida mangrove

forests hit by hurricanes (Krauss et al., 2005).

Table 3.4 Volume estimates of LWD (metre cubed per hectare) from different

mangrove forests. LWD volume data reported from mangrove forests outside of

Indonesia were obtained from other studies.

Country Mangrove LWD volume (m3 ha-1) Reference Micronesia Yap 35.1 Allen et al., 2000 Pohnpei 43.1 Kosrae 104.2 USA Taylor River 12.6 Krauss et al., 2005 Johnson Mound 56 Creek Everglades 147.9 Malaysia 160 Chan, 1982 Indonesia Laulua 0.8 This study Langira 4.6 Galua 8.1 Sombano 102.9

3.5 Discussion

The volume of LWD estimated from the Galua and Langira forests are similar to the small forests found in Florida. However, in the Laulua forest the volume of LWD was very low. This is not surprising as this mangrove forest had been heavily harvested. In addition, the volume of living biomass reported in chapter two reflect the volume of LWD produced from those same forests. The carbon 98 contained within the LWD is important, as carbon is retained within the sediments

(Donato et al., 2011), and the mineralization of LWD items provide energy for many different trophic pathways when degraded (Benner and Hodson, 1985; Robertson and

Daniel, 1989a).

In a study of mangrove wood degradation in the Caribbean the mineralization rate of mangrove wood was found to be ten times slower when compared to the rates of leaf litter (Benner and Hodson, 1985). When in anaerobic sediments, the rate of wood mineralization is thirty times lower compared with aerobic mineralization rates

(Benner and Hodson, 1985). This means that LWD will remain in anaerobic sediments for a long time, which may explain the very large volume of LWD in the

Sombano forest – as the sediment in Sombano is anoxic.

The environmental factor that determines differences of dead wood volumes between Sombano and Langira is emersion time. In a two year study from a tropical

Australian mangrove forest more than fifty percent of the mass of LWD was lost in the mid- and low- intertidal regions of the forest (Robertson, 1990). However, only five percent of the woody mass had been lost in the high intertidal zone (Robertson,

1990). The large loss of LWD was attributed to the wood-boring activity of the

Teredinidae. This study demonstrates that tidal inundation is the environmental factor contributing to the differences of woody-mass from the panels – as the distribution of wood-degrading organisms within mangrove ecosystems are affected by immersion (Kohlmeyer et al., 1995). Other authors have reported upon the influence of tidal inundation as a driving factor for wood-dependent mangrove animals. A study of mangrove forests in East Africa determined that numbers of wood-boring isopods increase with decreasing emersion times (Svavarsson et al.,

2002). Furthermore, in Caribbean (Kohlmeyer et al., 1995), and Australian mangrove 99 forests (Robertson and Daniel, 1989a; Robertson, 1990) teredinids also increase with decreasing emersion times.

In Sombano, the decay of mangrove wood was very slow – no loss of wood was found from the panels even after one year exposure. The slow degradation of

LWD may be due to the anoxic substrata and the kinds wood-degrading organisms found within LWD in that forest. It is known that wood decay in anaerobic environments is very slow (Benner and Hodson, 1985). However, woody debris in

Langira may be exported by tidal water flow, meaning that the usual wood degradation processes may not have time to initiate the lignocellulose breakdown before the physical processes such as tidal flow exports wood from the forest. Tidal inundation is particularly important in the mid- to low-intertidal areas of mangrove forests, as tidal flow will frequently export much of the woody debris (Boto and

Bunt, 1982; Robertson and Daniel, 1989a). Differences of degradation processes magnify the importance of nutrient recycling in mangrove ecosystems (Lang and

Knight, 1979), as wood degrades at different rates from the high to low intertidal regions of mangrove forests (Sherman et al., 2003) - due to differences of wood- degrading organisms from the high to low intertidal regions (Kohlmeyer et al.,

1995).

Wood decay is accelerated further inland when compared to wood decay in the seaward zones (Sherman et al., 2003). Wood-degradation is dependent on habitat characteristics such as decomposer communities, tidal energies, site fertility and ambient temperatures (Davis et al., 2003). In the Sombano forest, the biodegradation processes that break down LWD are conducted by organisms that are also frequently found in terrestrial forests such as beetles (Novotny et al., 2006), termites (Jones and

Eggleton, 2000) and basidiomycete fungi (Tedersoo et al., 2003). 100

However, in the Langira forest the wood-degrading organisms were different.

Unlike the Sombano, in the high-intertidal, the Langira forest extends down to mean sea-level. Waterlogged LWD will be inhibited from saprophytic decay by obstructing fungal growth due to the limited availability of oxygen, therefore decreasing respiration rates - further explaining the lack of basidiomycetes in areas of the forest with reduced emersion times (Progar et al., 2000).

In the landward regions of the Galua forest, as in Sombano basidiomycetes were found upon many items of LWD, and this wood was very soft due to the high level of decay. Wood-decaying fungi are essential for the functioning of forest ecosystems (Lonsdale et al., 2008; Ananda and Sridhar, 2004), as they control the primary breakdown of wood. This means that basidiomycetes are particularly important in the high-intertidal regions of Galua and in the Sombano forest as there was a large volume of LWD. Of the known 900 marine fungi, 358 have been recorded from mangrove ecosystems (Maria and Sridhar, 2002; Maria and Sridhar,

2003). However, both terrestrial and marine fungi are involved in the decomposition of litter in mangrove ecosystems (Ananda and Sridhar, 2004). Marine fungi are able to tolerate salinity, anoxic sediments and tidal inundation (Kohlmeyer et al., 1995;

Gilbert and Sousa, 2002). Salt-tolerant fungi are found on a variety of substrata such as LWD, macroalgae, seagrasses and even deep sea sediments (Ravikumar et al.,

2009). Terrestrial fungi within mangrove ecosystems are often ignored as transient mycoflora (Maria and Sridhar, 2003; Ananda and Sridhar, 2004), even though they contribute to substantial woody degradation due to their lignin-modifying enzymes

(Nilsson, 1973).

Differences of the mangrove mycoflora were found between the mangrove forests. A variety of fungi develop within an assortment of different ecological 101 niches in mangrove ecosystems such as recurrently submerged decaying wood, exposed decaying wood and damaged mangrove roots (Sarma and Vittal, 2001;

Ananda and Sridhar, 2004). These differences may have also been determined by the distribution of the host substratum within each mangrove forest, combined with prolonged exposure of woody debris to direct sunlight, heat and strong winds - resulting in desiccation (Prasannarai and Sridhar, 2003). No basidiomycete species were recorded upon the LWD in the Laulua forest. This forest has been exposed to extensive harvesting pressures, leading to large open areas exposing the sediments and LWD making the impacted areas arid and dry.

In Caribbean mangrove forests, which have low mangrove tree diversity, the mycoflora have been shown to be host-specific (Gilbert and Sousa, 2002). Decaying mangrove wood colonised by particular fungi often show preference for a specific substratum (Sarma and Vittal, 2001). Thus, if the host is site-specific, then the fungi processing that woody substratum will have a restricted distribution (Ananda and

Sridhar, 2004). This was evident with the mycoflora in Sombano and Galua as these forests contained tree species specific to each site which had distinct basidiomycete species found only on that particular tree.

Environmental factors will also determine the distributions of the basidiomycetes, such as freshwater input (Ananda and Sridhar, 2004; Vittal and

Sarma, 2006) and height above sea-level (Ejechi, 2002). Salinity and emersion times were found to be correlated with the distribution of the mycoflora in the Sombano and Galua forests. The basidiomycete fruiting bodies were always found above the high tide mark on LWD in Sombano and Galua, in regions of low salinities and high freshwater input from rainwater. Thus, the dominance of terrestrial fungi upon mangrove LWD may be dependent on the position of the LWD in the intertidal 102 region (Ejecti, 2002), and the duration and intensity of rain storms (Ananda and

Sridhar, 2004).

In the Sombano and Galua strandline, environmental conditions for basidiomycete communities were favourable, with reduced salinities and prolonged emersion times. A change in salinity and emersion will alter the distributions of the basidiomycetes (Castillo and Demoulin, 1997; Ejechi, 2002). Factors influencing rates of saprophytic decay directly control nutrient regeneration and indirectly affect the rates of secondary production by animals in mangrove ecosystems (Benner and

Hodson, 1985). Thus, the flux of carbon and energy from ligninolytic activity to the animal consumers is the foundation of an intricate and essential trophic food-web

(Babu et al., 2010). The pathways of wood breakdown were different in the Langira forest. The reduced levels of wood degradation recorded from Langira may have been attributed to the lack of white-rot and brown-rot decay. The basidiomycetes were not found on LWD in the Langira forest, meaning that the breakdown of lignin due to ligninolytic activity did not occur within LWD from Langira. The level of wood degradation was less and the breakdown of LWD in Langira was caused by the wood-boring activity of the Teredinidae.

The basidiomycete distributions in Galua were determined by the distance from land and salinity. The basidiomycetes were found only at the strandline with hightide salinities not above 16 PSU. In all cases, the basidiomycete fruiting bodies identified from the Galua forest were found more than 2 metres above sea level, meaning that the basidiomycete fruiting bodies were not immersed by seawater. The limited exposure to seawater determines the presence of the basidiomycetes within the intertidal zone of the Galua forest. This limited exposure explains the presence of

Trametes versicolor in Galua, whose growth is strongly inhibited by seawater (Jones 103 and Byrne, 1976 cited in Castillo and Demoulin, 1997). The distributions of

Microporus xanthopus and Pycnoporus sanguineus in coastal forests from Papua

New Guinea were limited by their restricted abilities to cope with high salinity concentrations (Castillo and Demoulin, 1997), which corroborates this study.

Nonetheless, P. sanguineus had similar distributions to the salt-tolerant

Schizophylum commune, a euryhaline species (Castillo and Demoulin, 1997).

Pycnoporus sanguineus were able to survive in areas of increased salinities due to the large growth forms of the fruiting body, allowing it to maximize exposure to periods of high rainfall that may likely limit salt accumulation (Castillo and

Demoulin, 1997).

Mechanical degradation also occurred within the Galua forest. Coleopteran larvae and the termite communities found within LWD from Galua had very similar distributions to each other. It is likely that both coleopteran larvae and termites were not able to tolerate sustained periods of immersion. Coleopteran larvae are responsible for canopy-gap formations, increased litter-falls and nutrient enrichment due to their boring within trees (Feller, 2002). Evidence of the mangrove longhorn beetle, Aeolesthes holosericeus was found in many standing and fallen Excoecaria agallocha trees in the Sombano forest. The larvae of this beetle create large characteristic circular blisters – these blisters have also been reported on E. agallocha trees from mangrove forests in Singapore (Ng and Sivasothi, 1999). False blister beetle larvae from the family Oedemeridae were very common in LWD from

Langira, Galua and Laulua in the high- to mid- intertidal regions of the forests. These larvae are commonly found world-wide in marine drift wood and pilings (Arnett,

1951). Furthermore, oedermerids are able to resist tidal immersion (Arnett, 1951), which explains their wide distribution in the intertidal regions of the forests. 104

Termites were also found in LWD and in arboreal nests in the Sombano,

Galua and Laulua forests. Arboreal termites live in well-defined nests that are separated from their foraging grounds (Korb, 2008). Tree-dwelling termites are considered to be long distance foragers (Roisin et al., 2006), as they are able to travel in relative safety because their runways are covered (Merritt and Starr, 2010).

Furthermore, when on the forest floor foraging for food during low tide, the arboreal termite genera in the Galua and Laulua forests, Prorhinotermes, Coptotermes and

Nasutitermes are able to escape the incoming tide by retreating back up the trees to their nests. However, wood-dwelling termites live within the galleries they create in

LWD from their feeding activities (Adams et al., 2007). Wood-dwelling termites,

Cryptotermes were found frequently within decaying wood in the landward zones of the Galua and Laulua forests. Cryptotermes colonies remain within a single item of decaying wood, which serves as both food and shelter (Korb and Lenz, 2004).

Termites within decaying wood, such as the drywood genus Cryptotermes, have worker termites that are totipotent (Korb, 2008). Remaining within LWD will have benefits if members of the termite colony are able explore all caste options (Korb,

2008) such as; predator and risk avoidance due to no external foraging. Cryptotermes are also able to detect substrate-borne vibrations and these termites are able to determine the size of their detrital food item (Adams et al., 2007; Evans et al., 2007).

Thus, vibration detection may help these termites to avoid potential immersion by escaping areas within their galleries that may be submerged before the incoming tide reaches them as the termite galleries in many LWD items extended from below to above the high tide mark. The well-developed nests also provide environmental buffering from outside unsuitable temperatures (Korb, 2008).

Other wood-degrading organisms were found in LWD in the low- to mid- 105 intertidal areas of the mangrove forests, and those organisms - such as limnoriid isopods and teredinids, are predominantly aquatic. However, these crustaceans are not usually found in forests with freshwater input such as those forests in this study, and are more common within Caribbean mangrove forests (Cragg, 2007). Analysis of

LWD from the Loho forest revealed a new limnoriid species, Limnoria sellifera

(Cookson et al., 2012). Further research of wood-dependent animals in the mangrove forests of the Tukang Besi may produce other new species.

Life history strategies between teredinid species vary between those with non-brooded and brooded larvae. Species belonging to the genera Lyrodus and

Teredo retain their larvae to the late veliger stage. After spawning, larvae from larviparous species can be widely distributed, even though they have a short free- swimming period (Turner, 1966). Short and long term brooded species have a successful recruitment rate and are more likely to become world-wide in distribution within their limits of temperature and salinity (Brearley et al., 2003). Other studies also demonstrated that Lyrodus and Teredo have a greater abundance in LWD than non-brooded species (MacIntosh et al., 2012). The Lyrodus and Teredo larvae are protected during their early critical stages within the parent. Studies from Aqaba demonstrated that T. bartschi retain their larvae in the gills. Larval release at the non- feeding pediveliger limits larval dispersal (Cragg et al., 2009). The advanced development of T. bartschi larvae means that they are able to settle and colonise wood more rapidly compared to less developed larvae. Thus, brooded young have advanced development when released, and are therefore less sensitive to environmental variables (Turner, 1966). In contrast, Spathoteredo obtusa produce non-brooded planktotrophic larvae. When compared to settlement rates with brooded species, S. obtusa were less successful, as very few S. obtusa pallets were found. 106

Although planktotrophic development would appear to deliver the greatest dispersal range, oviparous teredinids are not more widely distributed compared to brooded teredinids (Cragg et al., 2009). The teredinids, Bactronophorus thoracites and

Dicyathifer manni were found only in LWD. Other authors also reported that these teredinids were found only in LWD and not in experimental panels (Rayner, 1983).

B. thoracites and D. manni may preferentially settle on wood with high levels of tannins even though toxic extractives have also found to be fatal for other wood digesting animals, Limnoria isopods (Borges et al., 2008).

Adult teredinids are sessile, and have isolated breeding populations maintained by fragmented areas of LWD. The availability of LWD to teredinids in mangrove ecosystems is stochastic, meaning that LWD items can be few and far between in some areas and dense in other areas (Duke et al. 1981).

Teredinid densities in mangrove ecosystems are higher with decreasing emersion times - due to an increased exposure to seawater (Rimmer et al., 1983;

Leonel et al., 2006). The Teredinidae are more numerous in the low intertidal, similar to the distributions of sphaeromatid populations recorded from mangrove forests in East Africa (Svavarsson et al., 2002). In areas of the Langira forest with higher emersion times above nineteen hours, no loss of wood was recorded from the panels. However, in areas with lower emersion times the dry-weight lost from the panels was due to the activities of the Teredinidae. Larval teredinids were not able to reach the high-intertidal areas of the forest – as evidenced by the lack of attack but larvae might reach but not survive. Thus, the distributions of LWD from the high- to the low- intertidal areas within mangrove ecosystems will have a large influence upon the amount of wood-degradation and mechanical recycling - due to the inability of the teredinids to extend up into the high-intertidal and the inability of terrestrial 107 wood-degrading organisms to extend down into the low-intertidal. These differences are attributed by the position of the LWD within the intertidal region of the forest; therefore the sharp change of wood-loss between terrestrial and marine biodegradation processes upon LWD, are controlled by tidal inundation.

The differences of dry-weight losses observed between the panels mirrored the estimates of the LWD volumes recorded from Sombano and Langira. Losses of wood were greater with increasing distance from the land. Within the low intertidal zones of the forests, items of LWD were smaller and more fragmented than the LWD within the high intertidal. Furthermore, the higher degree of panel fragmentation recorded from the sub-tidal panels may have also been due to high wave energy input upon the reefs. Indeed, the tunneling activity of the Teredinidae will reduce the structural integrity of fallen wood. Thus, coupled with high wave-energies the fragmentation process may be enhanced

Sharp differences of degradation processes were found between the terrestrial and marine wood-degrading organisms in Sombano and Langira. A fossil-coral plateau separates the Sombano and Langira forests. Thus it was not possible to examine the environmental gradients that create the sharp differences between terrestrial and marine wood-degrading organisms. It could be argued that the fossil- coral plateau created the change. However, in the Galua and Laulua forests the same terrestrial and marine wood-degrading organisms were found; Galua and Laulua are not separated by a fossil-coral plateau, providing ideal environments to determine the likely factors that facilitate the changes in wood-degrading organisms. Similar distributions of wood-degradation processes between fungi and teredinds were also reported from mangrove forests in Belize (Kohlmeyer et al., 1995). Teredinids and fungi work in parallel – with fungi decaying wood in the high intertidal and 108 teredinids processing wood in the low- intertidal regions (Kohlmeyer et al., 1995).

Each wood-degrading organism has a specific niche, although terrestrial and marine organisms are occasionally found processing the same LWD item. However, the niches occupied by these organisms are controlled by tidal inundation within the intertidal mangrove forests of the WMP.

The basidiomycetes, coleopteran larvae, termites and the Teredinidae are the essential catalysts for the breakdown of LWD from the high- to low- intertidal regions. Those processes contribute to the release and flux of organic carbon in mangrove habitats and maintain a complex trophic web (Benner and Hodson, 1985;

Bouillon et al., 2008; Kristensen et al., 2008). This may mean that wood-degrading processes spanning different tidal heights within mangrove forests may be affected by global warming, given that the change between the terrestrial and marine wood- degrading organisms can be seen over one – two metres on the forest floor.

Rising sea-levels and increasing atmospheric CO2 leading to decreasing pH concentrations (Guinotte and Fabry, 2008; Ries et al., 2009) may alter the distributions of wood-degrading organisms as emersion and salinity influence their distribution. This may have consequences for mangrove forests that will not permit landward accretion, as mangrove-tree distribution and mangrove-tree zonation are sensitive to different salinities, site fertility (Tomascik et al., 1997), and tolerance to water-logging (Pi et al., 2009) . Furthermore, the effect of decreasing pH may have a marked effect upon the success of teredinid larval settlement. The bivalves Mytilus edulis and Crassostrea gigas both exhibit declines of calcification with increasing pCO2 levels of 25 and 10% respectively (Gazeau et al., 2007) – the same effect may impact the development of teredinid larvae. Decalcification of teredinid shells may impede their ability to tunnel in to LWD, therefore reducing the amount of wood- 109 degradation in the mid- to low- intertidal regions of mangrove forests.

The terrestrial processes – basidiomycetes, termites, and coleopteran larvae are the major catalysts to the very slow turnover of LWD in the high intertidal zone.

The biodegradation of LWD attributed by these organisms is much slower when compared with the marine wood-degraders, teredinids.

Teredinids have a remarkable ability to breakdown woody detritus, therefore influencing the flux of carbon within intertidal mangrove forests. If teredinids were not found in the mid- to low- intertidal areas of mangrove forests there would be a massive build-up of detritus, leading to a reduced flow of nutrients to adjacent coastal ecosystems. The predominant breakdown of LWD attributed by the

Teredinidae has revealed their crucial ecological niche.

110

Chapter 4

THE EFFECT OF TEREDINID TUNNELLING ON MANGROVE PROP-ROOTS

4.1 Introduction

Herbivorous animals from terrestrial to aquatic environments consume a range of photosynthetic organisms from algae to trees. In terrestrial tropical forests, insects such as caterpillars and beetles consume much of the foliage and trunks of the trees (Novotny et al., 2007). Some trees are protected from herbivorous attack as they have an active anti-herbivore defence provided by ants that predate the attacking herbivores (Gaume and McKey, 1997), and other trees have chemical and structural anti-herbivore defence mechanisms (Turner, 1976; Brooks and Bell, 2002). Within forested intertidal habitats, mangrove trees are attacked by both terrestrial and marine animals (Feller, 2002; Ellison and Farnsworth, 1996). The mangrove herbivores range from insects such as beetles (Perry, 1988; Feller and Mathis, 1997) to wood- boring aquatic molluscs, teredinids (Robertson and Daniel, 1989a). Mangroves also have non-herbivorous aquatic wood-boring isopods such as sphaeromatids (Si et al.,

2002).

Damage to prop-roots may either be caused by herbivorous and non- herbivorous attack or physical actions such as abrasion (Gill and Tomlinson, 1977), or by falling branches from the canopy. Increased tidal inundation may also enhance the breakdown of damaged mangrove root tissues – as densities of wood borers become more numerous with increasing immersion times (Svavarsson et al., 2002).

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Isopods process less wood, when compared with teredinids. Sphaeromatids bore large conspicuous holes in live prop-roots (Simberloff et al., 1978). Teredinids on the other hand, predominantly process dead wood and create many tunnels within the logs through their boring activity (Filho et al., 2008).

Studies of wood-boring animals upon mangrove root tissues are primarily focussed upon sphaeromatid isopods (Perry, 1988). The destructive effect of the sphaeromatid, S. peruvianum on live mangrove root-tissues in Costa Rica can reduce the growth rates of Rhizophora aerial prop-roots by fifty percent (Perry, 1988). The decrease in root growth rate is caused by sphaeromatids damaging root tips – where somatic growth of the root occurs (Gill and Tomlinson, 1971). It is thought that sphaeromatids are unable to burrow in to older developed roots that reach the substratum due to the development of woody tissue in older roots (Rehm, 1976;

Perry, 1988).

Roots burrowed by the isopods at their tips may branch, (Rehm, 1976; Ribi,

1982) though it is not clear whether this harms the trees. On the one hand, the effect of the attack of sphaeromatids upon mangrove root tissues may benefit or harm the trees (Simberloff et al., 1978). The death of roots may lead to tree collapse (Rehm,

1976; Svavarsson et al., 2002). Rehm and Humm (1973) go further and say that the mangrove forests are shrinking because of the destruction of the prop roots by S. terebrans in south-western Florida. However, other authors found that sphaeromatids may benefit mangrove forests (Simberloff et al., 1978). Isopods may enhance tree survival by triggering an increase branching of lateral aerial roots, therefore securing the tree from falling due to an increase of root structures (Simberloff et al., 1978;

Ribi, 1982). 112

To combat increasing attack from herbivores plants release tannins (Bloch,

1952). Plants are successfully able to survive in environments where herbivores are common due to their ability to resist or recover from repeated herbivory (Brooks and

Bell, 2002; Hanley et al., 2007). Direct defence mechanisms may reduce herbivory or prevent herbivory on root tissue from further exacerbating root damage caused by feeding upon the plant tissues. Post-attack, mangroves may produce chemical defences such as secondary metabolites such as tannins (Alongi, 1987; Feller, 2002).

It is beneficial for the plant to produce these chemical compounds only when they are required as herbivore attack is random and variable (Hol et al., 2004). As, the energy required to produce these chemical defence exudates can be costly (Agrawal et al., 1999; Karban et al., 1997). Indeed, particular mangrove trees may express resistance only when it is needed to reduce energetic cost into producing the tannins.

Energy expenditure may eventually lead to a decrease in fitness - in this case, root death. The loss of fitness may be due to increased energy investments and use of resources (Agrawal, 1999). Conversely, herbivory may be beneficial (Paige and

Whitham, 1987; Paige, 1992). Herbivory on plants elicits a physiological response which can create an overcompensation of tissue regrowth (Lennartsson et al., 1998).

Increased plant growth may ultimately achieve greater fitness (Paige and Whitham,

1987).

Though damage by wood-boring isopods upon mangrove roots is widely reported, there are few reports of teredinids damaging Rhizophora prop-roots

(Roonwal, 1954, see Figure 4.1). Larval teredinds are unable to infest living trees, as the bark contains high tannin content (Rimmer et al., 1983; Kohlmeyer et al., 1995).

However, the teredinid Bactronophorus thoracites (Gould) is able to penetrate living woody tissues (Roonwal, 1954), as this species may settle preferentially on wood 113 with a lot of tannins. Other teredinid species are not so successful, larval Lyrodus pedicellatus died when trying to settle upon Dalbergia retusa, Panamanian rosewood

(Turner, 1976). The dominant heartwood extractive of Dalbergia is obtusaquinone, a quinone. Quinones are produced by many tropical trees, and they prevent the production of well-formed shells and pallets of teredinids and therefore hinder settlement of Lyrodus larvae (Turner, 1976). Toxic extractives have also been found to be fatal for the wood boring isopod Limnoria (Borges et al., 2008), and chemicals such as terpenoids, triterpenes and flavonoids - present in the root bark of Bruguiera gymnorhiza trees (Han et al., 2004) are toxic to other invertebrates (Spurr and

McGregor, 2003).

Tannins are defence mechanisms. Terrestrial plants also express resistance to herbivory via strategies labelled as tolerance mechanisms (Paige, 1999; Brooks and

Bell, 2002). Tolerance mechanisms are defined as tissue repair and regrowth.

Mangrove plants may likely use the tolerance strategy against damage (Brooks and

Bell, 2002). Tolerance to herbivory or the ability to recover and regrow tissues after damage is an additional plant strategy, which may have evolved in response to herbivore attack (Strauss and Agrawal, 1999).

Teredinids contribute to the breakdown of large woody detritus in the mid- to low-intertidal regions of the mangrove forests (Robertson, 1990). Much of the dead wood consumed by teredinids is replaced by the many tunnels they create through their boring of the wood (Robertson, 1990; Distel, 2003; Filho et al., 2008).The

Teredinidae are particularly important in mangrove forests where there are high levels of detrital woody biomass consisting of fallen trunks, branches and prop-roots as they mechanically break down the woody biomass (Robertson, 1990; Kohlmeyer et al., 1995). Although, teredinid attack may occur only after the bark has 114 deteriorated (Kohlmeyer et al., 1995) and the toxic extractives have leached away

(Borges et al., 2008). The colonised decaying wood may take anywhere from two to fifteen years to completely break down (Robertson and Daniel, 1989a; Kohlmeyer et al., 1995). Eventually, the broken-down woody fragments, teredinid faeces and teredinid biomass may be consumed by marine organisms in adjacent ecosystems

(Robertson and Daniel, 1989a; Granek et al., 2009). This means that wood-borers may benefit mangrove ecosystems by breaking down dead plant material, even though they do cause damage to living tissues (Barkati and Tirmizi, 1991).

Figure. 4.1, Rhizophora stylosa prop-roots. a, undamaged healthy prop-roots – a typical

view of roots demonstrating no obvious evidence of damage to their epiblema (outer layer),

and b, a compromised root showing extensive damage resulting in root death, not

commonly found. The arrow indicates the opening of an old teredinid tunnel.

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As long as their epiblema remains intact, mangrove roots are very effective at keeping wood-borers such as teredinids at bay (Kohlmeyer et al., 1995). However, soon after the outer layer is damaged and the inner root-tissues are exposed, larval teredinids can settle on the wood, and begin to process the tissues (Kohlmeyer et al.,

1995).

This study aims to determine the colonisation of teredinids and the regenerative capabilities of mangrove prop-root tissues under different levels of damage. This work tested the prediction that Rhizophora roots with a low level of damage will not be attacked by teredinids after one year and will make a full recovery, roots with a moderate level of damage will elicit an over-compensatory level of tissue re-growth, and roots with a severe level of damage will die due to teredinid attack after one year.

Rhizophora prop roots were experimentally damaged in the forests to determine which prop-root surfaces are colonisable by teredinids. The protective layers of the roots were removed to measure the ability of prop-roots to recover and the extent of infestation by teredinids exposed to different levels of damage. This investigation reveals how prop-root tissues respond to different levels of damage and the affect of teredinids upon those damaged areas.

4.2 Materials and methods

4.2.1 Root scar experiment

The long-term effects of experimental damage to mangrove roots were examined. The experiments took place in three inter-tidal Rhizophora stylosa- dominated mangrove forests in the Tukang Besi archipelago, Kaluku, Loho and One 116

Onitu (see chapter 2 for site details). Three ten by ten metre plots within each forest were designated for the experimental damage. The plots were chosen in these areas because teredinids were found and no evidence of disturbance was recorded. The damage consisted of one of three surgical treatments of increasing severity at the base upon individual roots. Within each plot, each treatment was conducted on three roots, totalling nine roots per plot, and twenty-seven roots per site. Within all the sites, a total of eighty-one roots were used.

The first surgical treatment (superficial) consisted of removal of the outermost layer of the bark, the epiblema (Figure 4.2 a). The second treatment

(moderate) consisted of additionally removing the starch-rich parenchyma cells

(Figure 4.2 b). The third treatment, (severe) consisted of additionally removing the inner radial and tangential walls of the endodermis containing lignin and suberin-rich tissues (Figure 4.2 c). All surgical treatments were cut into live Rhizophora roots using a diver’s knife. Each root-scar measured 4 cm in diameter by 10 cm in length.

Digital images were taken of each treated root over twelve months. All root circumferences and scar depths were measured before and during the twelve month period using a tailors tape and callipers.

4.2.2 Root biscuit analysis

Root scars on twenty-four of the experimentally damaged roots were cut transversely into biscuit sections after twelve months after the damage. For each of the levels of damage, eight roots were used. For the superficially damaged roots, a total of forty biscuits were cut, and for the moderately damaged roots, thirty-six biscuits and for the severely damaged roots a total of thirty-five biscuits were cut.

Each biscuit from the damaged roots was inspected for evidence of necrotic cellular 117

damage and then measured to calculate the area of cellular rehabilitation or loss. If

present, all teredinid tunnels in each biscuit were counted, and percentage surface

area of tunnels within each biscuit were estimated using ImageTool Version 3.00

(The University of Texas Health Science Centre at San Antonio).

Ep Co En

superficial moderate severe

Figure. 4.2 Three treatments of root scars characterised by different levels of damage. a, the removal of the outer epiblema layer, superficial damage, b, removal of the carbohydrate-rich cortex tissue , moderate damage, exposing the epidermis layer, c, removal of the lignin-rich endodermis layer, severe damage. Ep = epiblema, Co = cortex, En = endodermis, Ls = longitudinal section and Ts = transverse section.

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4.3 Data analysis

Biscuit analysis

Values of percentage surface area of tissue regrowth/loss and the number of teredinid tunnels within root-biscuits were tested for differences of teredinid tunnels in roots exposed to different surgeries using Box Plots.

Live roots

Upon the live roots, differences of the percentages of root tissue regrowth/loss over one year were tested using a 2-way ANOVA with treatment and time as factors. All percentage data were normalized using the arcsin transformation.

All statistical analyses were performed using MINITAB (MINITAB Inc, version

13.20).

4.4 Results

4.4.1 Root scar analysis upon live roots

Twelve months after the initial surgeries more than 95 % of roots exposed to the superficial damage had a complete recovery of tissues and total regrowth of the epiblema. No roots died when exposed to this level of damage. Almost 80% of roots exposed to a moderate level of damage had a full recovery with 30% of those roots exceeding the initial size of the root before surgery (Table 4.1). However, 14.8% of the roots exposed to the moderate level of damage died. More than 60% of the roots exposed to a severe level of damage made a full recovery with 25% of those roots exceeding the initial size of the root before surgery. However, 7.4% of the roots exposed to the severe level of damage died. The greatest percentage regrowth of root-tissue was recorded from roots exposed to a moderate and severe level of damage (mean ± SE). 119

Table 4.1, Tissue regrowth measurements (%) of roots exposed to three levels of damage.

The percentage of root tissue remaining after the initial surgery (0 days) and the percentage of tissue regrowth of those tissues after 12 months from the three levels of damage. Exposure after surgery Root Damage

Superficial Intermediate Severe

0 days 95.7 ± 0.5 70.9 ± 0.6 64.8 ± 1.1

12 months 102.4 ± 0.4 90.2 ± 7.6 93.4 ± 5.6

Differences were found with the percentages of root-tissue regrowth after 12

months when compared from the initial surgery between the roots with increasing

levels of damage (2-way ANOVA, Table 4.2).

Table 4.2, 2-way ANOVA results. Root treatments = A,

B and C, and Time (treatments at zero days and

treatment regrowth after 12 months).

Factors DF F-value P-value Treatment 2,156 8.2 < 0.001

Time 1,156 18.8 < 0.001

Treatment*Time 2,156 5.7 < 0.005

Tissue colour changes were observed and timed over 2 hours after each

surgery. The colour changes within the outer cortex layer after removal of the

epiblema occurred within thirty minutes on roots with superficial damage (Figure 4.3

a). These colour changes signify an almost immediate physiological defence-

response by the plants. In comparison to the superficial damaged roots, colour

changes observed from the moderately (Figure 4.3 b) and severely damaged roots

(Figure 4.3 c) were more rapid and conspicuous.

Figure 4.3 Timed colour changes and wound healing of the release of tannins in prop-roots

exposed to increasing levels of damage. a, roots exposed to superficial, b, moderate and c,

severe levels of damage over 12 months. Continues from page 114 to 116.

120 a 0 days 15 min 45 min 5 days

10 days 20 days 35 days 12 months

121 b 0 days 25 min 65 min 105 min

3 days 6 days 40 days 12 months

122 c

0 days 20 min 35 min 65 min

2 days 7 days 40 days 12 months 123

4.4.2 Root biscuit analysis

Differences were found between roots exposed to increasing levels of damage with the percentage of tissue rehabilitation or loss (Figure 4.4 a). Roots exposed to superficial damage had the lowest recorded percentage of tissue regrowth.

The percentage of tissue regrowth for the 40 biscuits from superficially damaged roots showed no changes with quartiles (Q) 1 and 2. Q3 and Q4 had between 1.5 and

2.4 % of tissue regrowth after 12 months. The percentage of tissue regrowth for the

36 biscuits cut from moderately damaged roots demonstrated that Q1 and Q2 had between -7 and 5.7 % of tissue loss and regrowth respectively. Q3 and Q4 of the biscuits from moderately damaged roots had between 13.5 and 21.3 % of tissue regrowth respectively. Two outliers were found with the biscuits from moderately damaged roots, with losses of -9.1 and -15.7 % of tissues over one year. The percentage of tissue regrowth for the 35 biscuits from severely damaged roots demonstrated that Q1 and Q2 had between -41.1 and -21.8 % of tissue loss respectively over one year. Q3 and Q4 had between 11.2 and 16.5 % respectively of tissue recovery over one year.

Differences of teredinid tunnels within the root-biscuits were also found

(Figure 4.4 b). No teredinid tunnels were recorded within the superficially damaged biscuits. With each quartile of the biscuits from moderately damaged roots, no teredinid tunnels were recorded. However, two outliers were found with 1 and 2 teredinid tunnels. Biscuits from roots severely damaged were different from both the biscuits from superficial and moderately damaged roots with Q3 and Q4 having 1 and 2 teredinid tunnels respectively. The analysis of the biscuits from severely damaged roots revealed three outliers, with 3, 6 and 10 tunnels.

124 a

Figure. 4.4 Box plot analyses of the roots exposed to increasing levels of damage. a, The variation

of root tissue loss and regrowth estimated from biscuits obtained from roots with superficial, moderate

and severe levels of damage and b, the number of teredinid tunnels within sections exposed to

superficial, moderate and severely damaged roots. The boundaries of the boxes indicate the 25th and

75th percentiles. The line within the box marks the median. Outliers are marked as *.

Regression for the root biscuit analyses highlighted that with increasing tissue

losses leading to root death, teredinid tunnels also increased (Figure 4.5 a). The

regrowth of tissues for root biscuits not attacked by the Teredinidae was 4.74 ± 1.6 % 125

(mean ± SE) of new tissue after twelve months. The loss of tissues for root biscuits attacked by the Teredinidae was 10.23 ± 8.6 % (mean ± SE) of root tissue loss over twelve months. The greatest numbers of teredinid tunnels observed in individual root biscuits were six and ten tunnels, associated with – 41.15 % and – 27.11 % losses of tissue over twelve months respectively.

Regression analyses for root-tissue regrowth after 12 months vs. number of teredinid tunnels for each scar treatment demonstrated that there were no correlations with the root-biscuit analyses of the superficial and moderately damaged roots.

However, a strong negative correlation was found with decreasing root-tissues and increasing teredinid tunnels with severely damaged roots (Figure 4.5 b).

4.4.3 Atypical growth of lignified tissues

A lateral out-growth of lignin-like tissue growing from the cylindrical endodermal layer was found after analysis of the moderate and severely damaged root-biscuits (Figure 4.6). Three-quarters and one quarter of all moderately and severely damaged biscuits respectively, had the atypical growth of the cylindrical lignin-rich endodermal layer.

126 a

30 Without teredinid tunnels With teredinid tunnels

0 0 2 4 6 8 10 -10 y = 5.5x + 4.8 -20 R2 (adj) = 0.34

-30

-40 Root tissue regrowth(%) after12months

-50 Number of teredinid tunnels

b 30 Superficial Intermediate Severe

0 0 2 4 6 8 10 -10 Severe damage -20 y = 6.35x + 5.81 R2 (adj) = 0.48 -30

-40 Root tissue regrowth(%) after12months

-50 Number of teredinid tunnels

Figure. 4.5 Regression analyses of root tissue loss and increasing teredinid

tunnels. a, The loss of root tissue exposed to increasing teredinid tunnels (mean ±

2 SE, R (adj) = 0.34, F1, 109 = 57.23, p = < 0.001 ) b, The loss of tissue of biscuits

from the superficial and moderately damaged roots revealed no significant

2 differences (R (adj) = 0.02, F1, 74 = 2.49, p = 0.119). The roots with severe levels of

damage had the greatest level of teredinid attack, coupled with the greatest loss of

2 tissues (mean ± SE, R (adj) = 0.48, F1, 33 = 32.03, p = < 0.001).

127

a b

Figure. 4.6 Atypical growth of new lignin-rich tissues. a, A biscuit section cut

from a prop-root exposed to a moderate level of damage. The white arrows

highlight the lateral lignin-like out-growths observed from the lignin-rich

cylindrical endodermal layer and b, A biscuit section cut from a prop-root

exposed to superficial damage (for comparison).

4.5 Discussion

The roots with a moderate level of damage, removal of the parenchyma tissues demonstrated an example of the tolerance mechanism – an over compensation of tissue regrowth (Brooks and Bell, 2002). Parenchyma cells have the ability to change; suggesting that post tissue damage will alter the cell physiology and regulate cell growth (Bloch, 1952). Parenchyma cell differentiation has the ability to change from parenchyma to cork and sclerenchyma cells (Wier et al., 1996). This study revealed similar results to those reported by Brooks and Bell (2002), demonstrating that the tolerance mechanism of the increases of tissue regrowth seen from the removal of the parenchyma tissues was remarkable recovery. Indeed, this increase of parenchyma tissue regrowth may be considered as an over-compensatory reaction of the damaged areas (Paige, 1992; Brooks and Bell, 2002). 128

Less energy investment is required by the plant needed for repair when exposed to superficial damage, when compared to the energy required by the plant to compensate for a moderate or severe level of damage, which may be exacerbated by teredinid tunnelling. Not all roots were tolerant to exposure of tissue removal.

Between fifteen and thirty percent of roots exposed to moderate and severe damage respectively demonstrated tissue degeneration and death after twelve months. This may be due to the variation and frequency of teredinid attack coupled with the severity of damage imposed upon each root.

Almost forty percent of all biscuits from roots with severe damage had a loss of nearly thirty percent of tissues. The variations of tissue regrowth between increasing severities of root damage may be due to the degree of tolerance the damaged mangrove roots were exposed to. The degree of tolerance can be expressed as compensation to the plants ability to tolerate herbivore attack (Strauss and

Agrawal, 1999). There are many extrinsic and intrinsic factors the plant needs to compensate for, such as frequency of attack from herbivores, chemical defence investment, nutrient availability and competition (Brooks and Bell, 2002). Thus, energy investment for wound healing, leading to a full recovery may vary spatially between the sites and within site, which may explain the variability of tissue regrowth between the damaged roots as the frequency of herbivorous attack from teredinids was found only in the moderate and severly damaged roots – but the greatest frequency of teredinid attack was found in the severly damaged roots.

Roots exposed to a moderate and severe level of damage demonstrated a dedifferentiation of cellular regrowth, attributed by a new lignified outgrowth from the endodermis layer. When tissue patterns arise which differ conspicuously from the norm, or when a lack of pattern is encountered the cellular regrowth can be classified 129 as atypical (Bloch, 1952). Other authors have reported the same lignified outgrowth of tissue in Rhizophora prop-roots when damaged (Wier et al., 1996). As renewed parenchyma tissue grows over the damaged area, a new formation of the radial endodermal layer is produced that appears to grow into the new cortex regrowth. The atypical growth of the new lignin-rich tissues may provide some structural rigidity for the new cellular regrowth of parenchyma cells. This adaptation may be induced by over compensation of the tolerance mechanism (Haukioja and Koricheva, 2000;

Brooks and Bell, 2002). The damaged area produces an increase of parenchyma tissues from that of the original volume recorded prior to the initial surgeries, the new outgrowth of non-radial woody endodermal layer may offer structural rigidity for the increased mass of cortex tissues as it develops. In addition, a constituent component of the endodermal layer is suberin. Mechanical wounding of the plant tissues activates genes that induce the production of proteins, such as phenylalanine ammonia lyase (PAL) (Karban and Myers, 1989; Doorn and Cruz, 2000). PAL can create the biosynthesis of phenols, such as suberin. An increase of suberin deposits produce increases of parasitic resistance or immunity to further infection in to the damaged tissues (Bloch, 1952; Karban and Myers, 1989); therefore enhancing the plants defence and increasing overall fitness and survival.

The plant’s abiotic and biotic environment such as salinity concentrations and immersion time can also affect its tolerance to herbivory (Strauss and Agrawal,

1999). All root-scar experiments within the three sites were conducted in areas of the forest where teredinid populations thrive – in areas of prolonged tidal immersion

(Rimmer et al., 1983; Kohlmeyer et al., 1995; Leonel et al., 2006).

Differences of degree of teredinid tunnelling between root-treatments were found: with increasing severities of tissue damage the level of teredinid tunnelling 130 increases. No teredinid tunnels were found in roots with superficial damage after twelve months. A few teredinid tunnels were found in the moderately damaged roots.

But, severely damaged roots had the greatest frequency of teredinid tunnels. The rapid recruitment of the teredinids found in severely damaged roots may have been facilitated by the removal of the lignin-rich endodermal layer that protects the inner vascular cylinder. Cells contained within the vascular cylinder are tannin-free (Gill and Tomlinson, 1971); making teredinid settlement possible (Turner, 1976), and thus compromising root fitness.

The morphological and physiological resilience of these unusual trees enhance their survival in mangrove habitats. This study demonstrates the regenerative capabilities of Rhizophora prop-roots, highlighting their resilience to damaged tissues. The regenerative abilities of Rhizophora prop-roots and the plant’s fitness are dependent upon the level of damage imposed to those tissues, coupled with extrinsic factors such as intensity of herbivory.

Sphaeromatid and teredinid activity is different within mangrove tissues.

Sphaeromatids may benefit the plant by inducing increases of lateral out-branching of new aerial-roots (Ribi, 1982). But, sphaeromatids are unable to bore in to woody prop-roots (Perry, 1988). Teredinids are able to attack prop-roots, and their affect within the root will frequently result in the death of that root, but only when the level of damage upon the root is severe. Thus, root death may generate woody detritus within the mangrove forests. The activity of the teredinids within woody debris creates many tunnels, which when vacant may provide niches for many fauna (Cragg and Hendy, 2010).

This study demonstrates the remarkable ability of mangrove trees to combat herbivorous activity by either deposition of anti-herbivore chemicals and/or an over- 131 compensatory of tissue regrowth. However, the tunnelling of teredinids in live damaged prop-roots in the low-intertidal regions of mangrove forests can be detrimental. These data demonstrate that Rhizophora prop-roots are able to defend against teredinid larval settlement when exposed to superficial damage, however when the loss of tissues exposes the tannin-free cells normally protected by the endodermal layers, teredinid larvae are able to settle and tunnel in to the prop-root.

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Chapter 5

HABITAT CREATION FOR ANIMALS BY TEREDINIDS

5.1 Introduction

Mangrove ecosystems have long been considered as not particularly biodiverse (Duke et al., 1998; Alongi, 2002), especially when compared with other tropical marine ecosystems such as coral reefs (Connell, 1978). However, mangrove forests provide a variety of niches to creatures that depend upon the mangrove ecosystem (Nagelkerken et al., 2008). Animals commonly found in mangrove forests include sesarmid crabs, which break down and recycle much of the leaf litter

(Robertson, 1990), beetles that processes large woody debris (LWD) in the high- to mid-intertidal areas of the forests (Feller, 2002). Thus, mangrove forests have organisms commonly found in terrestrial and aquatic habitats (Nagelkerken et al.,

2008). The structural complexity of mangrove forests provides a wide number of niches that maintain the high level of biodiversity, driven by substrata such as mangrove roots (Gratwicke and Speight, 2005).

The complexity of root structures provides cover and protection for small and juvenile fish communities (Ronnback et al., 1999; Correa and Uieda, 2008;

Wang et al., 2009), to decrease their risk of becoming predated upon (Verweij et al.,

2006; Tse et al., 2008; MacDonald et al., 2009). The structural heterogeneity provided by the roots may either impede the movement of hunting predatory fish or the prey-fish are able to reduce their visibility by using the roots to hide behind

(Laegdsgaard and Johnson, 2001; Kruitwagen et al., 2010) – therefore the diversity of animals and the abundance of individuals are largely considered to be attributed 133 by deterministic factors such as habitat complexity combined with animal recruitment rates (Syms and Jones, 2000).

Although the mangrove roots have been studied for their role as a nursery habitat, little is known about the mangrove faunal communities relying upon fallen wood as a habitat; in particular wood that has been attacked by teredinid bivalves.

Teredinids create many tunnels in LWD (Filho et al., 2008). However, when the teredinids die, the tunnels may support biodiversity when vacant for animals to exploit (Cragg and Hendy, 2010). This means that LWD may provide an ecosystem service as a micro-habitat within intertidal areas (Storry et al., 2006). When trees fall in to freshwater ecosystems they attract a high level of biodiversity. The riparian zone in freshwater ecosystems is important as this zone is habitat for many aquatic animals. Thus, LWD serves as an important link between terrestrial and aquatic ecosystems (Roth et al., 2007).

Initially, the study reported in this chapter aimed to compare the most biologically diverse substratum within the mangrove forests of the Tukang Besi

Archipelago by using biodiversity indices as an indicator of habitat importance. This study also aims to investigate the ecological role of the Teredinidae in creating niches for faunal communities within LWD. The complexity of the mangrove forest determines the habitat for many fauna (Gratwicke and Speight, 2005; Nagelkerken et al., 2010). Thus, the internal structural complexities within woody detritus created by teredinid tunnels was quantified and measured, and correlated with the level of animal diversity recorded from within each LWD sample.

To investigate the diversity of fauna within teredinid-attacked LWD samples of LWD were collected and the area occupied by teredinid tunnels within LWD was estimated as a measure of complexity. The fauna within the tunnels were removed to 134 determine if there are relationships of increasing amounts of teredinid tunnels within

LWD combined with increasing levels of faunal diversity within those vacant tunnels. Drilled wooden blocks mimicking teredinid tunnels were deployed in the mangrove forests to determine the recruitment and succession of fauna in to the tunnels.

This study tested the hypotheses that fallen wood will be the most bio-diverse substrata, when compared to the sediments and root structures within the forests. A subsidiary hypothesis tested that with increasing teredinid tunnels there will be an increasing level of animal diversity within fallen wood, and that the first animals to colonise the teredinid tunnels will have a greater mobility than animals that colonise later and the size of the tunnel will affect the species present.

5.2 Materials and methods

5.2.1 Assessing the sediment macro-infauna, root epifauna and LWD

cryptic fauna

Four mangrove forests, Kaluku, Loho, One Onitu and Gili were used to compare the biodiversity of sediment infauna, root epifauna and cryptofauna from

LWD (see chapter 2 for site descriptions). In each forest, three transects extending from the strandline and ending at the seaward edge, samples of LWD (n = 101) were collected and all cryptic macrofauna were removed. All items of LWD 2 metres each side of the transect tape were sampled. Adjacent to each LWD item in the forests, a sample of sediment (n = 101) was removed using a trowel for macro-infauna and five prop roots (n = 510) were inspected for epifauna. 135

For each sediment sample, a volume of 4000 cm3 was collected at a depth of

10 cm and placed on to a plastic sheet and all macro-infauna were carefully removed from the sediment. For each root, the length and circumference were measured to determine the surface area and all attached epifauna were identified. For each LWD sample, the volume of the wood piece was calculated. All animals were identified to the lowest achievable taxonomic level with resources available locally.

5.2.2 Assessing the effect of teredinid tunnelling on LWD in mangrove forests

Five mangrove forests were used to investigate animal communities within

LWD attacked by teredinids. These were, Langira, Kaluku, Loho, One Onitu and Gili

(see chapter 2 for site descriptions). Within each forest, LWD samples were collected from five transects extending from the strandline and extending out to the seaward edge (see chapter 2 for details). At ten metre intervals within the Langira forest, a total of 70 items of LWD were collected. Within Kaluku, Loho, One Onitu and the

Gili mangrove forests a total of 20, 44, 32 and 30 items of LWD respectively were collected at five metre intervals. A scale of teredinid attack intensity (Table 5.1) was used within the samples of LWD to categorise the level of teredinid tunnelling. The percentage of the surface occupied by teredinid tunnels was estimated across the longitudinal section of each woody detrital sample using the digital analysis package,

Image Tool Version 3.00 (The University of Texas Health Science Centre at San

Antonio). All animals were removed from each LWD sample and identified.

136

Table 5.1, The scale of teredinid-attack intensity used to categorise LWD within the mangrove forests of the Tukang Besi Archipelago. Levels 0 – 4 correspond to the percentage surface area of teredinid tunnels.

Level Teredinid tunnels in detrital samples 0 No sign of borer attack

1 Very few tunnels, covering no more than 15 % of the total area

2 No more than 25 % of the total area is comprised of tunnels

3 Integrity of wood compromised – between 25 % and 50 % of the total area are tunnels

4 Severe attack – more than 50 % of the sample is comprised of tunnels

5.2.3 The recruitment and succession of animals to drilled wooden blocks

mimicking teredinid-attacked LWD

A total of 135 drilled Bruguiera gymnorhiza blocks were exposed within three mangrove forests (Kaluku, Loho, and One Onitu) over the period from June

2009 to August 2009 to determine the recruitment and succession of fauna within tunnel-like niches. Three ten by ten metre plots were chosen at random in areas where teredinids dominate within each forest.

The volume of each test block was 512 cm3. To replicate the diameter of teredinid tunnels in naturally occurring LWD, each block had 3 x 20 mm drilled holes, 3 x 10 mm drilled holes, 3 x 5 mm drilled holes and 3 x 2.5 mm drilled holes

(Figure 5.1). All drilled holes were made using an electric drill. The blocks were attached to Rhizophora stylosa prop-roots within the three plots in each of the three forests, giving a total of forty-five blocks in each forest. Three blocks were removed from each plot every ten days over 50 days. All fauna present within the mimicked tunnels were collected and photographed. All animals were identified to the lowest achievable taxonomic resources available. 137

Figure 5.1, A drilled wooden block with various diameter tunnels to

replicate a teredinid-attacked piece of woody detritus. The wooden block

is tethered to a Rhizophora stylosa prop-root using a nylon cable-tie.

5.3 Data analysis

To test for differences of abundances of animals between the sites, a One-way analysis of variance (ANOVA) was used. A 1-way ANOVA was used to determine differences between values of Shannon-Weiner biodiversity indices calculated from root epifauna, sediment macro-infauna and LWD cryptic fauna. A 2-way ANOVA was used to test for differences of the abundance and numbers of species in drilled wooden blocks between three mangrove sites and over fifty days, with site and time as factors. Regression analyses were used to reveal relationships of faunal diversity within vacant teredinid tunnels and environmental factors, distance from the land, volume of LWD and surface area of teredinid tunnels. Tukey’s pairwise comparison test separated values into statistically distinct subsets for ANOVA. All data were 138 checked for normality, residuals were inspected to ensure that assumptions for

ANOVA were not compromised. Count data were square root transformed and all percentage data were arcsin transformed. Statistical analyses were performed using

MINITAB (MINITAB Inc, version 13.20).

Univariate and non-parametric multivariate techniques, biodiversity indices tests and a 1-way, 3-level PERMANOVA, based on a Bray-Curtis similarity matrix was used to determine differences between the sediments, roots and LWD in the

PRIMER 6 (Plymouth Routines in Multivariate Ecological Research). Post-hoc pairwise tests were applied after the PERMANOVA to highlight the similarities between the different substrata. Diversity indices were used for the fauna collected from different substrates, teredinid-attacked LWD, the fauna collected from within different diameter tunnels, and the recruitment of fauna to the tunnels (see chapters 2 and 3 for indices description).

5.4 Results

5.4.1 Comparisons of animal diversity in different substrata

Differences of the faunal communities between substrata were significant

(Table. 5.2 a, one-way PERMANOVA pair-wise test, pseudo-F 2,732 = 47.03, p =

< 0.01). The LWD and root fauna had the lowest similarity between the different

habitats with 38.5 % when compared with each of the substrata (Table. 5.2 b).

The sediments and roots were most similar with 70.5%.

139

Table 5.2 PERMANOVA comparisons of the pair-wise similarity between the fauna of the different substrata. a, the differences between the three substrata b, The

percentage similarity between sites for each substratum. a

Habitat t P (perm)

Roots, Wood 8.3 < 0.01

Roots, Sediment 3.8 < 0.01

Wood, Sediment 6.7 < 0.01

b Roots Wood Sediment

Roots 62.8

Wood 38.5 31.4

Sediment 70.5 43.5 83.0

The greatest levels of diversity (H`) and total number of species (S) were

recorded from LWD attacked by teredinids (Table. 5.3). The sediment samples had

the lowest recorded indices of diversity and numbers of species.

140

Table. 5.3, Biodiversity indices of animals recorded from roots, sediments and

LWD. S = number of species, J’ = Pielou’s evenness, H’ = Shannon-Weiner diversity

index.

Habitat Site S J' H'

Sediment Gili 5 0.84 1.36

Kaluku 3 0.95 1.04

Loho 7 0.96 1.87

One Onitu 4 0.68 0.94

Roots Gili 8 0.84 1.75

Kaluku 12 0.86 2.13

Loho 16 0.79 2.19

One Onitu 16 0.79 2.19

Wood Gili 13 0.85 2.18

Kaluku 26 0.83 2.72

Loho 28 0.84 2.81

One Onitu 21 0.86 2.62

Significant differences between the three substrata were found (Figure 5.2).

Sediments had the lowest recorded diversity and the mangrove roots had a slightly greater level of diversity, but not significantly different. However, woody detritus had the greatest recorded mean level of diversity within the mangrove forests. Fallen wood attacked by teredinids within the forests were on average 5.7 times more 141 biodiverse than the root structures and 9.3 times more diverse than the infauna within mangrove sediments.

0.9 n = 101 b 0.8

0.7 `) `)

H 0.6

0.5

Diversity ( 0.4

0.3 n = 510 n = 101 0.2 a a 0.1

0 Sediment Roots Wood Substrata

Figure 5.2 The Shannon-Weiner diversity of fauna recorded from four

mangrove sites within the Tukang Besi Archipelago. Tukey’s test reveals no

difference of biodiversity recorded between the sediments and root substrata

(letters above the bar). However, a clear significant difference was found with

biodiversity of cryptic fauna compared with other substrata from the

mangroves (mean ± SE, F2, 732 = 114.56, p = < 0.001).

The identified fauna from the roots, sediments and LWD were categorised by their trophic levels. The epifauna upon the roots consisted mainly of filter feeders.

The macro-infauna within the mangrove sediment was dominated by detritivores.

The fauna within the teredinid attacked LWD had a broad spectrum of trophic levels with carnivores, detritivores, filter feeders, herbivores and omnivores (Figure 5. 3). 142

100 90

80 Suspension feeders

70 Scavengers 60 Omnivores 50 Herbivores Filter feeders 40 Trophic level level (%) Trophic Detritivores 30 Carnivores 20 10 0 Roots Sediments LWD

Substrata

Figure 5.3 The trophic levels of fauna identified from mangrove roots, sediments and

LWD from the mangrove forests of the WMP. All epifauna, macro-infauna and cryptic

fauna identified from the roots, sediments and LWD respectively were categorised by their

feeding mode (%).

5.4.2 Biodiversity in teredinid tunnelled wood

Greater biodiversity was observed in more wood tunnelled by teredinids

(Figure 5.4 a and b, 1-way ANOVA, p = < 0.001). The Gili forest (Table 5.4 a) had a total of 30 LWD samples, and the greatest abundance of animals were found in LWD samples with a teredinid attack category of 3, with 63 animals. No level 4 teredinid attack categories were recorded in LWD from the Gili forest.

The greatest numbers of different species in Kaluku (Table 5.4 b) were found in LWD samples with a level 4 category of teredinid attack, with 25 species. No animals were recorded in the analyses of LWD within the level 0 category of 143 teredinid attack intensities. The greatest abundances of animals were also found in the level 4 categories, with a total of 315 animals.

The greatest numbers of different species in Langira (Table 5.4 c) were found in LWD samples with a level 4 category of teredinid attack, with 30 species. The greatest abundance of animals was also found within LWD samples with a level 4 category of teredinid attack, with a total of 478 animals. The lowest number of species was recorded with the level 0 category of teredinid attack, with 6 species.

The greatest numbers of different species in Loho (Table 5.4 d) were found in woody detrital samples with a level 4 category of teredinid attack, with 26 species.

The greatest abundance of animals was also recorded within LWD samples with a teredinid attack level of 4, with a total of 271 animals. No animals were found in

LWD samples with a teredind attack intensity level of 0.

The greatest numbers of different species in One Onitu (Table 5.4 e) were found in detrital samples with a teredinid attack intensity of level 4, with 18 species.

The greatest abundance of animals was also found in LWD samples with a teredinid attack level of 4, with a total of 73 animals. Whereas, the lowest recorded number of species and abundance was recorded within LWD with a level 0 category of teredind attack, with 5 species and a total of 6 animals respectively.

There is a strong relationship of increasing numbers of different species found within detrital samples of increasing percentage surface areas of teredinid tunnels (Figure 5.5 a). A similar relationship was found with the abundance of individuals with an increasing amount of teredinid tunnels. A regression analysis demonstrated that with increasing surface areas of teredinid tunnels within LWD, the abundance of animals will also increase (Figure 5.5 b). 144 a Gili Kaluku Langira Loho One Onitu 70 d

30 d

20 Abundance Abundance of individuals c abc c bc bc bc 10 b b b ab ab ab ab a a ab a a a a 0 * 0 1 2 3 4 b

10 d d 8 c

bc 6 c

c c c c c 4 bc bc Number Number species of bc ab ab ab 2 ab ab b a a a

0 * 0 1 2 3 4 Teredinid attack level

Figure 5.4, The abundance of animals (a) and numbers of species (b)

removed from teredinid-attacked LWD in mangrove forests from the

WMP. LWD from each site was categorised by the amount of teredinid

tunnels (refer to Table 5.1) and all fauna removed (mean ± SE). * = No LWD

collected with this level of teredinid attack intensity. Letters above the bars =

Tukeys Pairwise test. Table 5.4, Animals collected from teredinid-attacked LWD from five mangrove forests in the WMP. Taxonomic identification of the class, family145 and species

(where possible) of animals removed from LWD with different levels of teredinid attack. Abundances of species where classed as: * = rare, ** = common, and ***

= abundant within each category of attack. Blank spaces = none found. Total number of species, abundance and number of LWD samples analysed per category of teredinid attack are also demonstrated. Gili a Category of Teredinid attack Phylum Class Order Family Species 0 1 2 3 4 Mollusca Gastropoda Basommatophora Ellobiidae Pythia sp. A *** Cassidula sp. A * Annelida Polychaeta Aciculata Nereididae nereid spp. *** * Amphinomidae amphinomid sp. A * * Arthropoda Chilopoda Scolopendromorpha Cryptopidae centipede sp. A * ** Insecta Coleoptera Cerambycidae cerambycid sp. A *** *** *** *** Orthoptera Gryllidae Apteronemobius asahinai * Malacostraca Amphipoda Talitridae Microrchestia sp. A * * *** *** Decapoda Grapsidae Metapograpsus spp. * * * Xanthidae xanthoid spp. * * * *** Atyidae Caridina propinqua * Isopoda Cirolanidae Cirolana sp. A * *** Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus * Total species 4 6 10 8 0 Total abundance 17 10 36 63 0 Number of LWD samples 13 7 4 6 0

146 Kaluku b

Category of teredinid attack

Phylum Class Order Family Species 0 1 2 3 4 Platyhelminthes Rhabditophora Polycladida Gnesiocerotidae Styloplanocera sp. A * * Nemertea Enopla Hoplonemertea Prosorhochmidae Pantinonemertes sp. A * Mollusca Bivalvia Pterioida Isognomonidae Isognomon ephippium * * Mytiloida Mytilidae Xenostrobus sp. A *** Gastropoda Neogastropoda Marginellidae marginellid sp. A * Mitridae Strigatella sp. A * Littorinimorpha Assimineidae assimineid sp. A * Systellommatophora Onchidiidae Onchidium nigram * Heterostropha Pyramidellidae pyramidellid sp. A * Sipuncula Phascolosomatidea Phascolosomatida Phascolosomatidae Phascolosoma arcuatum * Annelida Polychaeta Aciculata Nereididae nereid spp. ** Amphinomidae amphinomid sp. A ** *** Eunicida Eunicidae eunicid sp. A * *** * *** Terebellida Terebellidae terebellid spp. ** Arthropoda Arachnida Araneae Desidae Desis martensi * Chilopoda Scolopendromorpha Cryptopidae centipede sp. A * Insecta Coleoptera Cerambycidae cerambycid sp. A * * Malacostraca Amphipoda Talitridae Microrchestia sp. A *** Decapoda Grapsidae Metapograpsus sp. A * * * *** Xanthidae xanthoid spp. * Atyidae Caridina propinqua ** * Alpheidae Alpheus sp. A * * Diogenidae Diogenes sp. A * Isopoda Cirolanidae Cirolana sp. A ** *** *** Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus * Total species 0 4 2 9 25

Total abundance 0 8 6 25 315

Number of LWD samples 3 4 2 3 8

147 Langira c

Category of Teredinid attack intensity

Phylum Class Order Family Species 0 1 2 3 4 Platyhelminthes Rhabditophora Polycladida Gnesiocerotidae Styloplanocera sp. A * Nemertea Enopla Hoplonemertea Prosorhochmidae Pantinonemertes sp. A * Mollusca Bivalvia Mytiloida Mytilidae Xenostrobus sp. A * * Pterioida Isognomonidae Isognomon ephippium * Gastropod Littorinimorpha Assimineidae assimineid sp. A * * Neotaenioglossa Cypraeidae Cypraea interrupta * Systellommatophora Onchidiidae Onchidium nigram * * Neogastropoda Mitridae Strigatella sp. A * Muricidae Thais gradata * Nassariidae Nassarius sp. A * Columbellidae Pseudanachis basedowi * Caenogastropoda Cerithiidae Clypeomorus sp. A * Cephalopoda Octopoda octopod sp. A * Sipuncula Sipunculidea Phascolosomatida Phascolosomatidae Phascolosoma arcuatum * * Annelida Polychaeta Aciculata Nereididae nereid spp. * * *** ** *** Amphinomidae amphinomid sp. A ** * *** Eunicida Eunicidae eunicid sp. A * ** * ** Terebellida Terebellidae terebellid spp. * ** Arthropoda Arachnida Araneae Desidae Desis martensi * ** Insecta Coleoptera Cerambycidae cerambycid sp. A *** ** * ** * Malacostraca Amphipoda Talitridae Microrchestia sp. A * Decapoda Grapsidae Metapograpsus sp. A * * *** ** ** Xanthidae xanthoid spp. ** * * ** Atyidae Caridina propinqua * * * Alpheidae Alpheus sp. A * * * Diogenidae Diogenes sp. A * * Pilumnidae Pilumnus vespertilio * * Isopoda Cirolanidae Cirolana sp. A * * ** * * Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus * * * Anguilliformes Muraenidae Gymnothorax richardsoni *

148

Echinodermata Echinoidea Echinoida Urchin sp. A * Ophiuroidea Ophiurida Amphiuridae Ophiactis sp. A * * Total species 6 13 14 8 30 Total abundance 16 45 34 26 487 Number of LWD samples 8 10 9 4 39 d Loho

Category of Teredinid attack intensity Phylum Class Order Family Species 0 1 2 3 4 Porifera Homoscleromorpha Porifera sp. A * Platyhelminthes Rhabditophora Polycladida Gnesiocerotidae Styloplanocera sp. A * * * Nemertea Enopla Hoplonemertea Prosorhochmidae Pantinonemertes sp. A * Annelida Polychaeta Aciculata Nereididae nereidae spp. ** *** *** *** Amphinomidae amphinomid sp. A * * * ** Eunicida Eunicidae eunicid sp. A * * * Terebellida Terebellidae terebellid spp. * * ** Canalipalpata Sabellidae sabellid sp. A * Arthropoda Arachnida Araneae Desidae Desis martensi * * Chilopoda Scolopendromorpha Cryptoidae centipede sp. A * Insecta Coleoptera Cerambycidae cerambycid sp. A ** * * * Malacostraca Amphipoda Talitridae Microrchestia sp. A * ** ** Decapoda Grapsidae Metapograpsus sp. A * * * ** Xanthidae xanthoid spp. * * * Alpheidae Alpheus sp. A * * Isopoda Cirolanidae Cirolana sp. A * ** *** Mollusca Bivalvia Mytiloida Mytilidae Xenostrobus sp. A ** * * *** Pterioida Isognomonidae Isognomon ephippium * Gastropoda Littorinimorpha Assimineidae assimineid sp. A * * Neotaenioglossa Cypraeidae Cypraea interrupta * Thiaridae Melanoides sp. A *

149

Neogastropoda Nassariidae Nassarius sp. A * Marginellidae marginellid sp. A * Polyplacophora Chiton sp. A * Cephalopoda Octopoda octopod sp. A * Sipuncula Sipunculidea Phascolosomatida Phascolosomatidae Phascolosoma arcuatum * Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus ***

Total species 0 7 12 12 26

Total abundance 0 24 38 79 271

Number of LWD samples 4 10 7 9 14 e One Onitu

Category of Teredinid attack intensity Phylum Class Order Family Species 0 1 2 3 4 Mollusca Gastropoda Littorinimorpha Assimineidae assimineid sp. A * Neogastropoda Marginellidae marginellid sp. A * *

Muricidae Thais gradata *

Systellommatophora Onchidiidae Onchidium nigram *

Heterostropha Pyramidellidae pyramidellid sp. A *

Sipuncula Phascolosomatidea Phascolosomatida Phascolosomatidae Phascolosoma arcuatum *

Annelida Polychaeta Aciculata Nereididae nereid spp. *

Amphinomidae amphinomid sp. A *** ***

Eunicida Eunicidae eunicid sp. A * *

Terebellida Terebellidae terebellid spp. * * **

Arthropoda Arachnida Araneae Desidae Desis martensi * * *

Insecta Coleoptera Cerambycidae cerambycid sp. A * * *** *

Malacostraca Amphipoda Talitridae Microrchestia sp. A *

150

Decapoda Grapsidae Metapograpsus sp. A * * *

Xanthidae xanthoid spp. * * *

Alpheidae Alpheus sp. A *

Atyidae Caridina propinqua *

Isopoda Cirolanidae Cirolana sp. A * *** * **

Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus ** *

Blenniidae blennie sp. A *

Ascidiacea Phlebobranchia Tunicate sp. A *

Total species 5 7 4 5 18 Total abundance 6 12 17 10 73

Number of LWD samples 11 6 4 3 8

151

a 14

Number species of Number 2 y = 0.07x + 1.09 R² (adj) = 0.45 0 0 20 40 60 80 100 % of section tunnelled

b 160

140

120 100 y = 0.25x + 2.17 80 R² (adj) = 0.32 60 40

20 Number individualsof Number 0 0 20 40 60 80 100 % of section tunnelled

Figure 5.5 Numbers of different species and abundance of animals in teredinid-

attacked LWD. a, The number of different species removed from detrital samples (R2

(adj): 0.45, F1,193 = 159.3, p = < 0.001) and b, the abundance of individuals of cryptic

fauna removed from detrital samples categorised by the level of teredinid attack intensity

2 (R (adj): 0.32, F1,193 = 93.3, p = < 0.001).

Regression analyses were used to determine the factors that contribute to increasing numbers of different species within LWD. No difference was found with 152 animals in wood with increasing distance from the land (P = 0.908) and no difference was found with the area of the mangrove forest (P = 0.816). A significant difference was found with increasing volume of LWD when correlated with increasing numbers of different species (F1, 193 = 8.12, p = < 0.01), although a low correlation was determined R2 (adj) = 0.04. The percentage surface area of teredinid tunnels was also

2 significant (F1, 193 = 159.34, p = < 0.001), however a strong correlation was found, R

(adj) = 0.45.

The greatest level of teredinid attack within LWD from One Onitu, Loho,

Langira and the Kaluku forests had the highest recorded diversity (Table 5.5). No

LWD samples were located in the Gili forest with a level 4 category of teredinid attack. LWD samples with increasing levels of teredinid attack had lower evenness values (J`) compared with detrital samples with a lower level of teredinid attack.

LWD samples with a percentage surface area of teredinid tunnels between 0 –

19 % had the lowest levels of diversity (H’) with 0.35 ± 0.06 (mean ± SE, n = 97,

Figure 5.6). No difference of diversities were found in detrital samples with 20 – 35,

50 – 65 and more than 65 % surface area of teredinid tunnels, although there was a trend with increasing percentage surface area of teredinid tunnels and increasing diversity. Detrital samples with more than sixty percent of teredinid tunnels have on average 4.5 times more species, 6.8 times more individuals and 4 times the diversity when compared to detrital samples with twenty percent or less of teredinid tunnels.

153

Table 5.5, Biodiversity indices of fauna removed from LWD in the WMP. S

= number of species, N = total individuals, J’ = Pielou’s evenness, H’ =

Shannon-Weiner diversity index.

Site Attack level S N J' H' One Onitu 0 5 5 0.99 1.59 1 7 9 0.97 1.88 2 4 8 0.94 1.31 3 5 6 0.95 1.52 4 18 29 0.97 2.80 Loho 0 0 0 0 0 1 7 12 0.97 1.89 2 12 20 0.97 2.40 3 12 27 0.94 2.34 4 26 68 0.93 3.03 Langira 0 6 9 0.93 1.67 1 13 22 0.96 2.47 2 13 19 0.96 2.47 3 8 14 0.97 2.02 4 29 91 0.92 3.09 Kaluku 0 0 0 0 0 1 4 6 0.99 1.37 2 2 3 0.98 0.68 3 9 14 0.96 2.12 4 25 68 0.92 2.95 Gili 0 4 7 0.85 1.18 1 6 7 0.98 1.75 2 10 17 0.96 2.20 3 8 19 0.90 1.88 4 0 0 0 0

154

1.8

1.6 bc 1.4 bc 1.2 b 1

0.8 Weiner Weiner diversity (H`) - 0.6 a 0.4

Shannon 0.2 0 < 20 20 - 35 35 - 50 50 - 65 > 65

% of section tunnelled

Figure 5.6, The diversity of animals recorded from detrital samples categorised by the

level of teredinid-attack intensity. Faunal diversity indices are expressed by the Shannon-

Weiner index (mean ± SE, 1-way ANOVA, diversity vs. % surface area of teredinid

tunnels, F4,190 = 24.83, p = < 0.001). Letters above the bars = Tukey’s comparisons.

To determine if the volume of LWD had influenced the biodiversity, all LWD samples were standardised to 1 m3 to remove the variable of LWD volume. The abundance of individuals was tested (Figure 5.7). The abundance of animals was lower in LWD per cubic meter with less than 20 % surface area of teredinid tunnels.

No significant difference was found with the abundance of animals from LWD samples with levels of teredinid attack intensity increasing from 20 - 35 % to more than 65 % surface area of teredinid tunnels. However, LWD with more than sixty- five percent surface area of teredinid tunnels will have 7.4 times the individuals compared to wood of the same volume with twenty percent or less surface area of teredinid tunnels.

155

90000

) 80000 c 3 70000 60000 50000 b 40000 30000 Animals Animals LWD in (m b b 20000 a 10000 0 < 20 20 - 35 35 - 50 50 - 65 > 65 % of section tunnelled

Figure 5.7 Variation in numbers of infauna with degree of teredinid attack

collected from LWD (metre3). Abundance of individuals removed from LWD

categorised by the level of teredinid attack intensity standardised to 1m3(mean ± SE,

F4,190 = 9.3, p = < 0.001).

LWD with higher levels of teredinid-attack had greater numbers of different species, and abundance of individuals (Figure 5.8). The curve representing the level

4 category of teredinid attack begins to level-out.

156

35 Level 0

30 Level 1

25 Level 2

20 Level 3 Number Number species of

15 Level 4

0 0 200 400 600 800 1000 1200 Total number of individuals

Figure 5.8, Species accumulation curves of infauna removed from LWD categorised by

the level of teredinid attack. Infauna removed from all levels of teredinid-attacked LWD

within five mangrove forests. Each level of teredinid attack is categorised by different coloured

symbols.

The Gili and One Onitu forests had the least abundance and number of species removed from teredinid-attacked LWD of the five mangrove forests.

However, Langira, Loho and Kaluku have the greatest abundance and number of species within teredinid-attacked LWD (Figure 5.9). 157

Abundance Species 800 35

700 30

600

500 20 400 15

300 Totalabundance 10 Number species of 200

100 5

0 0 Gili One Onitu Langira Loho Kaluku

Site

Figure 5.9, The total abundance and number of species removed from teredinid-attacked

LWD from each mangrove forest. Animals were removed from teredinid-attacked LWD and

counted.

The degree of teredinid-tunnelling in LWD samples within each mangrove forest was considered (Figure 5.10). Teredinid attack levels in LWD in four of the forests were statistically indistinguishable, while the level of attack in Gili was significantly lower than Langira, Loho and Kaluku.

158

45 b

35 b b 30 25 ab 20 a

15 % % of SA teredinid tunnels 10 5 0 Gili One Onitu Langira Loho Kaluku Site

Figure 5.10, The variation between five mangrove forests in degree of

teredinid tunnelling. The percentage surface area of teredinid tunnels within each

piece of LWD from each forest (mean ± SE1-way ANOVA, teredinid attack

intensity vs. site highlighted significant differences, F4,191 = 6.52, p = < 0.001).

Letters above the bars indicate a Tukey’s comparison test.

5.4.3 Recruitment and succession of animals within drilled wooden blocks mimicking teredinid tunnels

The removal of animals within Kaluku, Loho and the One Onitu forests from drilled wooden blocks (Table 5.6) demonstrated that as time increases, the numbers of species recorded from the drilled wooden blocks also increase. After 10, 20, 30,

40 and 50 days, 3, 6, 11, 8 and 11 species were recorded, respectively. The earliest recruiting animals were the grapsoid crabs, attributed by their high abundance after ten days. Suspension feeders (the bivalve Xenostrobus sp. and tunicates) only appeared after fifty days.

No significant differences of the numbers of species and individuals were found between the three sites (2-way ANOVA, F2, 128 = 1.89 and 2.53, p = 0.155 and 159 p = 0.084 respectively). However, significant differences were found with the numbers of species and abundance of individuals recruiting to the drilled wooden blocks mimicking teredinid tunnels over the fifty days (F4, 128 = 5.29, and 6.59, p = <

0.01 and p = < 0.001 respectively).

Day ten had the lowest recorded number of different species (Figure 5.11 a, mean ± SE, n = 27). The greatest numbers of different species were recorded from day forty. The abundance of animals recorded from the three forests also increased with time (Figure 5.11 b), with the greatest abundance also found at day forty.

a 1.8 b

1.6 ab 1.4 ab 1.2 a 1 a 0.8 0.6

Number Number species of 0.4 0.2 0 Day 0 10 Day 20 Day 30 Day 40 Day 50 Time (d) b 3

2.5 b b

2 ab 1.5 a a 1

0.5 Abundance Abundance of animals 0 Day 0 Day 10 Day 20 Day 30 Day 40 Day 50 Time (d)

Figure 5.11, Numbers of species and individuals removed from drilled wood blocks

over an exposure of 50 days. a, The number of species (1-way ANOVA, number of

species vs. time: F4, 130 = 5.21, p = < 0.01), and b, abundance of individual animals

removed from drilled wooden blocks over ten day intervals in the Kaluku, Loho and One

Onitu forests (mean ± SE, n = 27, 1-way ANOVA, abundance vs. time; F4, 130 = 6.44, p =

< 0.001). Letters above the bars are the results from a Tukey’s test. 160

Table 5.6, Animals collected from drilled wooden blocks in three mangrove forests in the Tukang Besi over a period of 50 days after installation. All animals were removed every 10 days from drilled holes in wooden blocks. Species abundance; *** = very common, ** = common, * = rare, blank cells = none found.

PHYLUM CLASS ORDER FAMILY SPECIES Day 10 Day 20 Day 30 Day 40 Day 50 Porifera Homoscleromorpha Porifera sp. A * * Mollusca Gastropoda Littorinimorpha Assimineidae assimineid sp. A * * Caenogastropoda Cerithiidae cerithiid sp. A * * * Basommatophora Ellobiidae ellobiid sp. A * * Heterostropha Pyramidellidae pyramidellid sp. A *

Cephalopoda Octopoda octopod sp. A * Bivalvia Mytiloida Mytilidae Xenostrobus sp. A * Annelida Polychaeta Aciculata Nereididae nereid spp. A ** * * Arthropoda Malacostraca Decapoda Alpheidae Alpheus sp. A * * * * Diogenidae Diogenes sp. A ** * *** ** Grapsidae grapsoid spp. *** *** *** *** *** Xanthidae xanthoid spp. * Isopoda Cirolanidae Cirolana sp. A * ** * * * Amphipoda Talitridae Microrchestia sp. A ** ** *** ** Chordata Actinopterygii Perciformes Ptereleotridae Parioglossus interruptus * Ascidiacea Phlebobranchia Tunicate sp. A * Total species 3 6 11 8 12

161

The biodiversity indices of the fauna removed from the drilled wooden blocks demonstrates that the evenness increases rapidly and then drops at day 30, meaning that the dominance of a particular species has increased (Figure 5.12). The biodiversity increases, as time increases.

Evenness Shannon Weiner

0.98 2.5

0.96 2

0.94 1.5

0.92

Weiner Weiner diversity - Evenness 1 0.9

0.88 0.5 Shannon

0.86 0 Day 10 Day 20 Day 30 Day 40 Day 50 Number of days

Figure 5.12, Community indices of fauna removed from drilled holes in wooden

blocks over time. Evenness is a measure of the faunal community structure and the

Shannon-Weiner diversity index measures the level of biodiversity.

In the fifty day experiment, within each site the most dominant species recruiting to the wooden blocks were grapsoid crabs, demonstrated by their early recruitment in each site (Figure 5.13). The high dominance of the Grapsidae within the sample units is further illustrated by their initial high peaks of the rank abundance curves that rapidly taper off.

The least dominant species to occur from the sample units differed between each of the three sites. The bivalve, Xenostrobus sp. were the least abundant and least dominant species collected from the sample units in the Kaluku forest, found only at day 50 of the experiment. The least abundant and least dominant species recorded 162 from the Loho forest were tunicates, also observed at day 50. The least dominant species recorded from the One Onitu forest within the sample units were the detritivore isopods, Microrchestia sp., also found at day 50.

The first and most dominant animals to recruit in to the drilled wooden blocks were; Grapsidae>Taltridae>Diogenidae>Cirolanidae> Alpheidae (Malacostraca)

>Nereididae (Polychaeta) >Cerithiidae>Ellobiidae = Assimineidae (Gastropoda)

>Porifera >Tunicate (Ascidiacea) = Mytilidae (Bivalvia). The order by which the animals are arranged are categorised by decreasing dominance. The most dominant species recorded from the sample blocks belonged to the Malacostraca, a highly motile class of animals such as grapsoid crabs, Microrchestia sp., mud hoppers and cirolaniid isopods, fish lice. The least dominant species recorded were Xenostrobus sp., sessile filter feeding bivalves. Of the motile fauna recorded, the least dominant class of animals were the Gastropoda, such as the slow-moving ellobiids and assimineids.

The number of different species and diversity of animals increased with increasing diameter of tunnels (Table 5.7). Drilled tunnels 2.5 mm in diameter had the lowest number of species and diversity of fauna. However, the greatest numbers of species and diversity were recorded from tunnels measuring 10 mm diameter, with

12 species. Species commonly found in 2.5 mm and 5 mm diameters drilled holes were cirolanids and Microrchestia sp. The most common species found in 10 mm and 20 mm diameter drilled holes were grapsoids.

Figure 5.13, The frequency and rank abundance of species within drilled wooden blocks within the mangrove forests. Numbers of animals are the total abundance recorded from the drilled wooden blocks within each forest. Pages 159 – 161.

163

Kaluku

10 days

Species

Abundance

20 days

Abundance

Species

30 days

Species

Abundance

40 days

Species

Abundance

50 days

Abundance

Species

Number of individuals Species rank

164

Loho

10 days

Abundance

Species

20 days

Species

Abundance

30 days

Species

Abundance

40 days

Species

Abundance

50 days

Species

Abundance

Number of individuals Species rank 165

One Onitu

10 days

Abundance

Species

20 days

Abundance

Species

30 days

Species

Abundance

40 days

Species

Abundance

50 days

Species

Abundance

Number of individuals Species rank

166 Table 5.7, Biodiversity indices of fauna removed from different diameter drilled tunnels after 50 days . S= number of species, J’ Pielou’s evenness, H’ = Shannon-Weiner diversity index.

Tunnel Diameter Number of Secies J' H'

2.5 1 0 0

5 5 0.9 1.5

10 12 0.9 2.2

20 10 0.9 2.0

5.5 Discussion

Within the mangrove forests of the Tukang Besi Archipelago very few

animals were found in the sediments, mainly consisting of Uca spp. Although

mangrove sediments are generally anoxic, fiddler crabs bioturbate the sediments

(Bouma et al., 2009) and even enhance the growth of the mangroves (Smith et al.,

2009). Burrowing by fiddler crabs increases the growth of Caribbean trees and

decreases pore-water salinity (Smith et al., 2009). Upon the mangrove roots, a large

proportion of the epifauna were sessile filter-feeders, sponges, mussels and oysters.

Mangrove roots with epifauna have been found to enhance fish abundance

(MacDonald et al., 2008). Furthermore, sponges not only take-up much of the

available hard surfaces of the roots, they also provide an important ecosystem service

by protecting the roots from herbivorous attack by isopods (Ellison and Farnsworth,

1990; Ellison and Farnsworth, 1992). The root epifauna within the Gili mangrove

forest was distinct from the root epifauna recorded from all other sites. Gili is

established further inland compared to the other mangrove forests and no sessile

filter-feeding animals were found as tidal conditions were unsuitable for sessile

filter-feeding invertebrates to colonise. Mangrove roots within Gili had a greater

number of motile fauna ranging from grapsoid crabs to potamid gastropods. The 167

LWD fauna-taxon and density was the richest compared with the other substrata.

LWD is more structurally complex compared to the sediments and roots. A greater faunal richness is found with increasing habitat complexity (Taniguchi and Tokeshi,

2004).

The fauna collected from the teredinid-attacked LWD in the mangrove forests were recruited during sporadic episodes of colonisation, but maintained by a habitat- specific structure – teredinid tunnels inside the LWD. Non-deterministic environmental interactions may also have an effect upon site-specific animal communities (Poff and Allan, 1995). For example, hydrological processes shape waterways, by creating random processes that characterise and structure the animal assemblages (Poff and Allan, 1995; Grossman et al., 1998). Thus, the animals recorded from each site from this study may have varied due to the intricate small tidal channels and streams within the forests.

Species diversity is positively correlated with increasing habitat complexity

(Petren and Case, 1998; Saha et al., 2009), such as is created by teredinids. LWD within aquatic ecosystems are important constituents in the structure and complexity of water ways and LWD will probably become colonised by a wide variety of animals (Hilderbrand et al., 1997; Storry et al., 2006). Thus, LWD represents a vital resource providing a spatial subsidy within the ecosystem (Polis et al., 1997). Wood enhances habitat complexity (Shirvell, 1990; Brooks et al., 2004), the deposition of sediments and the retention of organic matter (Bilby and Likens, 1980; Smock et al.,

1989). Thus, removal of wood increases sediment discharge and leads to a reduction of habitat structural complexities (Larson et al., 2001; Brooks et al., 2004). LWD is therefore an important component within aquatic ecosystems (Shields et al., 2006).

Thus, a major threat to the mangrove ecosystem is wood harvesting (Valiela et al., 168

2001; Sanchirico and Mumby, 2009), which could reduce ecosystem-level mangrove faunal diversity due to the reduction of wood (Benke et al., 1985; Wright and

Flecker, 2004).

Indeed, the results from this study corroborate previous research that woody detritus does increase animal diversity (Everett and Ruiz, 1993; Wright and Flecker,

2004). However, this study links the role of the Teredinidae with increasing ecosystem-level biodiversity. Teredinids create niches as they consume LWD, by creating tunnels for a wide range of animals when those tunnels become vacant. As the number of teredinid tunnels increase within LWD, the abundance and diversity of animals also increase. The increasing species diversity with increasing numbers of teredinid tunnels may likely be due to a higher proportion of refuges from predation

(Willis et al., 2005). The refuge provided by woody detritus exploited by grass shrimp significantly reduces their risk of predation (Everett and Ruiz, 1993).

Vulnerable animals have more options for avoiding and escaping potential predation in habitats containing a greater number of niches. Structurally complex habitats may reduce visual contact, encounter rates and aggressive behaviour between competitors

(Jones et al., 2001; Willis et al., 2005).

The results from this study magnify the contribution of woody detritus in mangrove ecosystems to biodiversity maintenance. Mangrove forest biodiversity is significantly enhanced by a large volume of teredinid attacked LWD. Non-tunnelled

LWD has a limited number of species and abundance of individuals due to the reduced niche availability. A lack of teredinid tunnels within woody detritus will create a reduced habitat complexity that may likely increase predator-prey encounters. Differences of spatial structure will influence the frequency of interactions such as predation or niche exploitation for animals (Warfe and Barmuta, 169

2004; Nurminen et al., 2007). This may also be the case for the spatial structure teredinid tunnels provide in LWD, which explains the sharp change in animal assemblages and diversity of animals in LWD without tunnels when compared to

LWD that have been attacked by the Teredinidae. These data clearly illustrate that the faunal assemblages change and the diversity is increased when the surface area of

LWD increases, attributed by the activity of the Teredinidae.

When considering species diversity, the Shannon-Wiener Index is the most frequently used diversity indices when measuring the fauna of benthic ecology (Reiss and Kroucke, 2005). Shannon-Wiener combines species richness with the community evenness. However, this index is dependent on the sample size

(Magurran, 2004), as larger samples are likely to contain additional resources and therefore increased species. To effectively rule out the factor of sample size from this study all LWD samples were calculated to the same volume. Teredinid tunnels significantly enhance the animal diversity within detrital samples of the same volume.

The Teredinidae also influence faunal diversity on whole ecosystem level in the WMP. The Gili forest had the lowest level of teredinid attack on LWD, and the lowest level of cryptic biodiversity of the sites investigated. Kaluku and Loho had the highest degree of teredinid tunnelling in LWD. This study demonstrates that, although LWD is essential for the biodiversity of both the specialist and more generalist animals (Hilderbrand et al., 1997; Kappes et al., 2009); teredinid tunnels will increase the internal structural complexity within LWD and the tunnels significantly enhance ecosystem-level biodiversity within these particular Indonesian mangrove forests. 170

Teredinid tunnels are home to many animals which cannot bore into the very hard, undecayed wood. It is for this reason that the Teredinidae can be classed as ecosystem engineers.

The colonisation and recruitment of fauna in to the drilled wooden blocks revealed that as time increased the number of different species and number of individuals colonising also increased. The initial low diversity of animals recorded from day ten may be due to a lack of microorganism colonisation and thus a lack of microbial biofilms (Gulis et al., 2004) or that it takes time for organisms to find the wood. The formations of microbial biofilms have been linked with the colonisation and presence of invertebrates (Spanhoff et al., 2001; Gulis et al., 2004). As biofilms increase, invertebrates will colonise (Tank and Winterbourn, 1995; Tank and

Wintrbourn, 1996; Collier and Halliday, 2000).

Other factors may also determine the rate at which animals colonise new habitats, by means of their mobility threshold (Arsuffi and Suberkropp, 1989), expressed as the foraging range of an animal deemed safe within the limits of potential predation risks and foraging for food. Animals are likely to be less selective feeders when food items and other resources are widely dispersed relative to their foraging range because the risk of predation increases as distance between food items become greater (Schoener, 1971). Thus, less mobile animals will likely be the least frequent to colonise new habitats. Animals with a greater mobility are found to be more selective about the quality of food items than their less mobile congeners.

However, mobile species become less selective when their mobility threshold is breached (Karieiva, 1982).

The invertebrate fauna examined by Arsuffi and Suberkropp (1989) ranked fauna by their mobility capabilities, they demonstrated that the more mobile the 171 forager, the greater its ability to be more discriminate (Arsuffi and Suberkropp,

1989). The recruitment of fauna to drilled wooden blocks also revealed a mobility gradient, with the most common recorded species from the drilled wooden blocks being the most mobile to the least common fauna being sessile. Thus, more mobile species will have a greater chance of occupying the available tunnel niche.

The level of complexity also had an influence upon the colonisation of the drilled wooden blocks. This was demonstrated by the larger holes within the drilled blocks, as they had the greatest number of species and diversity. It is likely that the increased surface area of tunnels were less difficult to locate compared to smaller tunnels.

The tunnelling of teredinids within the mangrove forests of the Tukang Besi provides many niches within LWD. Thus teredinids increase microhabitat heterogeneity due to their tunnelling action. It is the behaviour of the teredinids that create a high complexity of tunnels that lead to a broad range of co-existing animals within LWD. The activity of the Teredinidae within Indonesian mangrove forests will increase the faunal diversity at whole ecosystem level and the fauna from the teredinid-attacked LWD range in animals found in tropical terrestrial forests to subtidal marine habitats in the mangroves from the WMP. Spatial heterogeneity is a fundamental property of the natural world (Kostylev et al., 2005) and heterogeneity within an ecosystem is a vital component for the interaction of co-existing animals

(Petren and Case, 1998; Gratwicke and Speight, 2005). By comparison structurally simple habitats are not able to support the same levels of biodiversity when compared with habitats consisting of high levels of complexity and rugosity (Levin,

1992).

172

Chapter 6

BIODIVERSITY MAINTENANCE BY VACANT TEREDINID TUNNELS

6.1 Introduction

In mature mangrove forests there are considerable quantities of large woody debris (LWD) in the intertidal zone which are tunnelled by teredinids (Cragg, 2007;

Filho et al., 2008). These animals are prolific excavators of wood and through their tunnelling they influence the flux of carbon in mangrove forests by the amount of wood they process (Robertson and Daniel, 1989a; Filho et al., 2008). Within the logs the numbers of teredinids can be huge (Distel, 2003). Teredinids will create a complex mass of tunnels (Figure 6.1 a). The tunnels are cylinders that range in size and diameter which taper to a small opening at the wood surface. However, many openings emerge when wood is heavily tunnelled and the surface is broken open.

Teredinid death may support biodiversity as teredinids leave niches for animals to exploit the vacant tunnels (Cragg and Hendy, 2010).

In general, mangrove forests are known to support biodiversity as an ecosystem service (Kathiresan and Bingham, 2001; Nagelkerken et al., 2008), provided by the huge amount of wood. Prop-roots belonging to the mangrove trees

Rhizophora provide cover to many vulnerable and small fish, which are drawn to the complexity of the roots to decrease their risk of predation (Verweij et al., 2006).

However, at low tide this refuge may not be available. Animals living within teredinid-attacked LWD probably do so to avoid predators. The main purpose for seeking refuge for animals in cryptic habitats is to avoid predation (Ruxton et al., 173

2004). Many animals seek refuge in teredinid-attacked LWD in the Wakatobi mangrove forests (Cragg and Hendy, 2010). Two examples of animals found in vacant teredinid tunnels were of particular interest, a fish Parioglossus interruptus, and a spider Desis martensi.

In a study of the animals living within mangrove wood on islands of the

Tukang Besi Archipelago, Indonesia, the interrupted dartfish, Parioglossus interruptus (Suzuki and Senou, 1994) and the intertidal spider, Desis martensi (L.

Koch, 1872) were commonly found within teredinid tunnelled wood (Figure 6.1 b and c, Cragg and Hendy, 2010). Interrupted dartfish are also known from Iriomote

Island, Japan (Suzuki and Senou, 1994) and New Guinea (Rennis and Hoese, 1985).

Dartfish (Ptereleotridae) are bathypelagic gobioids inhabiting warm, temperate to tropical tidal estuaries and mangrove forests throughout the Western Pacific and

Indian Oceans (McDowall, 2001). These fish typically occupy structurally complex habitats such as seagrass beds or mangrove roots that provide protection from predators as well as access to the zooplankton upon which they feed (Keith et al.,

2004). Desis martensi (Desidae) are known from Singapore. These spiders are thought to forage for food at night and seek refuge in rock crevices and shells to avoid tidal immersion (The Singapore Red Data Book, 2008). Desids have a wide distribution spanning from the Galapagos to the South China Sea (Roth, 1967).

Both of these species are interesting due to their unusual nature; an aquatic species, that frequently experiences episodes of emersion and an intertidal species that frequently experiences episodes of immersion. Fish follow the ebbing tide out

(Sheaves, 2005) and wetland spiders migrate to higher elevations (Adis and Junk,

2002) to avoid emersion and immersion; however P. interruptus and D. martensi 174 remain within the Tukang Besi mangrove habitats throughout the complete tidal cycle.

The mangrove forests belonging to the islands of the Tukang Besi

Archipelago experience large diel changes in water temperature, oxygen levels and salinity (Taylor et al., 2005; Cragg and Hendy, 2010). Such rapid changes of salinity and temperature are common with intertidal mangrove forests (Tomascik et al.,

1997). These environmental fluctuations may be buffered within the teredinid tunnel niche. Dartfish and spiders may seek refuge within tunnelled wood for protection from high temperatures, while avoiding desiccation as well as predation. This study examined whether the niche provided by teredinid-tunnelled wood utilised by dartfish and desids is critical for their distribution. This study tested the hypotheses that dartfish and spiders 1. reduce their risk of predation by taking refuge in teredinid-tunnelled wood at low tide, and 2. are protected by the tunnels from extreme temperatures. This study describes a previously unreported cryptic, wood- dwelling behaviour for dartfish and desids. In addition, this study also reveals the biodiversity maintenance provided by the cryptic habitat of vacant teredinid tunnels and assesses the significance of this behaviour in the autecology of the fish and spiders.

Samples of LWD from the Langira forest were collected to determine the abundance and distribution of fish and spiders during low tides. Numbers of dartfish not in wood were quantified at high tide and foraging desids were searched for during day and night time hours. Size comparisons of dartfish collected from within wood and dartfish caught at hightide, not in wood were determined. Measurements of dartfish and desid diameters and tunnels were compared to establish if they have a close geometric fit within the tunnels. 175

Laboratory experiments investigating fish and spiders were conducted to provide insights of their behaviour and ecology. Individual spiders were observed in plastic tubs to determine if they emerge from tunnels during day time and night time hours. Spiders were coaxed in to glass tubes to observe their leg arrangement when inside teredinid tunnels. When inside the glass tubes, the spiders were immersed in seawater to demonstrate the water-tight properties of their webs. Dartfish were observed in aquaria that contain the calcareous linings of teredinid tunnels to determine their behaviour within those tunnels. In addition, fewer tunnels were placed into each aquarium than dartfish to determine if the fish bunch within the tunnels. A range of different diameter tunnels were used in each aquarium to establish if dartfish behaviour changes with increasing area of cryptic habitat. Using data loggers, temperature was recorded within woody detritus. Ambient air temperatures and water temperatures adjacent to each detrital sample containing temperature loggers were also recorded. Temperature cooling in wood was examined using a thermometer and a wind tunnel. To compare fluctuations of the temperature data, inside and outside LWD, non-invasive temperature tolerance experiments were used for P. interruptus. Critical Thermal Maxima (CTM) of P. interruptus determined if their physiological temperature tolerance is suited to the changes of abiotic temperatures within the mangrove environment.

6.2 Materials and methods

6.2.1 Field studies

Samples of LWD, defined as wood pieces more than 20 mm in diameter, were collected from all LWD encountered along eight transects in the Rhizophora stylosa-dominated mangrove forest, Langira, Kaledupa Island. The transects extended perpendicularly from the strandline extending out to the seaward fringe for 176 a mean distance of 380 m. A total of 196 pieces of LWD were collected from the transects. These were split lengthwise and photographed to estimate the proportion of wood removed by teredinid tunnelling, then broken open to remove all fauna occupying teredinid tunnels. The same transects were used to estimate the abundance and distribution of Desis martensi not in LWD during day-time (8am – 3pm) and night-time (10pm – 3am) low tides. Dartfish (n = 188) and desids (n = 58) living within the LWD were counted and measured (± 0.1 mm). Measurements of dartfish in wood were compared with lengths of dartfish (n = 271) not in wood and caught at high tide. The internal diameters of 1455 randomly chosen teredinid tunnels in LWD from Langira were measured using callipers. Measurements of dartfish and desid diameters (for fish, standard diameter recorded at the operculum base and for spiders, the diameter of the abdomen was recorded) and tunnels were also compared.

Measurements of length and diameters were made from images of the sampled fish and spiders using the digital analysis package ImageTool Version 3.00 (The

University of Texas Health Science Centre at San Antonio). Within this text D. martensi body length refers to the cephalothorax + abdomen.

Numbers of free-swimming dartfish in tidal mangrove creeks and in the water between Rhizophora prop roots fringing open water were counted by snorkel visual census at high tide during day time and at night using a dive torch. During the day, counts were made in belt transects measuring 25 x 4 metres placed at random along creeks in the low- (n = 9), mid- (n = 9) and high intertidal (n = 9) areas of the forest, and at night, transects (n = 4) were established in the same areas and tidal zones respectively. Visual census was not extended beyond the fringing areas as densely packed roots prevented access. 177

Temperatures inside tunnelled wood and in adjacent shallow tidepools were measured (± 0.5 °C, hourly intervals) over a period of one month. The temperature loggers, six Thermochron iButtons® were placed in the mid- and low-intertidal areas of the forest, 1.5 to 1.0 and less than 1.0 m above mean sea level, respectively for ten days each.

To compare temperature fluctuation inside tunnelled wood and adjacent ambient air accessible to sunlight-temperatures, loggers were used at 20 min intervals for 20 consecutive days in the mid-intertidal zone. Within tunnelled wood

(n = 6) and adjacent exposed areas (n = 6), one logger was placed within a fallen log and one logger was placed outside the log exposed to direct sunlight.

To determine the heights above sea level across the tidal gradients within

Langira the high tide marks on mangrove trees were recorded using high visibility string. At low tide, the distance from the substratum to the mark on each tree was measured and then subtracted from high-tide tidal-height to establish height above sea level. Emersion times for the LWD were estimated by relating their tidal height to data in regional tide tables.

6.2.2 Laboratory studies

Wood cooling

To replicate the temperature change of emersed fallen wood in the mangrove forest a 60 x 20 cm Perspex wind tunnel was used to determine if evaporative cooling occurred within LWD samples. The airflow-induced wind speed was 3 m sec-1. Cooling rates were recorded within and at the surface of teredinid tunnelled wood approximately 64 cm3 in volume (n = 45). Using a calibrated thermocouple thermometer (Oakton WD-35427-00 Temp 10 Type J), temperature changes were 178 recorded over 30 minutes within the wind tunnel. The thermometer had two temperature probes; one probe was placed upon the wood surface and the other 3 cm below the surface of the LWD sample. Within the wind tunnel, each wood sample was placed on a mesh-box to ensure that all sides received equal air-flow.

Critical thermal maxima of dartfish

The upper thermal tolerance of dartfish was determined using groups of twenty dartfish some with, and some without access to teredinid tunnels in three aquaria. The dartfish were acclimated for 7-10 days at 27 or 35 °C. The low and high acclimation temperatures chosen represent the approximate average temperature of open water and representative high temperatures in tidepools around Kaledupa, respectively (Taylor et al., 2005; Eme and Bennett, 2009a). The aquaria were placed within a 95-L bath. The water in each aquarium was aerated by a linear air pump.

Approximately 30% of the water was changed daily. Heat in the bath was provided by a 300 W glass AZOO® heater (AZOO, Taikong Corp, Taiwan), and water in the bath was circulated using two AZOO® pumps. Dartfish were fed Sera® Vegetarian

Diet flake food (Montgomery, PA, USA) once per day but not fed 24 h prior to the critical thermal maximum (CTM) trials. Fish acclimated to 35 C were brought to this constant acclimation temperature by increasing the water temperature 1.9 ± 0.9

°C d-1 for 4 d, and then held at this constant temperature for 7d. Fish acclimated to 27

C were held at this constant acclimation temperature for 7d.

Following 7 d of acclimation, each fish was exposed to a CTM trial which consisted of a temperature increase of 0.3 °C min-1 until loss of equilibrium (LOE) was observed. For each CTM trial, randomly selected fish were placed into 250-ml

Nalgene® beakers filled with clean seawater at appropriate constant acclimation 179 temperature. Beakers were suspended within a 20-L, insulated, recirculating water bath and provided with moderate aeration to prevent thermal stratification. A mercury thermometer (certified Fisherbrand® NIST) was used to monitor the temperature in each beaker. Water temperature in the CTM bath was raised using

600-800W of aquarium heater wattage until final LOE was reached. LOE was defined as the inability to maintain dorso-ventral orientation for at least one minute

(Beitinger et al., 2000). For trials with teredinid tunnels, fish were placed into CTM beakers with a piece of calcareous lining from a teredinid tunnel measuring 3-5 cm in length and kept at acclimation temperature at least 6 hours prior to the trial.

Following experiments, fish were released at their site of capture.

Dartfish behaviour observations

Ten dartfish were placed into each of four aquaria. Each aquarium contained nine teredinid tunnels. Observations of dartfish locations were recorded hourly each day during day-time (8am-3pm) and night-time (9pm-2am). After four days, the dartfish in each aquarium were replaced three times with a new group of ten fish, and given 24 hours to acclimate. All fish were released at their site of capture following experiments.

Desid behaviour

Spiders collected from decaying logs were placed in plastic tubs with a small piece of teredinid-attacked wood. To determine if spiders remained within wood day- time and night-time observations were made. Observations for each spider were recorded for 3 - 5 days. The total number of observation days for the desids was 75 days. 180

Fish and spiders were collected under permit LIPI

04/TKPIPA/FRP/SM/IV/2011. Animals in the laboratory studies were treated according to a protocol approved by the University of West Florida, Animal Care and Use Committee (reference #2010-002).

6.3 Data analysis

6.3.1 Field studies

One-way analysis of variance was used to compare the quantity of LWD between the low, mid and high intertidal in the Langira mangrove forest. A two-way

ANOVA was used to compare the temperature within wood and adjacent tidepools at night and day times when immersed and emersed. Regression analyses were used to determine correlations between D. martensi body length versus the ratio of body length to body width and to investigate correlations between wood and pool water/air temperatures. Dartfish lengths were compared between fish collected from LWD at low tide and fish collected at high tide, not in wood using a two-sample t-test. Paired t-tests were used to compare animal abundance and distributions for fish and spiders collected from LWD at low tide and fish collected at high tide between Rhizophora roots and open channels. Chi-Squared tests were used to compare dartfish abundance in wood with different levels of teredinid attacked wood defined by the percentage surface area of tunnels and in areas of the mangrove forests with different emersion times. Chi-Squared tests were also used to determine if desids prefer to remain in

LWD or venture away. 181

Figure 6.1 The tunnelled niche created by teredinids and fauna removed from

vacant teredinid tunnels. a, the prolific effect of the wood-boring teredinid activity

demonstrated by the extensive tunnels, b, a fully exposed dartfish. In the background

openings of tunnels within the fallen log can be seen created by the teredinids and c, an

exposed D. martensi individual removed from within its teredinid tunnel.

6.3.2 Laboratory studies

Analysis of Covariance (ANCOVA) was performed across the two temperature treatments, 27 and 35C, with body mass as the covariate and CTmaximum as the response variable. 1-way ANOVA was used to compare CTmaxima across 182 acclimation treatments and between tunnel treatments. Comparisons of day and night time numbers of dartfish within tunnels were tested using Chi-Square tests. Tukey’s

Studentised Range (TSR) post-hoc test separated values into statistically distinct subsets for all ANOVAs. Count data were square root transformed and all measurements were log transformed. All statistical significance were determined based on α = 0.05. Statistical analyses were performed using MINITAB (MINITAB

Inc, version 13.20).

6.4 Results

6.4.1 Temperature analysis

Mean temperatures measured within emersed wood were lower than in tide pools adjacent to the wood (Figure 6.2 a-b: 2-way ANOVA, with wood/pool, emersion/immersion and night/day as factors, F1,1272 = 61.68, p = < 0.001). This difference was more marked in the mid intertidal zone (Figure 6.2 a). The temperature differential between tunnels in wood and adjacent tide pools was most marked when tide pool temperatures peaked (10am – 4 pm), which occurred at low tide during day-time in the mid intertidal. Temperature lowering within wood was most marked when tide pool temperatures peaked (Figure 6.2 c and d), which occurred at low tide, during day-time (10am – 4pm) in the mid-intertidal when compared with night time hours and immersed wood (Tukey’s pairwise comparisons).

183

a Mid-intertidal b Low-intertidal

30 a 30 pool wood

pool wood ab a

C) b a ° 29 C) 29 a c cd ° b b 28 d 28 d c c 27 27 e d 26 26 25 25

Temperature Temperature ( 24 Temperature Temperature ( 24 23 23 Peak Night Peak Night Peak day Night Peak day Night day day Emersed Immersed Emersed Immersed

c Temperature (°C) of tide pool d Temperature (°C) of tide pool

25 30 35 40 25 30 35 40

1 1

-1 -1 y = -0.42x + 11.32

y = -0.33x + 7.35 C) within wood within C)

° R² = 0.52 C) within wood within C)

-3 R² = 0.16 ° -3

-5 -5 Difference ( Difference -7 ( Difference -7

Figure 6.2, Analysis of variance and regressions of temperatures recorded within

the Langira mangrove. a Temperature differences recorded at night (6 pm – 6 am) and day (10 am – 4 pm) times during high and low tides within wood and tide pools

from the mid-intertidal zone (mean ± SE), and, (b) the low-intertidal zone. Letters

above the bars = Tukey’s pairwise comparisons, (c) the temperature differential within

the mid and (d) low zones between wood and tide pools.

Temperature lowering in wood was most marked when temperatures of sun- exposed wood substratum was emersed during low tide at day-times (10am – 4pm) in the mid-intertidal (Figure 6.3, R-Sq (adj) = 0.69, F1, 1182 = 2646.25, p = < 0.001).

184

Temperature (°C) of ambient air

15 20 25 30 35 40 45 6

-2 C) within wood

° -4

-6

-8 Difference ( -10 y = -0.66x + 16.8 -12 R² = 0.69 -14

Figure 6.3, Variation in temperature differential between air and

wood recorded from Langira. Variation across the range of air

temperatures measured at Langira in the temperature differential between

interior of tunnelled wood and air outside the wood.

When tunnelled wood was subject to airflow, temperatures were slightly reduced (approximately 1 degree C) at the surface, but by over two degrees C in the tunnels (Figure 6.4). The rate of cooling was 6.7 times greater inside teredinid tunnelled wood compared to the surface. The mean rate of cooling on the wood surface was 0.03 ± 0.01 °C min-1 (mean ± SE, n = 45) and the rate of cooling within the centre was 0.2 ± 0.01 °C min-1.

185

2.5

1.5

C) ° 0.5

surface,

- 0 5 10 15 20 25 30 -0.5

-1 (internal

-1.5 Temperature Temperature differential -2 -2.5 -3 Time (min)

Figure 6.4, Wind-tunnel induced cooling of LWD. The temperature

differential between the surface and centre of tunnelled wood samples (n =

45) over 30 minutes (mean ± SE).

6.4.2 Parioglossus interruptus autecology analysis

The CTmaxima of dartfish acclimated to 35 °C, was significantly higher compared to dartfish acclimated to 27 °C (Figure 6.5: ANOVA, with acclimation temperature and tunnel position as factors; F1, 53 = 183.47, P ≤ 0.001). This is an acclimation response of an increase of 1.8 °C, or 0.23 °C of tolerance for every 1 °C increase in acclimation temperature. 186

No tunnels Tunnels

35°C

° 42.5

27°C 41.5

Critical Thermal Maximum CriticalThermal Maximum ( 40.5

Figure 6.5, Critical thermal maxima of dartfish. Fish acclimated to 27 °C and 35 °C had no difference of temperature tolerance with or without out

tunnels. Significant differences of temperature tolerance between acclimated

fish to 27 °C and 35 °C were found (mean ± SE).

Very few of the teredinid tunnels had a smaller diameter than the dartfish, and most dartfish had diameters that overlapped the diameter range of the most numerous tunnel sizes (Figure 6.6 a). The lengths of dartfish collected at high tide, not in LWD tended to be longer than that of dartfish collected from within wood at low tide, 2.6 ± 0.02 cm (mean ± SE, n = 271) compared with 2.1 ± 0.03 cm (n = 188) respectively (Figure 6.6 b: Two-Sample t-test; dartfish at high tide vs. dartfish in wood, t = 14.48, P ≤ 0.001). 187 a

50 Tunnel 45 Dartfish 40

20 Frequency (%) 15

Diameter (mm) b

30 in wood

25 in water

Frequency (%) 10

Length (cm)

Figure 6.6 Measurements of dartfish and teredinid tunnels. a, Size distributions

of dartfish diameters (n = 168) compared against teredinid tunnel diameters (n =

1455), b, Size distributions of dartfish removed from woody detritus (n = 188) at low

tide and dartfish (n = 271) captured during high tide between prop-root systems.

188

Analysis of the LWD samples revealed that wood with a higher percentage surface area of teredinid tunnelling had a greater frequency of dartfish occupying the vacant tunnels (Figure 6.7: Chi-Squared P ≤ 0.001).

45 12 40

35 30 25 36 20 17 15 34 10

% % Occupied by dartfish 5 97 0 < 20 20 - 35 35 - 50 50 - 65 > 65 % of section tunnelled

Figure 6.7, Samples of teredinid-attacked LWD categorised by the level of teredinid

activity imposed upon each woody detrital sample. The percentage of total woody items

(n = 196) occupied by dartfish. Numbers above the bars indicate the number of detrital

samples analysed.

The degree of tidal emersion also affected the degree of LWD occupancy of dartfish (Figure 6.8). With increasing emersion time (h), the amount of wood occupied by dartfish was reduced. The greatest proportion of wood occupied by dartfish was collected in areas of the forest with less than five hours of emersion

(Chi-Squared P ≤ 0.001).

189

30 11

20 44 15

10 83

% % Occupied by dartfish 22 5 36

0 < 5 5 - 7 7 - 9 9 - 11 > 11

Period of emersion (h)

Figure 6.8, The variation of dartfish occupying LWD removed from the forest in

areas exposed to different periods of emersion (hours). With decreasing emersion,

LWD (n = 196) containing dartfish increase. Numbers above the bars indicate the

number of detrital samples analysed.

Estimated abundances of dartfish during high tide within the Langira forest were significantly higher from between the root structures (11433 ± 1690.7 fish ha-1, mean ± SE) than abundances of dartfish from the channels (222 ± 155.5 fish ha-1,

Paired t-test: open channels vs. roots, t = -5.57, P < 0.001). No dartfish were found between roots or in open channels during high tides at night-time (10 pm – 3am).

The numbers of dartfish estimated to be in LWD during day-time low tides in the

Langira forest were 4304 ± 805. 7 fish ha-1.

Observations of fish in aquaria containing pieces of teredinid tunnel linings revealed a marked and significant difference of use of the tunnels between day and night time. The majority of night-time observations of fish location were of fish within the linings (88.2%) while the majority day-time fish location observations 190 were of fish free in the water (84.4%) (percentages relate to 4200 observations each for day and night time conditions, Chi-Squared, P< 0.001).

6.4.3 Desis martensi autecology analysis

Desis martensi were not observed outside wood during lowtide from transects

(n = 8) totalling 1.6 ha within the Langira forest either during day time hours or night time hours. The D. martensi population density within teredinid attacked LWD was

1161 ± 587 (mean ± SE) individuals per ha within the Langira forest.

Laboratory observations of D. martensi distributions revealed that over 95 %

(n = 21) of the total individuals remained within teredinid-attacked LWD (Paired t- test, in tunnels vs. foraging: p = < 0.001). Of the 22 observed individuals, 1 specimen ventured outside from within its tunnel. No significant difference (Chi-Squared, P =

0.699) was found with D. martensi distributions between night and day-time observations.

The largest proportion of desids were found in decaying logs in areas of the mangrove forest immersed in water 15 – 20 hours per day during a spring high tide

(Figure 6.9). The shortest immersion times of LWD containing individuals of D. martensi was 5.6 hours per day and the longest immersion time was estimated to be

21.9 hours per day.

191

90 80 70

60 D. martensi D. 50 40

population 30 20 10

Proportion (%)of Proportion 0 0 - 5 5 - 10 10 - 15 15 - 20 >20 Immersion (h)

Figure 6.9, Proportions of the desid population found in LWD in areas of the

forest subject to a range of immersion times. Desid individuals removed from

teredinid-attacked LWD from Langira.

Size frequency analysis of the desid populations revealed that the most common diameter was between 1-2 mm. The smallest individual found within the detrital samples measured 0.6 mm, and the largest specimen measured 4.5 mm in diameter. When compared with size ranges of teredinid tunnel diameters (Figure

6.10), the size distributions of spiders had a similar fit to the majority of tunnel diameters.

60 Tunnels Spiders

Frequency (%) 20

0 0-1 1-2 2-3 3-4 4-5 5-6 6-7 7-8 8-9 9-10 10-11 11-12 12-13 13-14 > 14 Diameter (mm)

Figure 6.10, The size ranges of desid and teredinid tunnel diameters. Size

distributions of D. martensi individuals (n = 58) removed from teredinid-attacked

woody detritus (n = 82) compared against teredinid tunnel diameters (n = 1455). 192

Replicating teredinid tunnels, desids coaxed in to glass tubes revealed that their legs I, II, and III bunch anteriorly, and the posterior legs “leg IV” are positioned directly behind of the opisthosma (Figure 6.11 a) . Figure 6.11 b, reveals a D. martensi individual with a body length of 12 mm (19.4 mm including legs) turning

180° inside an 8 mm diameter glass tube. D. martensi also have dense scopulate

(bush-like) plumose hairs on each tarsus and metatarsus (Figure 6.11 c), which probably provides an increase in surface area enabling D. martensi to have the ability to walk on water (Figure 6.11 d).

Figure 6.11 Images of D. martensi illustrating morphological characteristics. a, The body

and leg arrangement of D. martensi in a constricted space (dorsal view), and b, a D. martensi

(12 mm body length) turning 180°C with ease within an 8 mm glass tube, c, an S.E.M image

of dense scopulate plumose hairs on the leg I tarsus of D. martensi and d, an individual of D.

martensi walking on water.

193

Regression analysis of body length vs. body/width revealed a significant correlation (F1, 40 = 26.05, p = < 0.001, R-Sq = 0.39). As D. martensi develop from juveniles in to adults they become more elongate (Figure 6.12).

20 y = 1.27x + 2.47

10 R² = 0.39 Ratio(body length/width) 0 0 2 4 6 8 10 12 14

Body length (mm)

Figure 6.12, The change in body proportions of D. martensi

individuals. The regression demonstrates the ratio of body length and

width against the body length.

6.5 Discussion

The effects of teredinid activity may extend after death as the solid wood becomes riddled with tunnels that can be exploited by other organisms. The use of teredinid tunnels by dartfish is an unusual and virtually unknown behaviour among fishes. Juvenile morays, Gymnothorax richardsoni have also been reported to occasionally occupy tunnelled wood, as do a range of marine invertebrates and terrestrial arthropods (Cragg and Hendy, 2010). In tropical Atlantic mangrove ecosystems, the mangrove rivulus, Kryptolebias marmoratus exploits galleries created by terrestrial insects in decaying mangrove logs (Taylor et al., 2008); a 194 similar niche, but distinctly higher in the intertidal than the teredinid-tunnelled wood examined in this study. Indeed, the mangrove rivulus are amphibious even to the point of foraging on land and may also emerge under damp leaf litter, in crab tunnels and beneath plastic debris (Davis et al., 1995; Taylor et al., 2008). Dartfish, however, appear specifically bound to the teredinid-tunnel niche as no other refuge besides decaying wood was available. In addition, dartfish occupy the tunnel niche during low tides, which result in those exposed logs and the adjacent tidepools to experience rapid fluctuations in temperature (minutes to hours). Indeed, mangrove forests frequently have large-diel fluctuations in temperature (Taylor et al., 2005).

However, the Tukang Besi islands have a diverse array of hyperthermic fishes common to the Indo-Pacific region, and mangrove forests in this area contain many fishes with unusual behaviour (Bennett, 2010).

Dartfish show similar thermal tolerance values to many shallow water fishes in the Tukang Besi islands, particularly tidepool-resident goby species (Eme and

Bennett, 2009a), as they exhibit approximately twice the acclimation response of tidepool-resident gobies from around Kaledupa Island. Dartfish also have critical thermal maxima that are higher than those of reef fish species (Eme and Bennett,

2009b; Bennett, 2010). Thus, the dartfish’s thermal biology is that of a robust, shallow-water fish that has an upper thermal tolerance sufficient to deal with temperatures above 40 C. However, dartfish may obtain environmental buffering by avoiding the warmest tidepool temperatures by taking refuge within the cooler vacant teredinid tunnels. Dartfish would probably have a lower resting metabolic rate inside the vacant tunnels. Avoiding the most extreme fluctuations in tidepool temperatures allows for a relatively stable thermal environment. As temperatures become lower 195 within the tunnels the fish may also benefit from evaporative cooling, making desiccation within LWD unlikely for dartfish.

Considerable numbers of vacant teredinid tunnels were observed within LWD from the shore to the seaward edge of the mangrove forest. Dartfish probably use teredinid tunnels to avoid moving into open water during the ebbing tide, therefore the tunnels may provide a novel predation refuge. Within teredinid-attacked wood, dartfish occlude the majority of teredinid tunnels; maintaining a close geometric fit within this cryptic habitat may reduce predation risks as larger predators are not able to access dartfish within the narrow tunnels. Dartfish will bunch within teredinid tunnels during the night time. Similar behaviour described as log packing was reported with K. marmoratus in decaying logs attacked by termites and coleopteran larvae (Taylor et al., 2008). Dartfish bunching in groups within teredinid tunnels may reduce their risks of predation, as higher numbers of fish will plug the teredinid tunnel and may prevent access by predators.

Length measurements of dartfish reveal that the populations removed from decaying logs were smaller than those caught from between the roots at high tide.

Thus, the more vulnerable smaller individuals in a population are more likely to seek refuge in wood, and the larger fish make less use of these cryptic refugia.

Furthermore, adult dartfish range up to 3 cm in length (Allen et al., 2003), suggesting that the smaller dartfish found in the LWD represents an important refuge function for the vulnerable smaller stages of this species in mangrove forests. It is likely that the dartfish found in LWD belong to the same population of fish that is found between the roots and open channels during high tide. Thus, estimates of fish density in LWD at day-time low tides indicate that almost 40% of dartfish from the Langira mangrove seek teredinid-attacked LWD. Observations of dartfish swimming at high- 196 tide during night-time revealed zero encounters during extensive surveys of the mangrove forest. This corroborates our laboratory experiments, which showed that

88% of dartfish were observed within the refuge of teredinid tunnels during night- time.

During high tide, large numbers of dartfish were frequently seen swimming in schools between and around the complex Rhizophora root-systems. Other authors report that small and juvenile fish reduce their risk of predation by hiding among the mangrove roots (Laegdsgaard and Johnson, 2001; Verweij et al., 2006; MacDonald et al., 2009). Although dartfish use teredinid tunnels to avoid being carried in to permanent mangrove creeks or being stranded on mud flats, it appears that they do follow the ebbing tide out to the outermost seaward edge of the forest. Fallen wood exposed at low tide at the seaward edge of the forest would have the least exposure to high temperature fluctuations meaning that the cryptic dartfish have less probability of desiccation due to lower emersion times, which may explain the higher numbers of dartfish found in wood at the seaward fringe.

Non-aquatic, intertidal species such as Desis martensi are challenged by immersion. Many species of spiders live in aquatic ecosystems such as saltmarshes

(Morse, 1997; Petillon et al., 2008; Petillon et al., 2009), and flood plains

(Rothenbucher and Schaefer, 2006) but spider ecology has not been studied in mangrove ecosystems.

Commonly within the LWD, nests created by D. martensi were found containing juveniles along with adults in groups from 2 – 5. No foraging desids were found away from teredinid tunnels at low tide in the forest at day or night times.

However, D. martensi have been reported to forage on the shore at low-tide during night-times in intertidal regions found in Singapore (The Singapore Red Book, 197

2008). Remaining in tunnels will provide relative safety for the spiders, and the wood may provide an important energy subsidy; due to high levels of fauna (see chapter 5;

Cragg and Hendy, 2010), which acts as a prey aggregation device.

When in the tunnels, D. martensi produces a water-proof web that protects them from being immersed. Within wetland habitats spiders either migrate to the shore or move to higher ground in order to avoid immersion. For example, Pirata hygrophilus (Lycosidae) will climb trees to avoid immersion (Adis and Junk, 2002).

However, for Desis remaining in teredinid-attacked wood may also reduce the risk of being predated. Being exposed and visible to predators will have obvious risks, especially in less complex environments (Marczak and Richardson, 2007). Desis also prefers to remain in LWD found in the low intertidal, where there is a greater level of the tunnel niche due to increased teredinid activity in the low-intertidal. Furthermore, due to high availability of the many tunnels in the low intertidal, there may be a high level of prey items for the spiders.

D. martensi may also use the LWD to avoid extreme temperatures, therefore reducing their risk of desiccation. During low tide, day time ambient air temperatures within the mangrove forest peaked at more than 40°C. During the same time the internal temperature of LWD was recorded below 30°C, a difference of more than

12°C. Spiders are more abundant and active in cooler, shaded areas within their environment (Riechert and Tracy, 1975). Extreme temperature may be a limiting factor for the foraging capabilities of spiders. Spiders prefer cooler areas with less prey items than areas of increased temperatures with higher densities of prey

(Riechert and Tracy, 1975). Thus, the cooler temperatures within wood coupled with the high level of prey items provide an avoidance strategy from high temperatures combined with an enhanced exposure to food items. The benefits of avoiding adverse 198 temperatures and having plenty of prey items at your doorstep highlight teredinid- attacked LWD as an ideal habit for D. martensi.

Very few spiders are obligately aquatic. Those that are show a range of interesting adaptations, Argyroneta aquatica, is an air-breathing spider that spends the majority of its life submerged in freshwater habitats (Seymour and Hetz, 2011).

A. aquatica builds a dome-shaped web between aquatic plants. The spider fills the dome-web with air carried from the surface by use of a plastron, hydrofuge hairs that trap air bubbles. The air-filled dome can take up dissolved O2 from the surrounding water acting as a physical gill; pO2 diffuses from the water and through the web in to the spiders nest (Seymour and Hetz, 2011). Terrestrial spiders also have this ability to build underwater webs that have similar properties to an anatomical gill (Rovner,

1987). Dysdera crocata and Ariadna bicolor are both terrestrial spiders that have the ability to survive underwater for up to ten days when in their nests (Rovner, 1987).

Nonetheless, for the intertidal specialist Desis marina, the physical gill effect did not increase their survival time when underwater as these desid species are found living within their web nests in the holdfast cavities of kelp - which have adequate air supplies (McQueen and McLay, 1983 cited in Rovner, 1987). When out of water, D. marina has been reported to venture out of their kelp holdfasts and forage for food

(Rovner, 1987), though this present study suggests this may not be necessary. Unlike

D. marina, D. martensi may likely be obligate to decaying wood as they appear not to venture out of the teredinid tunnels when their woody habitat is emerged.

More than 40% of desids within teredinid-attacked LWD were between 1 - 2 mm. The maximum sizes of desids ranged up to 4 – 5 mm diameter. This suggests that Desis martensi has a life history strategy centred on strong parental care, with lots of energy invested in to its young. D. marina individuals from a rocky shore in 199

New Zealand laid eggs from September – January, and hatching is complete by May

(McLay and Hayward, 1987). The first two instars after hatching remain in the nest until July – August. It was also noted that female D. marina laid 3 to 4 eggs (McLay and Howard, 1987). Thus, the common frequency and occurrence of 2 – 5 individual

D. martensi, usually 1 adult with 2 – 4 juveniles within teredinid tunnels from

Langira suggest that D. martensi have a similar reproductive strategy to D. marina.

D. martensi are physically and morphologically adapted to their intertidal habitat within teredinid tunnels. These data demonstrate that as D. martensi mature from juveniles to adults they become increasingly longer than wide; producing elongate body morphology. Other authors state that the body, cephalothorax and opisthosoma of D. martensi is longer than it is wide (Roth, 1967). The elongate body structure is the ideal shape for living within a vacant teredinid tunnel. Moreover, similar to dartfish, desids also occlude the most frequent size ranges of teredinid tunnels. When the spiders enter teredinid tunnels they adopt a constrained leg arrangement making a narrow stance within the tunnel. When in the tunnels the anterior limbs, legs I, II and III are forward positioned. In addition, the anterior limbs have tarsal trichobothria, sensory hairs (Roth, 1967), that may aid for prey capture when the spider is in its tunnel. D. martensi presumably waits in ambush within its tunnel for small crustaceans such as cirolanids and/or taltrids to come in to contact with the long sensory hairs on the anterior legs which may in turn trigger D. martensi to strike.

The spiders have other diagnostic features and characteristics that make life within teredinid-attacked LWD easier. D. martensi have a malleable body. Spider flexibility is not uncommon (personal communication, Dr Hannah Wood). The flexible turning ability within narrow diameter tunnels is advantageous for D. 200 martensi as it enables them to turn with ease within narrow tunnels when hunting prey items. Suitable tunnel location is also essential. The complex inter-twining teredinid tunnels within fragmented decaying wood when emersed can be water- filled and/or drained, closed off by either sediment or from the calcareous cap the teredinids produce at the distal end of their tunnels or from other predators. Thus, the ability to be hyper-flexible and remaining mobile is a huge advantage in order to avoid potential risks within teredinid tunnels, such as immersion or being predated upon.

D. martensi have large tracheae confined to the abdomen (Roth, 1967).

Desids are aquatic and amphibious, and their tracheae are important adaptation to being able to spend many hours immersed (Selden, 2001). More than eighty percent of D. martensi populations within the Langira mangrove forest are immersed for more than eighteen hours per day during spring high tides. It has been hypothesised that the tracheae are more efficient than other respiratory organs, such as book lungs in preventing water loss from spiders (Levi, 1967); this is vital for D. martensi as the saline environment may increase osmotic stress (Gross, 1957; Pierce, 1982;

Gimenez, 2003). D. martensi also possess heavily scopulate plumose hairs, which are dense bush-like hairs at the tips of their legs. The scopulate plumose hairs may enhance D. martensi’s ability to walk on water; an advantageous ability for an intertidal animal within an aquatic environment.

This study is the first to report a likely obligatory exploitation by a fish and a spider of an intertidal niche provided by teredinid tunnels and unequivocally reveals the fundamental role of the Teredinidae in niche creation. Vacant teredinid tunnels provide a novel ecosystem service within LWD in seaward mangrove forests. This study also demonstrates that seeking environmental buffering by remaining cryptic 201 within teredinid tunnels is an important issue for these species. Lastly, the exploitation of LWD by dartfish and desids emphasises the detrimental impact of forest harvesting. Mangrove forests that are logged will likely impact the abundance and distribution of dartfish and desids in Indonesian mangrove forests, as well as numerous other cryptic, wood dwelling organisms.

202

Chapter 7

GENERAL DISCUSSION

The mangrove forests in the WMP are connected to sea grass beds and coral reefs. Such connectivity of important ecosystems has rarely been studied in

Indonesia (Tomascik et al., 1997). These ecosystems are important for their nursery function (Xavier et al., 2012), and flow of energy between the habitats (Granek et al.,

2009). Indeed, damage to these ecosystems, whether natural or anthropogenic, will disrupt the flow of nutrients across the ecosystems and reduce the biodiversity. The most biologically diverse substratum in the WMP mangrove forests is large woody detritus (LWD), in particular LWD that had been attacked by teredinids. These items of LWD provide a spatial subsidy that permits fish to remain in the mangrove forests at low tide and non-aquatic animals to remain at high tide. Remaining within the mangrove forests limits migrations, therefore remaining within the forests may reduce predation risks (Wang et al., 2009), particularly for vulnerable species and juveniles. LWD within the forests provides alternative refuges for the mangrove fauna if the complexity of the mangroves has declined (Sheaves, 2005).

If the rate of wood harvesting within Indonesian mangrove forests continues then this may have a direct threat to the abundance and distribution of fauna that depend upon the vacant teredinid tunnels within LWD. Fallen wood is an important resource within the mangrove forests, and the tunnelling of teredinids enhances the habitat complexity. Chapters 3, 5 and 6 reveal that teredinids are ecosystem engineers. Chapter five demonstrated that increases of habitat complexity driven by teredinids can positively affect ecosystem-level species-richness. To increase species richness at ecosystem level the engineering species must create conditions not 203 present elsewhere in the landscape and other animals must be able to live only in the engineer-created patches (Wright et al., 2002). This means that for an animal to be considered as an ecosystem engineer, species other than that of the engineering species must be obligate to the habitat created by the ecosystem engineer; such animals are discussed in Chapter Six.

Chapter six demonstrated that vacant teredinid tunnels within LWD are used by many animals within mangrove forests in the WMP, including juveniles and vulnerable species. Thus, LWD enhances biodiversity maintenance and predation refuges in mangrove ecosystems, as a large proportion of individuals exploit this resource. The role of mangrove forests as a predation refuge may be a consequence of the relationship between habitat complexity and predation success (MacDonald et al., 2008). The success of juvenile and vulnerable animals in mangrove habitats are due to the increased niche availability, which reduces predator-prey interactions

(Warfe and Barmuta, 2004; Saha et al., 2009).

Within natural undisturbed ecosystems, teredinids may be considered as ecosystem engineers. Ecosystem engineering accounts for the interaction between an organism and its environment. The engineering species activity results by directly or indirectly transforming the availability of resources to other species (Jones et al.,

1994). Ecosystem engineers modify their environment by creating changes in biotic or abiotic materials, which will ultimately result in the modification, maintenance and creation of new habitats or niches leading to increased habitat heterogeneity

(Jones et al., 1994; Wright et al., 2002; Berkenbusch and Rowden, 2003). Although, in areas of marine environments that necessitate the use of man-made wooden substrata teredinids are considered as pests because they incur much damage to wharfs and pilings (Laverty, 2002). However, the findings reported in parts of 204 chapters three, five and six suggest that teredinids are ecosystem engineers within the seaward edge of mangrove forests of the Tukang Besi as they break down large volumes of LWD, increase ecosystem-level biodiversity of these otherwise species- poor habitats (Duke et al., 2007) and create a predation refuge for vulnerable and juvenile animals. Thus, the abundance and distribution of the fauna dependent upon

LWD may be affected by harvesting of the mangrove wood (Donato et al., 2011).

The biodegradation of LWD by herbivory and decay within the mangrove ecosystems serves as the basis for an elaborate food-web. A consortium of organisms break down the fallen trees in the WMP mangrove forests. Such organisms range from coleopterans, termites, basidiomycetes and teredinids. These organisms process much of the fallen trees that enter the detrital food-chain, recycling the organic carbon back into the ecosystem. LWD will increase habitat complexity (O’Connor,

1991), nutrient enrichment from mineralisation (Gulis et al., 2004), and enhance deposition and stabilisation of fine sediments (Diez et al., 2000; Haschenburger and

Rice, 2004). Within the mid- to low- intertidal areas of mangrove ecosystems, teredinids have a key role in the mechanical-breakdown of LWD (Robertson and

Daniel, 1989a; Kohlmeyer et al., 1995) as they enhance the recycling of nutrients after they have processed the fallen wood (Robertson and Daniel, 1989a).

Recent research has highlighted the fact that mangrove forests are under threat, with an estimated 1 – 2 % each year disappearing globally as a result of wood- harvesting practices (Duke et al., 2007; Donato et al., 2011; Lewis et al., 2011), mainly due to production of prawn farms (Martinez-Alier, 2001). Of the 146,500 km2 of global mangrove forests, 34,931 km2 or 23.8% are located in Indonesia (Spalding et al., 1997). It is estimated that more than 31% of Indonesian mangrove forests have disappeared in the last 30 years (Wilkie and Fortuna, 2003). Thus, more than 1 % of 205

Indonesian mangrove forests are being cleared every year with no sign indicating that it is reducing. Indeed, mangroves cleared for prawn farms incur much damage to the natural environment with declines of invertebrates and fish due to the loss of habitat from uprooted mangrove trees (Martinez-Alier, 2001). The loss of ecosystem services as a result from forest harvesting will be reduced predation avoidance for juvenile fish and invertebrates and reduced deposition of fine sediments as mangroves stabilise sediments. Thus, forest harvesting will decrease species richness and increase water turbidity of adjacent ecosystems. Of the mangrove forests surveyed from this study the greatest impact was found in the Laulua forest (Figure

7.1 a). This forest was heavily cleared with vast areas of bare substrata, and mangrove forest dead-zones with no fauna observed compared to undisturbed areas.

Within the WMP mangrove forests, local demand of mangrove-derived materials for artisanal fishing and construction is great (Cullen, 2010). The harvesting of mangrove wood may lead to a socio-economic issue. Livelihoods are dependent upon fish catches, meaning that the maintenance and up-keep of fish catching devices is necessary (Figure 7.1 b). Thus, the WMP mangroves are under constant harvesting pressure (Figure 7.1 c and d). This harvesting will lead to an ecological negative feed-back loop, as the harvesting of mangroves will impact the biodiversity, therefore human livelihoods will decline (Martinez-Alier, 2001).

206

Figure 7.1 Mangrove forest clearing and wood harvesting within the WMP. a, large exposed

areas of bare sediment from the aftermath of up-rooting trees in the Laulua mangrove forest, b, an

artisanal fishing method, a fish fence constructed from mangrove wood, c, Bruguiera saplings cut

and put in bundles ready to be collected to make fish fences and d, a store of harvested mangrove

wood removed from the WMP mangrove forests.

Global warming may also be a threat to mangrove forests (Mcleod and Salm,

2006). Melting polar-ice leading to rising sea levels, and increasing atmospheric CO2 concentrations leading to decreasing seawater pH (Fabry et al., 2008; Doney et al.,

2009), may impact the forests (Kjerfve and Macintosh, 1997; Mcleod and Salm,

2006). Studies on the impact of mangrove forests on global warming is focused upon the inland migration of mangrove forests driven by rising sea levels (Ellison, 2000), erosion and increased immersion to mangrove plants (He et al., 2007; Yusef and 207

Francisco, 2009). However, the impact of global warming upon the biodegradation processes within mangrove ecosystems has not been considered. Sea-levels for the year 2050 are estimated to increase by 0.4 metres (McInnes et al., 2003), therefore consequences upon the distributions of trees and biodegrading organisms may alter the ecology of those habitats.

In the case of the Galua and Laulua forests, the boundaries between terrestrial and marine biodegradation processes from Laulua and Galua were recorded at 1.82 ±

0.05 and 1.89 ± 0.01 meters above sea level respectively. However, an estimated rise in sea level of 0.4 metres by the year 2050 means that the boundary between terrestrial and marine wood degrading organisms within Laulua and Galua is predicted to be ≤ 2.22 and 2.29 meters above sea level respectively (Figure 7.2).

208

4 (m) 3 Predicted biodegradation boundary 2050

2 Biodegradation boundary 2010 1

Height abovesea level 0 0 100 200 300 400 500 600 700 800 900 b Distance from strandline (m) 4

3 Predicted biodegradation boundary 2050

1 Biodegradation boundary 2010

Height abovesea(m) level 0 0 20 40 60 80 100 120 140 160 180 200 220 Distance from strandline (m)

Figure 7.2 The present day biodegradation boundaries in two mangrove forests of the WMP with predicted boundaries in 2050. a, the shore height

profile of Galua; and, b, Laulua, the black dashed line represents the present day estimated boundary of biodegradation of LWD between terrestrial and

marine organisms driven by emersion time, and the dashed red line represents the estimated biodegradation boundary at the year 2050. 209

As Figure 7.2 illustrates, the threat to Indonesian coastal and intertidal areas from global warming is sea level rise - with the greatest impact imposed upon mangrove ecosystems (Gilman et al., 2008; Yusef and Francisco, 2009). The range of projected sea-level rise is under some conjecture with estimated ranges between

0.1 – 0.4 m above present day levels for 2050 (Michener et al., 1997; McInnes et al.,

2003; Bosello et al., 2007). The majority of research of sea-level rise in mangrove forests is focused upon the upper and lower limits of potentially changing mangrove forest distribution (Ellison and Stoddart, 1991; Mcleod and Salm, 2006; Yusef and

Francisco, 2009). The distribution and community structure of biodegrading organisms discussed in chapter three within the Galua and Laulua forests will be altered. As a consequence, this may alter the breakdown pathways and flux of nutrients from LWD available to other consumers.

In the event of rising sea-levels, the breakdown of lignin by the activity of the basidiomycetes in the terrestrial region of the Galua forest will no longer be possible due to increased salinity caused by the rising seawater. The slow depolymerisation of lignin by white-rot and brown-rot decay (Worrall et al., 1997), will not take place in the landward regions of the Galua forest.

The hypothesis of the landward retreat of mangrove plants as an adaptation to rising sea-levels (Gilman et al., 2007; Gilman et al., 2008; Yusef and Francisco,

2009) does not apply with some of the forests within the WMP, as fossil-coral walls crown the shorelines of the archipelago making mangrove recruitment unlikely.

Increasing sea-levels will also increase immersion times for LWD in the intertidal regions of the WMP. This change would probably drive teredinid populations further into the high intertidal zone by increasing exposure of larval teredinid veligers to areas where there are greater volumes of LWD. The rise in sea 210 levels would probably decrease the abundance and distribution of the coleopteran larvae and termite communities within the intertidal regions due to increased exposure to immersion. The breakdown processes of LWD and how the consequently generated organic carbon may then be re-introduced back into the forest may be drastically altered by shifting wood biodegradation processes driven by rising sea-levels.

Mangrove forests such as those from this study will likely be sensitive to rising sea-levels due to their inability to migrate landward. If the rate of sea-level rise is greater than the rate of change in forest surface elevation, then this would hinder the landward retreat of trees (Gilman, 2004; Gilman et al., 2007). The landward succession of trees would likely depend upon the ability of those trees to colonise new habitats such as increasing shore height elevation (Gilman et al., 2007).

However, many of the forests discussed within the chapters of this research have a landward boundary characterised by fossil-coral walls. Thus, any potential landward migration of those forests would not be possible. The lack of landward migration, coupled with elevated rises in sea levels would likely have adverse consequences for the re-distribution of the biodegradation processes that breakdown the LWD.

Increasing levels of atmospheric CO2 are also linked to global warming, further contributing to problems within marine environments such as coral reefs

(Hoegh-Guldberg et al., 2007; Doney et al., 2009) and calcium carbonate-producing organisms (Parker et al., 2009). Although, increasing atmospheric CO2 may have significant ecological problems for coral reefs and animals such as oysters, such consequences have not been considered for teredinids within mangrove forest environments. The biodegradation boundary of LWD in mangrove forests between terrestrial and marine processes would be changed. The effect of decreased pH upon 211 teredinid populations will have complications for their life histories and ecology due to a reduction of their calcareous hard-parts.

This is evident with other biogenic calcium carbonate secreting organisms such as hard corals as they undergo a reduction of calcification, with loses of 20-60% when exposed to partial pressures that reach levels of 560ppmv which is double pre- industrial CO2 concentrations (Guinotte and Fabry, 2008). Atmospheric partial pressures of CO2 may increase from the present day levels of 380ppmv, to concentrations of 700ppmv by the end of the century (IPCC, 2001 and 2007). The primary concern of an elevated pCO2 in seawater is a net decrease of calcium carbonate saturation, which greatly alters the calcification rates of biogenic calcium- carbonate producing organisms such as molluscs (Royal Society, 2005). These changes create a change in seawater chemistry by decreasing the pH; generating a more acidic marine environment (Fabry et al., 2008; Jokiel et al., 2008; Ries et al.,

2009).

Experiments manipulating decreased pH levels in seawater demonstrated that calcifying species experienced reduced calcification rates. Oyster larvae exposed to seawater with decreased pH concentrations displayed malformed shells (Kurihara et al., 2007; Parker et al., 2009). The bivalves Mytilus edulis and Crassostrea gigas both exhibit declines of calcification with increasing pCO2 levels of 25 and 10% respectively (Gazeau et al., 2007). The emphases of these studies indicate that the early life history strategies of bivalve molluscs are greatly compromised by acidic oceans (Kurihara et al., 2004). The increased de-calcification rates of oysters and mussels can be reflected to other bivalve molluscs such as the Teredinidae. If the shells of the teredinid veliger are greatly compromised, successful recruitment within

LWD would not be possible – making survival unlikely as the postlarval shell is 212 needed for burrowing (Turner, 1976). Upon development, proteinaceous and calcareous material is added to the juvenile teredinid shell, which has teeth-like projections used for rasping. Teredinids also produce calcareous pallets - exclusive to the family (Turner, 1966). Teredinid boring rates ranges from 0.03 to 3.85mm day-1, which accelerates with increasing length of the body (Manyak, 1982).

If decreasing pH inhibits teredinid settlement, this may consequently transfer the competitive benefit to non-calcifying animals. If global warming increases sea- levels and ocean acidification as predicted; there would be a reduction with the volume of LWD processed in mangrove ecosystems. The reduced wood- biodegradation in the mid- to low- intertidal areas of the forests would have further detrimental consequences, as LWD serves as an integral and essential source of nutrients (Gee and Somerfield, 1997), and is the base of an elaborate food web

(Benner and Hodson, 1985).

As a consequence of global warming, increasing sea level rise and partial pressures of CO2 concentrations has revealed the need to investigate the effect of global warming upon the distribution of the terrestrial and marine biodegradation processes. Global warming may induce shifts of the biodegradation processes within mangrove ecosystems, with unknown consequences. The forests from the Tukang

Besi Archipelago have a wide tidal range extending from mean sea level to the high intertidal. Conversely, mangrove forests from the Caribbean have a comparatively limited tidal regime that range in height by 20 to 40 cm (Ilka C. Feller, personal communication). The investigation of the biodegradation boundary of LWD in forests from the Caribbean is necessary, due to the limited tidal flow. A different suite of biodegradation processes may be found within Caribbean mangrove forests due to the reduced emersion time. Differences of the biodegradation processes may 213 provide insights into the effect of wood-degradation within permanently and tidally inundated forests. Increasing ocean acidification may have ecological implications for the marine biodegrading detritivores in mangrove ecosystems, which may ultimately affect the flow of nutrients to other ecosystems.

Final conclusions

Despite being considered as economic pests, the Teredinidae are the dominant biodegraders of LWD within the seaward edges of mangrove forests. The fragmentation of LWD in the mid to low intertidal areas of the forests are enhanced by teredinids. Much of the LWD within the mangrove forests is processed and recycled back into the ecosystem due to the mechanical effect and digestion by teredinids. The tunnelling activity of teredinids significantly modifies the environment, increases the mangrove faunal diversity at ecosystem level and creates niches for animals that exhibit obligate behaviour to that substrate. Furthermore, teredinid-attacked LWD has the greatest faunal diversity compared to mangrove forest sediments and Rhizophora prop-roots within the mangrove forests of the

WMP. These data demonstrate that the teredinid populations within the forests of the

Tukang Besi Archipelago can be defined as ecosystem engineers.

Teredinids and their tunnelling of LWD within Indonesian forests will be affected by increasing wood harvesting pressures and global warming. If wood harvesting continues there will be a reduction of mangrove faunal diversity, due to the reduced woody habitat and lack of teredinid tunnel niches. If sea levels rise and ocean pH levels decrease as predicted there will be a redistribution of the wood- biodegrading organisms ultimately leading to potentially changing plant zonation, with fewer nutrients sequestered within the mangrove forest sediments combined 214 with a build-up of LWD and reduced flow of recycled mangrove-derived organic carbon to adjacent ecosystems.

Priorities for further research in this area are needed to address issues concerning global warming. Experiments to determine the de-calcification rates of larval and adult teredinid populations would provide insights into reduced boring rates and estimates of increasing volumes of LWD in mangrove forests. Experiments could be used to replicate global warming-induced ocean acidification, and to investigate the effect of decreased pH upon the Teredinidae – as the recycling of organic carbon may be compromised in these coastal habitats due to global warming.

Successful mangrove forest management depends on their current status and the prediction of detrimental ecological issues after the ecosystem structure and function has been altered. Thus, projects that maintain the mangrove forest habitat complexity after the trees have been harvested by way of implanting fallen logs may enhance the faunal diversity of impacted mangrove areas in the WMP and maintain the flow of organic carbon from LWD.

215

Reference list

Adams, E. S., Atkinson, L., and Bulmer, M. S. 2007 Relatedness, recognition errors, and

colony fusion in the termite Nasutitermes corniger. Behavioural Ecology and

Sociobiology. 61, 1195 – 1201

Adis, J., and Junk, W. J. 2002 Terrestrial invertebrates inhabiting lowland river floodplains

of Central Amazonia and Central Europe: a review. Freshwater Biology. 47, 711 –

731

Agrawal, A. A. 1999 Induced responces to herbivory in wild radish: effects on several

herbivore and plant fitness. Ecology. 80, 1713 – 1723

Agrawal, A. A., Strauss, S. Y., and Stout, M. J. 1999 Costs of reduced responses and

tolerance to herbivory in male and female fitness components of wild radish.

Evolution. 53, 1093 - 1104

Allen, J. A., Ewel, K.C., Keeland, B. D., Tara, T. and Smith III, T. J. 2000 Downed wood in

Micronesian mangrove forests. Wetlands. 20, 169 – 176

Allen, G., Roger, S., Human, H., Deloach, N. 2003 Reef Fish Identification - Tropical

Pacific. Florida, New World Publications, Inc. USA

Allen, G. R. 2008 Conservation hotspots of biodiversity and endemism for Indo-Pacific

coral reef fishes. Aquatic Conservation, Marine and Freshwater Ecosystems. 18, 541

– 556

Alongi, D. M. 1987 The influence of mangrove-derived tannins on intertidal meiobenthos in

tropical estuaries. Oecologia. 71, 537 – 540

Alongi, M. D. 2002 Present state and future of the world’s mangrove forests. Environmental

Conservation. 29, 331 – 349

216

Alongi, D., 2007 The mangrove forests of Papua. pp. 824-857. In: Marshall, A. J., and

Beehler, B. M. (eds) In The Ecology of Papua. The Ecology of Indonesia Series,

Volume 5, Periplus Editions Ltd

Alongi, D. M. 2009 Paradigm shifts in mangrove biology. Chapter 22, in Perillo, G. M. E.,

Wolanski, E., Cahoon, D. R., and Brinson, M. M. (eds.) Coastal Wetlands: An

Integrated Ecosystem Approach. Elsevier Press.

Ananda, K., and Sridhar, K. R. 2004 Diversity of filamentous fungi on decomposing leaf

and woody litter of mangrove forests in the southeast coast of India. Current Science.

87, 1431 - 1437

Arnett, R. H. 1951 A revision of the Neartic Oedemeridae. The American Midland

Naturalist. 45, 257 - 391

Arsuffi, T. L., Suberkropp, K. 1989 Selective feeding by shredders on leaf-colonising stream

fungi: comparison of macroinvertebrate taxa. Oecologia. 79, 30 – 37

Ashton, E. C., and Macintosh, D. J. 2002 Preliminary assessment of the plant diversity and

community ecology of the Sematan mangrove forest, Sarawak, Malaysia. Forest

Ecology and Management. 166, 111 - 129

Babu, R., Varadharajan, D., Soundarapandian, P., and Balasubramanian, R. 2010 Fungi

diversity in different coastal marine ecosystem along South East coast of India.

International Journal of Microbiological Research. 1, 175 - 178

Barkati, S., and Tirmizi, N. M. 1991 The environmental impact of wood-borers in mangrove

swamps. Symposia of the Zoological Society of London. 63, 251 – 263

Barnes, R. S. K. 2003 Interactions between benthic molluscs in a Sulawesi mangal,

Indonesia: the cerithiid mud-creeper Cerithium coralium and potamidid mud-

whelks, Terebralia spp. Journal of the Marine Biological Association of the United

Kingdom. 83, 483 – 487 217

Beitinger, T.L., Bennett, W.A., & McCauley, R.W 2000 Temperature tolerances of North

American freshwater fishes exposed to dynamic changes in temperature.

Environmental Biology of Fishes. 58, 237 – 275

Benke, A. C., Henry III, R. L., Gillespie, D. M., and Hunter, R. J. 1985 Importance of snag

habitat for animal production in Southeastern streams. Fisheries. 10, 8 – 13

Benner, R., and Hodson, R. E. 1985 Microbial degradation of the leachable and

lignocellulosic components of leaves and wood from Rhizophora mangle in a

tropical mangrove swamp. Marine Ecology Progress Series. 23, 221 - 230

Bennett, W.A. 2010 Extreme physiology of intertidal fishes of the Wakatobi. In Marine

Research and Conservation in the Coral Triangle: The Wakatobi National Park (eds

J. Clifton, R. K. F. Unsworth & D. J. Smith), pp. 111–128 Nova Science Publishers,

New York

Berkenbusch, K., and Rowden, A. A. 2003 Ecosystem engineering – moving away from

‘just-so’ stories. New Zealand Journal of Ecology. 27, 67 - 73

Bilby, E. R., and Likens, G. E. 1980 Importance of orbanic debris dams in the structure and

function of stream ecosystems. Ecology. 6, 1107 – 1113

Blasco, F., Saenger, P., and Janolet, E. 1996 Mangroves as indicators of coastal change.

Catena. 27, 167 - 178

Bloch, R. 1952 Wound healing in higher plants II. Botanical Review. 18, 655 – 679

Borges, L. M. S., Cragg, S. M., Bergot, J., Williams, J. R., Shayler, B., and Sawyer, G. S.

2008 Laboratory screening of tropical hardwoods for natural resistance to the marine

borer Limnoria quadripunctata: The role of leachable and non-leachable factors.

Holzforschung. 62, 99 - 111

Borojevic, K., Steiner Jr. W. E., Gerisch, R., Zazzaro, C., and Ward, C. 2010 Pests in an

ancient Egyptian harbour. Journal of Archaeological Science. 37, 2449 – 2458 218

Bosello, F., Roson, R., and Tol, R. S. J. 2007 Economy-wide estimates of the implications of

climate change: sea level rise. Environmental and Resource Economics. 37, 549 -

571

Bosire, J. O., Dahdouh-Guebas, F., Kairo, J. G., Wartel, S., Kazungu, J., and Koedan, N.

2006 Success rates of recruited tree species and their contribution to the structural

development of reforested mangrove stands. Marine Ecology Progress Series. 325,

85 - 91

Boto, K. G., and Bunt. J. S. 1982 Carbon export from mangroves. Journal of Marine and

Freshwater Research. 105 - 110

Bouillon, S., Connolly, R. M., and Lee, S. Y. 2008 Organic matter exchange and cycling in

mangrove ecosystems: Recent insights from stable isotope studies. Journal of Sea

Research. 59, 44 - 58

Bouma, T. J., Olenin, S., Reise, K., and Ysebaert, T. 2009 Ecosystem engineering and

biodiversity in coastal sediments: posing hypotheses. Helgoland Marine Research.

63, 95 – 106

Brearley, A., Chalermwat, K., and Kakhai, N. 2003 Pholadidae and Teredinidae (Mollusca:

Bivalvia) collected from mangrove habitats on the Burrup Peninsula, Western

Australia. In, Wells, F. E., Walker, D. I., and Jones, D. S. (Eds) The marine flora

and fauna of Dampier, Western Australia. Western Australia Museum, Perth,

Western Australia. 345 - 356pp

Brooks, R. A., and Bell, S. S. 2002 Mangrove response to attack by a root boring isopod:

root repair versus architectural modification. Marine Ecology Progress Series. 231,

85 – 90

219

Brooks, A. P., Gehrke, P. C., Jansen, J. D., and Abbe, T. B. 2004 Experimental

reintroduction of wood debris on the Williams River, NSW: geomorphic and

ecological responses. River Research and Applications. 20, 513 – 536

Brush, G. S. 2009 Historical land use, nitrogen, and coastal eutrophication: a paleoecological

perspective. Estuaries and Coasts. 32, 18 - 28

Bucher, V. V. C., Pointing, S. B., Hyde, K. D., and Reddy, C. A. 2004 Production of wood

decay enzymes, loss of mass, and lignin solubilisation in wood by diverse tropical

freshwater fungi. Microbial Ecology. 48, 331 - 337

Bunt, J. S., Boto, K. G., and Boto, G. 1979 A survey method of estimating potential levels of

mangrove forest primary production. Marine Biology. 52, 123 - 128

Castillo, G., and Demoulin, V. 1997 NaCl salinity and temperature effects on growth of

three wood-rotting basidiomycetes from a Papua New Guinea coastal forest.

Mycology Research. 101, 341 – 344

Cerri, C. E. P., Easter, M., Paustian, K., Killian, K., Coleman, K., Bernoux, M., Falloon, P.,

Powlson, D. S., Batjes, N. H., Milne, E., Cerri, C. C. 2007 Predicted soil organic

carbon stocks and changes in the Brazilian Amazon between 2000 and 2030.

Agriculture, Ecosystems and Environment. 122, 58 – 72

Chan, Y. H. 1982 Storage and release of organic carbon in Peninsular Malaysia.

International Journal of Environmental Studies. 18, 211 - 222

Clausen, R., and York, R. 2008 Global biodiversity decline of marine and freshwater fish: a

cross-national analysis of economic, demographic, and ecological influences. Social

Science Research. 37, 1310 – 1320

Clifton, J., and Unsworth, R. K. F. 2010 Introduction In: Clifton, J., Unsworth, R., and

Smith, D. (Eds.). Marine Research and Conservation in the Coral Triangle. Nova

Science Publishers, Hauppauge, New York 220

Clough, B. F. 1984 Mangroves. In: Control of Crop Productivity. Academic press, Sydney.

pp 253 -268

Collier, K. J., and Halliday, J. N. 2000 Macroinvertebrate-wood associations during decay of

plantation pine in New Zealand pumice-bed streams: stable habitat or trophic

subsidy? Journal of the North American Benthological Society. 19, 94 – 111

Connell, J. H. 1978 Diversity in tropical rain forests and coral reefs. Science. 199, 1302 –

1310

Cookson, L. J., Cragg, S. M., and Hendy, I. W. 2012 Wood-boring limnoriids (Crustacea,

Isopoda) including a new species from mangrove forests of the Tukang Besi

Archipelago, Indonesia. Zootaxa. 3248, 25 - 34

Coomans, E. H. 1969 Biological aspects of mangrove molluscs in the West Indies.

Proceedings of the Third European Malacological Congress. 9, 79 - 84

Correa, M. D. O. D. A., and Uieda, V. S. 2008 Composition of the aquatic invertebrate fauna

associated to the mangrove vegetation of a coastal river, analysed through a

manipulative experiment. Pan-American Journal of Aquatic Sciences. 3, 23 - 31

Crabbe, M. J. C., and Smith, D. J. 2005 Sediment impacts on growth rates of Acropora and

Porites corals from fringing reefs of Sulawesi, Indonesia. Coral Reefs. 24, 437 - 441

Cragg, S. M. In, The Ecology of Papua; Behler, B. M.; Marshall, A. J.; Eds, Periplus:

Singapore, 2007. pp539 - 563

Cragg, S. M., Jumel, M. C., Al-Horani, F. A. & Hendy, I.W. 2009 The life history

characteristics of the wood-boring bivalve Teredo bartschi are suited to the elevated

salinity, oligotrophic circulation in the Gulf of Aqaba, Red Sea. Journal of

Experiemental Marine Biology and Ecology. 375, 99 – 105

221

Cragg, S.M., Hendy, I.W. 2010 Mangrove Forests of the Wakatobi National Park. In:

Clifton, J., Unsworth R.K.F., and Smith, D.J. (Eds.), Marine Research and

Conservation in the Coral Triangle. Nova Science Publishers, Hauppauge, New

York

Cullen, L. C., Pretty, J., Smith, D., and Pilgrim, S. E. 2007 Links between local ecological

knowledge and wealth in indigenous communities of Indonesia: implications for

conservation of marine resources. Journal of Social Science. 2, 289 – 299

Cullen, L. C. 2010 Marine resource dependence and natural resource use patterns in the

Wakatobi In: Clifton, J., Unsworth, R. K. F., and Smith, D. J. (Eds.), Marine

Research and Conservation in the Coral Triangle. Nova Science Publishers,

Hauppauge, New York

Dahlgren, C. P., Kellison, T. G., Adams, A. J., Gillanders, B. M., Kendall, M. S., Layman,

C. A., Ley, J. A., Nagelkerken, I., and Serafy, J. E. 2006 Marine nurseries and

effective juvenile habitats: concepts and applications. Marine Ecology Progress

Series. 312, 291 – 295

Davis, W.P., Taylor, D.S. and Turner, B.J 1995 Does the autecology of the mangrove rivulus

fish (Rivulus marmoratus) reflect a paradigm for mangrove ecosystem sensitivity?

Bulletin of Marine Science. 57, 208 – 214

Davis, S. E., Corronado-Molina, C., Childers, D. L. and Day, J. W. 2003 Temporally

dependant C, N, and P dynamics associated with the decay of Rhizophora mangle L.

Leaf litter in oligotrophic mangrove wetlands of the southern Everglades. Aquatic

Botany. 75, 199 - 215

Dechart, G., Veldkamp, E., and Brumme, R. 2005 Are partial nutrient balances suitable to

evaluate nutrient sustainability of land use systems? Results from a case study in

Central Sulawesi, Indonesia. Nutrient Cycling in Agroforestry. 72, 201 - 212 222

Diez, J. R., Larranaga, S., Elosegi, A., and Pozo, J. 2000 Effect of removal of wood on

streambed stability and retention of organic matter. Journal of the North American

Benthological Society. 19, 621 - 632

Distel, L. D. 2003 The biology of marine wood boring bivalves and their bacterial

endosymbionts. American Chemical Society. 253 – 271

Donato, D. C., Kauffman, J. B., Murdiyarso, D., Kurnianto, S., Stidham, M., and Kanninen,

M. 2011 Mangroves among the most carbon-rich forests in the tropics. Nature

Geoscience. 4, 293 – 297

Doney, S. C., Fabry, V. J., Feely, R. A., and Kleypas, J. A. 2009 Ocean acidification: The

other CO2 problem. Annual Review of Marine Science. 1, 169 - 192

Donner, S. D., Knutson, T. R., and Oppenheimer, M. 2007 Model-based assessment of the

role of human-induced climate change in the 2005 Caribbean coral bleaching event.

Proceedings of the National Academy of Sciences. 1 – 6

Doorn, W. G. V., and Cruz, P. 2000 Evidence for a wounding-induced xylem occlusion in

stems of cut chrysanthemum flowers. Postharvest Biology and Technology. 19, 73 –

Duke, C. N., Bunt. S. J., and Williams. T. W. 1981 Mangrove litter fall in north-eastern

Australia. I. Annual totals by component in selected species. Australian Journal of

Botany. 29, 547 - 53

Duke, N. C. (1992), Mangrove floristics and biogeography, in Tropical Mangrove

Ecosystems, Coastal Estuarine Series., vol. 41, edited by A. I. Robertson and D. M.

Alongi, pp. 63–100, AGU, Washington, DC

Duke, N. C., Ball, M. C., and Ellison, J. C. 1998 Factors influencing biodiversity and

distribution gradients in mangroves. Global Ecology and Biogeography Letters. 7, 27

– 47 223

Duke, N. C. 2001 Gap creation and regenerative processes driving diversity and structure of

mangrove ecosystems. Wetlands Ecology and Management. 9, 257 – 269

Duke, N. C., Meynecke, J-O., Dittmann, S., Ellison, A. M., Anger, K., Berger, U., Cannicci,

S., Dielle, K., Ewel, K. C., Field, C. D., Koedam, N., Lee, S. Y., Marchand, C.,

Nordhaus, I., Dahdouh-Guebas, F. 2007 A World without mangroves? Science. 317,

41 – 42

Eggert, C., Temp, V., and Eriksson, K. L. 1996 The lignolytic system of the white rot fungus

Pycnoporus cinnaburinus: Purification and characterisation of the laccase. Applied

and Environmental Microbiology. 62, 1151 – 1158

Ejechi, B. O. 2002 Microbial deterioration of partially submerged service timbers in a

tropical inter-tidal zone. International Biodeterioration and Biodegradation. 51, 115

– 118

Ellison, A. M., and Farnsworth, E. J. 1990 The ecology of Belizean mangrove-root fouling

communities. I. Epibenthic fauna are barriers to isopod attack of red mangrove roots.

Journal of Experimental Marine Biology and Ecology. 142, 91 - 104

Ellison, J. C., and Stoddart, D. R. 1991 Mangrove ecosystem collapse during predicted sea-

level rise: Holocene analogues and implications. Journal of Coastal Research. 7,

151 - 165

Ellison, A. M., and Farnsworth, E. J. 1992 The ecology of Belizean mangrove-root fouling

communities: patterns of epibont distribution and abundance, and effects on root

growth. Hydrobiologia. 247, 87 - 98

Ellison, A. M., and Farnsworth, E. J. 1996 Spatial and temporal variability in growth of

Rhizophora mangle saplings on coral cays: links with variation in insolation,

herbivory, and local sedimentation rate. Journal of Ecology. 84, 717 – 731 224

Ellison, J. C. 2000 How South Pacific mangroves may respond to predicted climate change

and sea-level rise. In, Gillespie, A., and Burns, W. C. G. (eds.) Climate Change in

the South Pacific: Impacts and responses in Australia, New Zealand and small

Island states, Dordrecht, Netherlands, Kluwer Academic Publishers, The

Netherlands. 289 - 301

Eme, J., and Bennett, W. A. 2009a Acute temperature quotient responses of fishes reflect

their divergent thermal habitats in the Banda Sea, Sulawesi, Indonesia. Australian

Journal of Zoology. 57, 357 - 362

Eme, J. and Bennett, W. A. 2009b Critical thermal tolerance polygons of tropical marine fish

from Sulawesi, Indonesia. Journal of Thermal Biology. 34, 220 – 225

Engelhart, S. E., Horton, B. P., Roberts, D. H., Bryant, C. L., and Corbett, D. R. 2007

Mangrove pollen of Indonesia and its suitability as a sea-level indicator. Marine

geology. 242, 65 - 81

Evans, T. A., Inta, R., and Lenz, M. 2007 Foraging vibration signals attract foragers and

identify food size in the drywood termite, Cryptotermes secundus. Insectes Sociaux.

54, 374 – 382

Everett, R. A., and Ruiz, G. M. 1993 Coarse woody debris as a refuge from predation in

aquatic communities. An experimental test. Oecologia. 93, 475 – 486

Fabry, V. J, Seibel, B. A., Feely, R. A., and Orr, J. C. 2008 Impacts of ocean acidification on

marine fauna and ecosystem processes. ICES Journal of Marine Science. 65, 414 -

432

Feller, I. C., and Mathis, W. N. 1997 Primary herbivory by wood-boring insects along an

architectural gradient of Rhizophora mangle. Biotropica. 29, 440 - 451

Feller, I. C., and Mckee, K. L. 1999 Small gap creation in Belizean mangrove forests by a

wood boring insect. Biotropica. 31, 607 - 617 225

Feller, I. C. 2002 The role of herbivory by wood-boring insects in mangrove ecosystems in

Belize. Oikos. 97, 167 - 176

Filho, C. S., Tagliaro, C. H. & Beasley, C. R. 2008 Seasonal abundance of the shipworm

Neoteredo reynei (Bivalvia, Teredinidae) in mangrove driftwood from a northern

Brazilian beach. Iheringia Series of Zoology. 98, 17 – 23

Foley, J. A., Asner, G. P., Costa, M. H., Coe, M. T., DeFries, R., Gibbs, H. K., Howard, E.

A., Olson, S., Patz, J., Ramankutty, N., and Snyder, P. 2007 Amazonia revealed:

forest degradation and loss of ecosystem goods and services in the Amazon Basin.

Frontiers in Ecology and the Environment. 5, 25 – 32

Fryar, S. C., Davies, J., Booth, W., Hodgkiss, I. J., and Hyde, K. D. 2004 Succession of

fungi on dead and live wood in brackish water in Brunei. Mycologia. 96, 219 - 225

Gaume, L., and Mckey, D. 1997 Benefits conferred by “timid” ants: active anti-herbivore

protection of the rainforest tree Leonardoxa africana by the minute ant

Petalomyrmex phylax. Oecologia. 112, 209 - 216

Gazeau, F., Quiblier, C., Jansen, J. M., Gattuso, J. P., Middelburg, J. J., and Heip, C. H. R.

2007 Impact of elevated CO2 on shellfish calcification. Geophysical Research

Letters. 34, 1 - 5

Gee, J. M., and Somerfield, J. P. 1997 Do mangrove diversity and litter decay promote

meiofaunal diversity? Journal of Experimental Marine Biology and Ecology. 218,

13 - 33

Giesen, W., Wulfraat, S., Zieven, M., and Scholten, L. 2006 Mangrove guidebook for

Southeast Asia. RAP Publication 2006/07. Bangkok, Thailand. FAO regional office

for Asia and the Pacific and Wetlands international. ISBN 974-7946-85-8

Gilbert, G. S., and Sousa, W. D. 2002 Host specialisation among wood-decay polypore fungi

in a Caribbean mangrove forest. Biotropica. 34, 396 - 404 226

Gill, A. M., and Tomlinson, P. B. 1971 Studies on the growth of red mangrove (Rhizophora

mangle L.) 2. Growth and differentiation of aerial roots. Biotropica. 3, 63 - 77

Gill, A. M., and Tomlinson, P. B. 1977 Studies on the growth of red mangrove (Rhizophora

mangle L.) 4. The adult root system. Biotropica. 9, 145 – 155

Gilman, E. 2004 Assessing and managing coastal ecosystem response to change in sea-level

and climate. Prepared for the international research foundation for development

forum on small island developing states: challenges, prospects and international

cooperation for sustainable development. Contribution to the Barbados +10 United

Nations International meeting on sustainable development of small island

developing states, Port Louis, Mauritius, 10-14 January 2005. Honolulu, Blue Ocean

Institute

Gilman, E., Ellison, J., and Coleman, R. 2007 Assessment of mangrove response to

projected relative sea-level rise and recent historical reconstruction of shoreline

position. Environmental Monitoring Assessment. 124, 105 - 130

Gilman, E., Ellison, J., Sauni Jr, I., and Tuaumu, S. 2007 Trends in surface elevations of

American Samoa mangroves. Wetlands Ecology Management. 15, 391 - 404

Gilman, E. L., Ellison, J., Duke, N. C., and Field, C. 2008 Threats to mangroves from

climate change and adaptation options. Aquatic Botany. 89, 237 - 250

Gimenez, L. 2003 Potential effects of physiological plastic responses to salinity on

population networks of the estuarine crab Chasmagnathus granulate. Helgoland

Marine Research. 56, 265 – 273

Godwin, S. L. 2003 Hull fouling of maritime vessels as a pathway for marine species

invasions to the Hawaiian Islands. Biofouling. 19, 123 – 131

Granek, E., and Ruttenburg, B. J. 2008 Changes in biotic and abiotic processes following

mangrove clearing. Estuarine, Coastal and Shelf Science. 80, 555 - 562 227

Granek, E. F., Compton, J. E., and Phillips, D. L. 2009 Mangrove-exported nutrient

incorporation by sessile coral reef invertebrates. Ecosystems. 12, 462 - 472

Gratwicke, B., and Speight, M. R. 2005 The relationship between fish species richness,

abundance and habitat complexity in a range of shallow tropical marine habitats.

Journal of Fish Biology. 66, 650 - 667

Greenwood, J. L., Rosemond, A. D., Wallace, J. B., Cross, W. F., and Weyers, H. S. 2007

Nutrients stimulate leaf breakdown rates and detritivore biomass: bottom-up effects

via heterotrophic pathways. Oecologia. 151, 637 - 649

Gross, W. J. 1957 An analysis of response to osmotic stress in selected decapod crustacea.

Biological Bulletin. 1, 43 – 62

Grossman, G. D., Ratajczak Jr, R. E., Crawford, M., and Freeman, M. C. 1998 Assemblage

organization in stream fishes: effects of environmental variation and interspecific

interactions. Ecological Monographs. 68, 395 – 420

Guinotte, J. M., and Fabry, V. J. 2008 Ocean acidification and its potential effects on marine

ecosystems. Annual New York Academy of Science. 1134, 320 - 342

Gulis, V. V., Rosemond, A. D., Suberkropp, K., Weyers, H. S., and Benstead, J. P. 2004

Effects of nutrient enrichment on the decomposition of wood and associated

microbial activity in streams. Freshwater Biology. 49, 1437 – 1447

Hall, R., and Wilson, M. E. J. 2000 Neogene sutures in eastern Indonesia. Journal of Asian

Earth Sciences. 18, 781 – 808

Han, L., Huang, X., Sattler, I., Dahse, H-M, Fu, H., Lin, W., and Grabley, S. 2004 New

diterpenoids from the marine mangrove Bruguiera gymnorrhiza. Journal of Natural

Products. 67, 1620 - 1623 228

Hanley, M. E., Lamont, B. B., Fairbanks, M. M., and Rafferty, C. M. 2007 Plant structural

traits and their role in anti-herbivore defence. Perspectives in Plant Ecology,

Evolution and Systematics. 8, 157 – 178

Haschenburger, J. K., and Rice, S. P. 2004 Changes in woody debris and bed material

texture in a gravel-bed channel. Geomorphology. 60, 241 – 267

Hata, H., Kudo, S., Yamano, H., Kurano, N., Kayanne, H. 2002 Organic carbon flux in

Shiraho coral reef (Ishigaki Island, Japan). Marine Ecology Progress Series. 232,

129 - 140

Hauff, R. D., Ewel, K. C., and Jack, J. 2006 Tracking human disturbance in mangroves:

estimating harvest rates on a Micronesian Island. Wetlands Ecology and

Management. 14, 95 - 105

Haukioja, E., and Koricheva, J. 2000 Tolerance to herbivory in woody vs. herbaceous plants.

Evolutionary Ecology. 14, 551 – 562

He, B., Lai, T., Fan, H., Wang, W., and Zhang, H. 2007 Comparison of flooding-tolerance in

four mangrove species in a diurnal tidal zone in the Beibu Gulf. Estuarine, Coastal

and Shelf Science. 74, 254 - 262

Hilderbrand, R. H., Lemly, A. D., Dolloff, A., and Harpster, K. L. 1997 Effects of large

woody debris placement on stream channels and benthic macroinvertebrates.

Canadian Journal of Fish and Aquatic Science. 54, 931 – 939

Hoegh-Guldberg, O., Mumby, P. T., Hooten, A. J., Steneck, R. S., Greenfield, P., Gomez,

E., Harvell, C. D., Sale, P. F., Edwards, A. J., Caldeira, K., Knowlton, N., Eakin, C.

M., Iglesias-Prieto, R., Muthiga, N., Bradbury, R. H., Dubi, A., and Hatziolos, M. E.

2007 Coral reefs under rapid climate change and ocean acidification. Science. 318,

1737 - 1742 229

Hol, G. W. H., Macel, M., Van Veen, J. A., and Meijden, E. V. D. 2004 Root damage and

aboveground herbivory change concentration and composition of pyrrolizidine

alkaloids of Senecio jacobaea. Basic and Applied Ecology. 5, 253 – 260

Hughes, T. P. 1994 Catastrophes, phase-shifts, and large-scale degradation of a Caribbean

coral reef. Science. 265, 1547 - 1551

IPCC. Climate change 2001: the scientific basis contribution of working group 1 to the Third

assessment report of the intergovernmental panel on climate change. Cambridge

University Press. Cambridge

IPCC. Climate change 2007: the physical science basis. Contribution of working group 1 to

the Forth assessment report of the intergovernmental panel on climate change.

Cambridge University Press. Cambridge

Jimenez, A. J., and Sauter, K. 1991 Structure and dynamics of mangrove forests along a

flooding gradient. Estuaries. 14, 49 - 56

Jimenez, A. G., and Bennett, W. A. 2005 Respiratory physiology of three Indo-Pacific

fiddler crabs: metabolic responces to intertidal zonation patterns. Crustaceana. 78,

965 – 974

Jokiel, P. L., Rodgers., K. S., Kuffner., I. B., Andersson, A. J., Cox, E. F., and Mackenzie, F.

T. 2008 Ocean acidification and calcifying reef organisms: a mesocosm

investigation. Coral Reefs. 27, 473 - 483

Jones, E. B. G., and Byrne, P. J. 1976 Physiology of the higher marine fungi. In Recent

advances in aquatic mycology (ed. E. B. G. Jones), pp. 135-175. Elek Science:

London

Jones, C. G., Lawton, J. H., and Shachak, M. 1994 Organisms as ecosystem engineers.

Oikos. 69, 373 – 386 230

Jones, D. T., and Eggleton, P. 2000 Sampling termite assemblages in tropical forests: testing

a rapid biodiversity assessment protocol. Journal of Applied Ecology. 37, 191 - 203

Jones, M., Mandelik, Y., and Dayan, T. 2001 Coexistence of temporally partitioned spiny

mice: roles of habitat structure and foraging behaviour. Ecology. 82, 2164 – 2176

Joshi, H. and Ghose, M. 2003 Forest structure and species distribution along soil salinity and

pH gradient in mangrove swamps of the Sundarbans. Tropical Ecology. 44, 197 - 206

Kairo, J. G., Bosire, J., Langat, J., Kirui, B., and Koedam, N. 2009 Allometry and biomass

distribution in replanted mangrove plantations at Gazi Bay, Kenya. Aquatic

Conservation: Marine and Freshwater Ecosystems. 19, 63 - 69

Kalmis, E., Yasa, I., Kalyoncu, F., Pazarbasi, B., and Kocyigit, A. 2008 Ligninolytic enzyme

activities in mycelium of some wild and commercial mushrooms. African Journal of

Biotechnology. 23, 4314 - 4320

Kappes, H., Jabin, M., Kulfan, J., Zach, P., and Top, W. 2009 Spatial patterns of litter-

dwelling taxa in relation to the amounts of coarse woody debris in European

temperate deciduous forests. Forest Ecology and Management. 257, 1255 - 1260

Karban, R., and Myers, J. H. 1989 Induced plant responses to herbivory. Annual Review of

Ecology and Systematics. 20, 331 – 348

Karban, R., Agrawal, A. A., and Mangel, M. 1997 The benefits of induced defences against

herbivores. Ecology. 78, 1351 - 1355

Karieiva, P. 1982 Experimental and mathematical analyses of herbivore movement:

quantifying the influence of plant spacing and quality of foraging discrimination.

Ecological Monographs. 52, 261 – 282

Karlsson, O. M., Richardson, J. S., and Kiffney, P. M. 2005 Modelling organic matter

dynamics in headwater streams of south-western British Columbia, Canada.

Ecological Modelling. 183, 463 - 476 231

Kathiresan, K. and Bingham, L. B. 2001 Biology of Mangroves and Mangrove Ecosystems.

Advances in Marine Biology. 40, 81 - 251

Keith, P., Bosc, P. and Valade, P. 2004 A new species of Parioglossus (Gobioidei,

Ptereleotridae) from Seychelles islands. Cybium. 28, 341 - 344

Kenyon, R. A., Loneragan, N. R., Manson, F. J., Vance, D. J., and Venables, W. N. 2004

Allopatric distribution of juvenile red-legged banana prawns (Penaeus indicus H.

Milne Edwards, 1837) and juvenile white banana prawns (Penaeus merguiensis De

Man, 1888), and inferred extensive migration, the Joseph Bonaparte Gulf, northwest

Australia. Journal of Experimental Marine Biology and Ecology. 309, 79 – 108

Kjerfve, B., and Macintosh, D. J. 1997 The impact of climate change on mangrove

ecosystems. Proceedings of mangrove ecosystem studies in Latin America and

Africa sponsored by UNESCO, Brazil. 1993. pp.1 - 7

Kleypas, J. A. 1996 Coral reef development under naturally turbid conditions: fringing reefs

near Broad Sound, Australia. Coral reefs. 15, 153 - 167

Kohlmeyer, J., Bebout, B. and Volkmann-Kohlmeyer, B. 1995 Decomposition of mangrove

wood by marine fungi and teredinids in Belize. Marine Ecology. 16, 27 – 39

Komiyama, A., Poungparn, S., and Kato, S. 2005 Common allometric equations for

estimating the tree weight of mangroves. Journal of Tropical Ecology. 21, 471 - 477

Komiyama, A., Ong, J. E., and Poungparn, S. 2008 Allometry, biomass, and productivity of

mangrove forests: A review. Aquatic Botany. 89, 128 - 137

Korb, J., and Lenz, M. 2004 Reproductive decision-making in the termite, Cryptotermes

secundus (Kalotermitidae), under variable food conditions. Behavioural Ecology.

15, 390 – 395

Korb, J. 2008 The ecology of social evolution in termites. In: Ecology of Social Evolution

(eds. Korb, J., and Heinze, J.), Springer Press, Heidelberg, pp. 151 – 174 232

Kostylev, V. E., Erlandsson, J., Ming, M. Y., and Williams, G. A. 2005 The relative

importance of habitat complexity and surface area in assessing biodiversity: Fractal

application on rocky shores. Ecological Complexity. 2, 272 - 286

Krauss, K. W., Doyle, T. W., Twilley, R. R., Smith, T. T., Whelan, K. R. T., and Sullivan, J.

K. 2005 Woody debris in the mangrove forests of south Florida. Biotropica. 37, 9 -

Krauss, K. W., Lovelock, C. E., Mckee, K. L., Lopez-Hoffman, L., Ewe, S. M. L., and

Sousa, W. P. 2008 Environmental drivers in mangrove establishment and early

development: a review. Aquatic Botany. 89, 105 - 127

Kristensen, E., Bouillon, S., Dittmar, T., and Marchand, C. 2008 Organic carbon dynamics

in mangrove ecosystems: A review. Aquatic Botany. 89, 201 - 219

Kruitwagen, G., Nagelkerken, I., Lugendo, B. R., Mgaya, Y. D., and Bonga, W. 2010

Importance of different carbon sources for macroinvertebrates and fishes of an

interlinked mangrove-mudflat ecosystem (Tanzania). Estuarine, Coastal and Shelf

Science. 88, 464 – 472

Kurihara, H., Shimode, S., and Shirayama, Y. 2004 Sub-lethal effects of elevated

concentration of CO2 on planktonic copepods and sea urchins. Journal of

Oceanography. 60, 743 - 750

Kurihara, H., Kato, S., and Ishimatsu, A. 2007 Effects of increased seawater pCO2 on early

development of the oyster Crassostrea gigas. Aquatic Biology. 1, 91 - 98

Kwakkel-Hol, A. 2006 Tourism in Wakatobi Marine National Park. http://www.coral-

triangle.org. Accessed May 2012.

Laegdsgaard, P., and Johnson, C. 2001 Why do juvenile fish utilize mangrove habitats?

Journal of Experimental Marine Biology and Ecology. 257, 229 – 253

233

Lambert, R. L., Lang, G. E., and Reiners, W. A. 1980 Loss of mass and chemical change in

decaying boles of a subalpine balsam fir forest. Ecology. 61, 1460 - 1473

Lang, G. E., and Knight, D. H. 1979 Decay rates for the boles of tropical trees in Panama.

Biotropica. 11, 316 - 317

Larson, M. G., Booth, D. B., and Morley, S. A. 2001 Effectiveness of large woody debris in

stream rehabilitation projects in urban basins. Ecological Engineering. 18, 211 – 226

Laurance, W. F., Lovejoy, T. E., Vasconcelos, H. L., Bruna, E. M., Didham, R. K., Stouffer,

P. C., Gascon, C., Bierregaard, R. O., Laurance, S. G., and Sampaio, E. 2002

Ecosystem decay of Amazonian forest fragments: a 22-year investigation.

Conservation Biology. 16, 605 - 618

Laverty, R. 2002 Advanced engineered wood composits. Forest Products Journal. 52, 18 –

Lennartsson, T., Nilsson, P., and Tuomi, J. 1998 Induction of overcompensation in the field

gentian, Gentianella campestris. Ecology. 79, 1061 – 1072

Leonel, M. V. R., Lopes, S. G. B. C., de Moraes, D. T., and Aversari, M. 2006 The

interference of methods in the collection of teredinids (Mollusca, Bivalvia) in

mangrove habitats. Iheringia, Series of Zoology. 96, 25 - 30

Leponce, M., Roisin, Y., and Pasteels, J. M. 1995 Environmental influences on the arboreal

nesting termite community in New Guinean Coconut plantations. Community and

Ecosystem Ecology. 24, 1442 - 1452

Levi, H. W. 1967 Adaptations of respiratory systems of spiders. Evolution. 3, 571 – 583

Levin, S. A. 1992 The problem of pattern and scale in ecology: The Robert H. MacArthur

award lecture. Ecology. 73, 1943 - 1967

Lewis, M., Pryor, R., and Wilking, L. 2011 Fate and effects of anthropogenic chemicals in

mangrove ecosystems: A review. Environmental Pollution. 159, 2328 – 2346 234

Lira-Medeiros, C. F., Parisod, C., Fernandes, R. A., Mata, C. S., Cardoso, M. A., and

Ferreira, P. C. G. 2010 Epigenetic variation in variation in mangrove plants

occurring in contrasting natural environment. PloS One. 5, 1 - 8

Lonsdale, D., Pautasso, M., and Holdenrieder, O. 2008 Wood-decaying fungi in the forest -

conservation needs and management options. European Journal of Forest Research.

127, 1-22

Lugo, A. E. 1980 Mangrove ecosystems: Successional or steady state? Biotropica. 12, 65 -

Lugo, A. E., and Snedaker, S. C. 1974 The ecology of mangroves. Annual Review of

Ecology, Evolution and Systematics. 5, 39 - 64

Luo, J., Serafy, J. E., Sponaugle, S., Teare, P. B., and Kieckbusch, D. 2009 Movement of

grey snapper Lutjanus griseus among subtropical seagrass, mangrove, and coral reef

habitats. Marine Ecology Progress Series. 380, 255 – 269

MacDonald, J. A., Glover, T., and Weis, J. S. 2008 The impact of mangrove prop-root

epibionts on juvenile reef fishes: A field experiment using artificial roots and

epifauna. Estuaries and Coasts. 31, 981 – 993

MacDonald, J. A., Shahrestani, S., and Weis, J. S. 2009 Behaviour and space utilization of

two common fishes within Caribbean mangroves: implications for the protective

function of mangrove habitats. Estuarine, Coastal and Shelf Science. 84, 195 – 201

MacIntosh, H., Nys, R. D., and Whelan, S. 2012 Shipworms as a model for competition and

coexistence in specialized habitats. Marine Ecology Progress Series. 461, 95 - 105

Magurran, A. E. 2004 Measuring Biological Diversity. Blackwell Publishing: Oxford, UK

Malhi, Y., Meir, P., and Brown, S. 2002 Forests, carbon and global climate. Philosophical

Transactions of the Royal Society A: Physical, Mathematical and Engineering

Sciences. 360, 1567 - 1591 235

Manyak, D. M. 1982 A Device for the collection and study of Wood-boring Molluscs:

Application to Boring Rates and Boring Movements of the Shipworm Bankia gouldi.

Estuaries. 5, 224 - 229

Marchand, C., Lallier-Verges, E., and Baltzer, F. 2003 The composition of bulk sedimentary

organic matter in relation to the dynamic features of a mangrove-fringed coast in

French Guiana. Estuarine Coastal and Shelf Sciences. 56, 119 - 130

Marczak, L. B., and Richardson, J. S. 2007 Spiders and subsides: results from the riparian

zone of a coastal temperate rainforest. Journal of Animal Ecology. 76, 687 – 694

Maria, G. L., and Sridhar, K. R. 2002 Richness and diversity of filamentous fungi on woody

litter of mangroves along the West coast of India. Current Science. 83, 1573 - 1580

Maria, G. L., and Sridhar, K. R. 2003 Diversity of filamentous fungi on woody litter of five

mangrove plant species from the southwest coast of India. Fungal Diversity. 14, 109

- 126

Martinez-Alier, J. 2001 Ecological conflicts and valuation: mangroves versus shrimps in the

late 1990s. Environment and Planning C: Government and Policy. 19, 713 - 728

McArdle, B. H., and Anderson, M. J. 2001 Fitting multivariate models to community data: a

comment on distance-based redundancy analysis. Ecology. 82, 290 - 297

McDowall, R.M 2001 Parioglossus (Teleostei: Gobioidei: Microdesmidae) in New Zealand.

New Zealand Journal of Marine and Freshwater Research. 35, 165 – 172

McElhinny, C., Gibbons, P., Brack, C., and Bauhus, J. 2005 Forest and woodland stand

structural complexity: its definition and measurement. Forest Ecology and

Management. 218, 1 - 14

McInnes, K. L., Walsh, K. J. E., Hubbert, G. D., and Beer, T. 2003 Impact of sea-level rise

and storm surges on a coastal community. Natural Hazards. 30, 187 - 207 236

McKee, K. L., and Faulkner, P. L. 2000 Restoration of biochemical function in mangrove

forests. Restoration Ecology. 8, 247 – 259

Mckeon, C. S., and Feller, I. C. 2004 The supratidal fauna of Twin Cays, Belize. National

Museum of Natural History, Smithsonian environmental research center. Atoll

Research Bulletin. 526, 1 - 22

McLay, C. L., and Hayward, T. L. 1987 Reproductive biology of the intertidal spider Desis

marina (Araneae: Desidae) on a New Zealand rocky shore. Journal of Zoology. 2,

357 – 372

McLeod, E., and Salm, R. V. 2006 Managing mangroves for resilience to climate change.

IUCN. Gland, Switzerland. 64pp

McQueen, D. J., and McLay, C. L. 1983 How does the intertidal spider Desis marina

(Hector) remain under water for such a long time? New Zealand Journal of Zoology.

10, 383 – 392

McWilliams, J. P., Cote, I. M., Gill, J. A., Sutherland, W. J., and Watkinson, A. R. 2005

Accelerating impacts of temperature-induced coral bleaching in the Caribbean.

Ecology. 86, 2055 - 2060

Merritt, N. R. C., and Starr, C. K. 2010 Comparative nesting habits and colony composition

of three arboreal termites (Isoptera: Termitidae) in Trinidad and Tobago, West

Indies. Sociobiology. 56, 611 – 622

Michener, W. K., Blood, E. R., Bildstein, K. L., Brinson, M. M., and Gardner, L. R. 1997

Climate change, hurricanes and tropical storms, and rising sea level in coastal

wetlands. Ecological Applications. 7, 770 - 801

Middleton, A. B., and McKee, L. K. 2001 Degradation of mangrove tissues and implications

for peat formation in Belizean island forests. Journal of Ecology. 89, 818 - 828 237

Morse, D. H. 1997 Distribution, movement , and activity patterns of an intertidal wolf spider

Pardosa lapidicina population (Araneae, Lycosidae). The Jounal of Arachnology. 25,

1 – 10

Morton, B. 1983 Mangrove Bivalves. In: Wilbur, K. W. (ed.) The Mollusca, Vol. 2.

Academic Press, New York. p. 77 - 139

Muller-Karger, F. E., Varela, R., Thunell, R., Luerssen, R., Hu, C., and Walsh, J. J. 2005

The importance of continental margins in the global carbon cycle. Geophysical

Research Letters. 32, 1 – 4

Mumby, P. J., Edwards, A. J., Arias-Gonzalez, J. E., Lindeman, K. C., Blackwell, P. G.,

Gall, A., Gorczynska, M. I., Harborne, P. G., Pescod, C. L., Renken, H., Wabnitz, C.

C. C., and Llewellyn, G. 2003 Mangroves enhance the biomass of coral reef fish

communities in the Caribbean. Nature. 427, 533 – 536

Nagelkerken, I., and van der Velde, G. 2004 Are Caribbean mangroves important feeding

grounds for juvenile reef fish from adjacent seagrass beds? Marine Ecology Progress

Series. 274, 143 – 151

Nagelkerken, I., Blaber, S. J. M., Bouillon, S., Green, P., Haywood, M., Kirton, L. G.,

Meynecke, J. O., Pawlik, J., Penrose, H. M., Sasekumar, A. & Somerfield, P J. 2008

The habitat function of mangroves for terrestrial and marine fauna: A review.

Aquatic Botony. 89, 155 – 185

Nagelkerken, I., Schryver, A. M., Verweij, M. C., Dahdouh-Guebas, F., Van der Velde, G.,

and Koedam, N. 2010 Differences in root architecture influence attraction of fishes

to mangroves: A field experiment mimicking roots of different lengths, orientation

and complexity. Journal of Experimental Marine Biology and Ecology. 396, 27 - 34

Ng, P. K. L., and Sivasothi, N. 1999 A guide to the mangroves of Singapore. Singapore

Science Center, Singapore 238

Nilsson, T. 1973 Studies on wood degradation and cellulolytic activity of microfungi. Studia

Forestalia Suecica. No. 104

Nisbet, T. R. 2001 The role of forest management in controlling diffuse pollution in UK

forestry. Forest Ecology and Management. 143, 215 - 226

Nordhaus, I., and Wolff, M. 2007 Feeding ecology of the mangrove crab Ucides cordatus

(Ocypodidae): food choice, food quality and assimilation efficiency. Marine Biology.

151, 1665 – 1681

Novotny, V., Drozd, P., Miller, S. E., Kulfan, M., Janda, M., Basset, Y., and Weiblen, G. D.

2006 Why are so many species of herbivorous insects in tropical rainforests?

Science. 313, 1115 – 1118

Novotny, V., Miller, S. E., Hulcr, J., Drew, R. A. I., Basset, Y., Janda, M., Setliff., Darrow,

K., Stewart, A. J. A., Auga, J., Isua, B., Molem, K., Manumbor, M., Tamtiai, E,

Mogia, M., and Weiblen, G. D. 2007 Low beta diversity of herbivorous insects in

tropical forests. Nature. 448, 692 - 695

Nurminen, L., Horppila, J., and Pekcan-Hekim, Z. 2007 Effect of light and predator

abundance on the habitat choice of plant-attached zooplankton. Freshwater Biology.

52, 539 – 548

O’Connor, N. A. 1991 The effects of habitat complexity on the macroinvertebrates

colonising wood substrates in a lowland stream. Oecologia. 85, 504 - 512

Ong, J. E., Khoon, G. W., and Clough, B. F. 1995 Structure and productivity of a 20-year-

old stand of Rhizophora apiculata Bl. mangrove forest. Journal of Biogeography. 22,

417-424

Ong, J. E., Gong, W. K., and Wong, C. H. 2004 Allometry and partitioning of the mangrove,

Rhizophora apiculata. Forest Ecology and Management. 188, 395 - 408 239

Paige, K. N., and Whitham, T. G. 1987 Overcompensation in response to mammalian

herbivory: the advantage of being eaten. The American Naturalist. 129, 407 – 416

Paige, K. N. 1992 Overcompensation in response to mammalian herbivory: from mutualistic

to antagonistic interactions. Ecology. 73, 2076 – 2085

Paige, K. N. 1999 Regrowth following ungulate herbivory in Ipomopsis aggregata:

geographic evidence for overcompensation. Oecologia. 118, 316 – 323

Pandolfi, J. M., Bradbury, R. H., Sala, E., Hughes, T. P., Bjorndal, K. A., Cooke, R. G.,

McArdle, D., McClenachan, L., Newman, M. J. H., Paredes, G., Warner, R. R., and

Jackson, J. B. C. 2003 Global trajectories of the long-term decline of coral reef

ecosystems. Science. 301, 955 - 958

Parker, L. M., Ross, P. M., and O’Connor, W. A. 2009 The effect of ocean acidification and

temperature on the fertilization and embryonic development of the Sydney rock

oyster Saccostrea glomerata (Gould 1850). Global Change Biology. 15, 2123 - 2136

Perry, D. M. 1988 Effects of associated fauna on growth and productivity in the red

mangrove. Ecology. 69, 1064 - 1075

Petillon, J., Georges, A., Canard, A., Lafeuvre, J-C, Bakker, J. P., and Ysnel, F. 2008

Influence of abiotic factors on spider and ground beetle communities in different salt-

marsh systems. Basic and Applied Ecology. 9, 743 – 751

Petillon, J., Montaigne, W., and Renault, D. 2009 Hypoxia coma as a strategy to survive

inundation in a salt-marsh inhabiting spider. Biology Letters. 5, 442-445

Petren, K., and Case, T. J. 1998 Habitat structure determines competition intensity and

invasion success in gecko lizards. Proceedings of the National Acadamy of Science,

USA. 95, 11739 – 11744

Pi, N., Tam, N. F. Y., Wu, Y., and Wang, M. H. 2009 Root anatomy and spatial pattern of

radial oxygen loss of eight true mangrove species. Aquatic Botany. 90, 222-230 240

Pierce, S. K. 1982 Invertebrate cell volume control mechanisms: A coordinated use of

intracellular amino acids and inorganic ions as osmotic solute. Biological Bulletin.

163, 405 – 419

Pilgrim, S., Cullen, L., Smith, D., and Pretty, J. 2007 Hidden harvest or hidden revenue – a

local resource use in a remote region of Southeast Sulawesi, Indonesia. Indian

Journal of Traditional Knowledge. 6, 150 – 159

Plaziat, J.C. 1984 Mollusc distribution in the mangal. In: Por, F.D., Dor, I. (eds.),

Hydrobiology of the Mangal, The Ecosystem of the Mangrove Forests. Developments

in Hydrobiology 20. Dr. W. Junk Publishers, The Hague.

Poff, N. L., and Allan, D. J. 1995 Functional organization of stream fish assemblages in

relation to hydrological variability. Ecology. 76, 606 – 627

Polidoro, B. A., Carpenter, K. E., Collins, L., Duke, N. C., Ellison, A. M., Ellison, J. C.,

Farnsworth, E. J., Fernando, E. S., Kathiresan, K., Koedam, N. E., Livingstone, S. R.,

Miyagi, T., Moore, G. E., Nam, V. N., Ong, J. E., Primavera, J. H., Salmo III, S. G.,

Sanciangco, J. C., Sukardjo, S., Wand, Y., and Yong, J. W. H. 2010 The loss of

species: mangrove extinction risk and geographic areas of global concern. PLoS

ONE. 5, e10095.doi:10.1371/journal.pone.0010095

Polis, G. A., Anderson, W. B., and Holt, R. D. 1997 Toward an integration of landscape and

food web ecology: The dynamics of spatially subsidized food webs. Annual Review

of Ecology and Systematics. 28, 289 – 316

Prasad, M. B. K., and Ramanathan, A. L. 2008 Sedimentary nutrient dynamics in a tropical

estuarine mangrove ecosystem. Estuarine, Coastal and Shelf Science. 80, 60 – 66

Prasannarai, K., and Sridhar, K. R. 2003 Abundance and diversity of marine fungi on

intertidal woody litter of the West coast of India on prolonged incubation. Fungal

Diversity. 14, 127-141 241

Proffitt, E. C., and Devlin, D. J. 2005 Grazing by the intertidal gastropod Melampus coffeus

greatly increases mangrove leaf litter degradation rates. Marine Ecology Progress

Series. 296, 209 – 218

Progar, R. A., Schowalter, T. D., Freitay, C. M., and Morrell, J. J. 2000 Respiration from

coarse woody debris as affected by moisture and saprotroph functional diversity in

western Oregon. Oecologia. 124, 426-431

Ravikumar, M., Sridhar, K. R., Sivakumar, T., Karamchand, K. S., Sivakunmar, N., and

Vellaiyan, R. 2009 Diversity of filamentous fungi on coastal woody debris after

tsunami on the southeast coast of India. Czech Mycology Journal. 61, 107-115

Rayner, S. M. 1983 Distribution of teredinids (Mollusca: Teredinidae) in Papua New

Guinea. Records of the Australian Museum. 35, 61 – 76

Rehm, A. E. 1976 The effects of the wood-boring isopod Sphaeroma terebrans on the

mangrove communities of Florida. Environmental Conservation. 3, 47 - 57

Rehm, A., and Humm, H. J. 1973 Sphaeroma terebrans: a threat to the mangroves of

southwestern Florida. Science. 182, 173 - 174

Reiss, H., and Kroncke, I. 2005 Seasonal variability of benthic indices: An approach to test

the applicability of different indices for ecosystem quality assessment. Marine

Pollution Bulletin. 50, 1490 – 1499

Rennis, D.S. and Hoese, D. F 1985 A review of the genus Parioglossus, with descriptions of

six new species (Pisces: Gobioidei). Records of the Australian Museum. 36, 169–201

Riechert, S. E., and Tracy, R. C. 1975 Thermal balance and prey availability: bases for a

model relating web-site characteristics to sider reproductive success. Ecology. 56,

265 – 284 242

Ribi, G. 1982 Differential colonization of roots of Rhizophora mangle by the woodboring

isopod Sphaeroma terebrans as a mechanism to increase root density. Marine

Ecology. 3, 13 - 19

Ries, J. B., Cohen, A. L., and McCorkle, D. C. 2009 Marine calcifiers exhibit mixed

responses to CO2 induced ocean acidification. Geological Society of America. 12,

1131-1134

Rimmer, M. A., Battaglene, S. L., Dostine, P. L. 1983 Observations on the Distribution of

Bankia australis Calman (Mollusca: Teredinidae) in the Patonga Creek Mangrove

Swamp, New South Wales. Australian Journal of Marine and Freshwater Research.

34, 355–357

Roberts, C. M., McClean, C. J., Veron, J. E. N., Hawkins, J. P., Allen, G. R., McAllister, D.

E., Mittermeier, C. G., Schueler, F. W., Spalding, M., Wells, F., Vynne, C., and

Werner, T. B. 2002 Marine biodiversity hotspots and conservation priorities for

tropical reefs. Science. 295, 1280 – 1284

Robertson. I. A., 1985 The determination of trophic relationships in mangrove-dominated

systems: Areas of Darkness. Proceedings of the Research for Development Seminar

held at the Australian Institute of Marine Sciences. 291 - 304

Robertson, A. I., and Daniel, P. A. 1989a Decomposition and the annual flux of detritus

from fallen timber in tropical mangrove forests. Limnology and Oceanography. 34,

640 - 646

Robertson, A. I. 1990 Plant-animal interactions and the structure and function of mangrove

forest ecosystems. Australian Journal of Ecology. 16, 433 - 443

Roisin, Y., Dejean, A., Corbara, B., Orivel, J., Samaniego, M., and Leponce, M. 2006

Vertical stratification of the termite assemblage in a neotropical rainforest.

Oecologia. 149, 301 – 311 243

Ronnback, P., Troell, M., Kautsky, N., and Primavera, J. H. 1999 Distribution pattern of

shrimp and fish among Avicennia and Rhizophora microhabitats in the Pagbilao

mangroves, Philippines. Estuarine, Coastal and Shelf Science. 48, 223 – 234

Ronnback, P., Macia, A., Almqvist, G., Schultz, L., and Troell, M. 2002 Do penaeid shrimps

have a preference for mangrove habitats? Distribution pattern analysis on Inhaca

Island, Mozambique. Estuarine, Coastal and Shelf Science. 55, 427 – 436

Roonwal, M. L. 1954 The marine borer, Bactronophorus thoracities (Gould) [Mollusca,

Eulamellibranchiata, Teredinidae], as a pest of living trees in the mangrove forests of

the Sunderbans, Bengal, India. Proceedings of the Zoological Society. 7, 91 – 103

Roth, V. D. 1967 Descriptions of the spider families Desidae and Argyronetidae. American

Museum Novitates. 2292, 1 - 9

Roth, L. C. 1992 Hurricanes and mangrove regeneration: Effects of Hurricane Joan, October

1988, on the vegetation of Isla del Venado, Bluefields, Nicaragua. Biotropica. 24,

375 – 384

Roth, B. M., Kaplan, I. C., Sass, G. G., Johnson, P. T., Marburg, A. E., Yannarell, A. C.,

Havlicek, T. D., Willis, T. V., Turner, M. G., and Carpenter, S. R. 2007 Linking

terrestrial and aquatic ecosystems. The role of woody habitat in lake food webs.

Ecological Modelling. 203, 439 – 452

Rothenbucher, J., and Schaefer, M. 2006 Submersion tolerance in floodplain arthropod

communities. Basic and Applied Ecology. 7, 398 – 408

Rovner, J. S. 1987 Nests of terrestrial spiders maintain a physical gill: flooding and the

evolution of silk constructions. Journal of Arachnology. 14, 327 – 337

Royal Society, 2005 Ocean acidification due to increasing atmospheric carbon dioxide.

Policy document 12/05 Royal Society, London, The Clyvedon Press. Ltd, Cardiff

244

Ruxton, G. D., Sherrett, T. N., and Speed, M. P. 2004 Avoiding Attack: The Evolutionary

Ecology of Crypsis, Warning Signals and Mimicry. Oxford, UK: Oxford University

Press

Saha, N., Aditya, G., and Saha, G. K. 2009 Habitat complexity reduces prey vulnerability:

An experimental analysis using aquatic insect predators and immature dipteran prey.

Journal of Asia-Pacific Entomology. 12, 233 – 239

Sanchirico, J. N., and Mumby, P. 2009 Mapping ecosystem functions to the variation of

ecosystem services: implications of species-habitat associations for coastal land-use

decisions. Theoretical Ecology. 2, 67 – 77

Sarma, V. V., and Vittal, B. P. R. 2001 Biodiversity of manglicolous fungi on selected plants

in the Godavari and Krishna deltas, East coast of India. Fungal Diversity. 6, 115 -

130

Sasekumar, A. and Chong, V. C. 1998 Faunal diversity in Malaysian mangroves. Global

Ecology and Biogeography Letters. 7, 57 - 60

Scheltema, R. S. 1971 Dispersal of phytoplanktotrophic Shipworm larvae (Bivalvia:

Teredinidae) over long distances by ocean currents. Marine Biology. 11, 5 - 11

Schoener, T. W. 1971 Theory of feeding strategies. Annual Review of Ecology and

Systematics. 2, 369 – 404

Schories, D., Barletta-Bergan, A., Barletta, M., Krumme, U., Meklig, U., and Rademaker, V.

2003 The keystone role of leaf-removing crabs in mangrove forests of North Brazil.

Wetlands Ecology and Management. 11, 243 - 255

Selden, P. A. 2001 Eocene spiders from the Isle of Wight with preserved respiratory

structures. Palaeontology. 44, 695 – 729

Seymour, R. S., and Hetz, S. K. 2011 The diving bell and the spider: the physical gill of

Argyroneta aquatica. The Journal of Experimental Biology. 214, 2175 – 2181 245

Sheaves, M. 2005 Nature and consequences of biological connectivity in mangrove systems.

Marine Ecology Progress Series. 302, 293 - 305

Sherman, E. R., Fahey, J. T. and Martinez, P. 2003 Spatial patterns of biomass and

aboveground net primary productivity in a mangrove ecosystem in the Dominican

Republic. Ecosystems. 6, 384 - 398

Shields, J. R. F. D., Knight, S. S., and Stofleth, J. M. 2006 Large woody addition for aquatic

habitat rehabilitation in an incited, sand-bed stream, Little Topashaw Creek,

Mississippi. River Research and Applications. 22, 803 – 817

Shirvell, C. S. 1990 Role of instream rootwads as juvenile Coho salmon (Oncorhynchus

kisutch) and steelhead trout (O. mykiss) cover habitat under verying streamflows.

Canadian Journal of Fisheries and Aquatic Sciences. 47, 852 – 861

Si, A., Bellwood, O., and Alexander, C. G. 2002 Evidence for filter-feeding by the wood-

boring isopod, Sphaeroma terebrans (Crustacea: Peracarida). The Journal of

Zoology. 256, 463 - 471

Simberloff, D., Brown, B. J., and Lowrie, S. 1978 Isopod and insect root borers may benefit

Florida mangroves. Science. 201, 630 – 632

Smith, N. F., Wilcox, C., and Lessmann, J. M. 2009 Fiddler crab burrowing affects growth

and production of the white mangrove (Laguncularia racemosa) in a restored Florida

coastal marsh. Marine Biology. 156, 2255 - 2266

Smock, L. A., Metzler, G. M., and Gladden, J. E. 1989 Role of debris in the structure and

functioning of low-gradient headwater streams. Ecology. 70, 764 – 775

Sousa, W. P., Quek, S. P., Mitchell, B. J. 2003 Regeneration of Rhizophora mangle in a

Caribbean mangrove forest: interacting effects of canopy disturbance and a stem-

boring beetle. Oecologia. 137, 437 – 445

246

Spalding, M. D., Blasco, F., and Field, C. D. (eds) 1997 World Mangrove Atlas. The

international society for mangrove ecosystems, Okinawa, Japan

Spanhoff, B., Alecke, C., and Meyer, E. I. 2001 Simple method for rating the decay stages of

submerged woody debris. Journal of the North American Benthological Society. 20,

385 – 394

Spurr, E. B., and McGregor, P. G. 2003 Potential invertebrate antifeedants for toxic baits

used for vertebrate pest control. Science for Conservation. 232, 1 - 36

Storry, K. A., Weldrick, C. K., Mews, M., Zimmer, M., and Jelinski, D. E. 2006 Intertidal

coarse woody debris: A spatial subsidy as shelter or feeding habitat for gastropods?

Estuarine, Coastal and Shelf Science. 66, 197 – 203

Strauss, S. Y., and Agrawal, A. A. 1999 The ecology and evolution of plant tolerance to

herbivory. Tree. 14, 179 – 185

Suzuki, T. and Senou, H. 1994 Parioglossus interruptus, a new species of goby from the

western Pacific. Japanese Journal of Ichthyology. 41, 281–286

Svavarsson, J., Melckzedeck, O. K. W., and Olafsson, E. 2002 Does the wood-borer

Sphaeroma terebrans (Crustacea) shape the distribution of the mangrove Rhizophora

mucronata ? Ambio. 31, 574 – 579

Syms, C., and Jones, G. P. 2000 Disturbance, habitat structure, and the dynamics of a coral-

reef fish community. Ecology. 81, 2714 – 2729

Taniguchi, H., and Tokeshi, M. 2004 Effects of habitat complexity on benthic assemblages

in a variable environment. Freshwater Biology. 49, 1164 - 1178

Tank, J. L., and Winterbourn, M. J. 1995 Biofilm development and invertebrate colonization

of wood in four New Zealand streams of contrasting pH. Freshwater Biology. 34,

303 – 315 247

Tank, J. L., and Winterbourn, M. J. 1996 Microbial activity and invertebrate colonisation of

wood in a New Zealand forest stream. New Zealand Journal of Marine and

Freshwater Research. 30, 271 – 280

Taylor, J. R., Cook, M. M., Kirkpatrick, S. N., Galleher., Eme, J., and Bennett, W. A. 2005

Thermal tactics of air-breathing and non air-breathing Goboids inhabiting mangrove

tidepools on Palau Hoga, Sulawesi, Indonesia. Copeia. 4, 886 - 893

Taylor, D.S., Turner, B.J., Davis, W.P. & Chapman, B.B. 2008 A novel terrestrial fish

habitat inside emergent logs. American Naturalist. 171, 263 – 266

Tedersoo, L., Koljalg, U., Hallenberg, N., and Larsson, K. H. 2003 Fine scale distribution of

ectomycorrhizal fungi and roots across substrate layers including coarse woody

debris in a mixed forest. New Phytologist. 159, 153 - 165

The Singapore Red Data Book. 2008 www. wildsingapore. com/wildfacts /arachnida/

desis.htm. Accessed May 2010

Tomascik, T., Mah, A. J., Nontji, A., and Moosa, M. K. 1997 The Ecology of the Indonesian

Seas Part Two. The ecology of Indonesia series, volume VIII. Periplus Editions,

Singapore.

Tse, P., Nip, T. H. M., and Wong, C. K. 2008 Nursery function of mangrove: a comparison

with mudflat in terms of fish species composition and fish diet. Estuarine, Coastal

and Shelf Science. 80, 235 – 242

Turner, R. D. 1966 A Survey and Illustrated Catelogue of the Teredinidae (Mollusca:

Bilvalvia). Harvard, The Museum of Comparative Zoology. pp. 265

Turner, R.D., 1971 Identification of Marine Wood Boring Molluscs. In: Jones, E.B.G.,

Eltringham, S.K. (Eds.), Marine Borers and Fouling Organisms of Wood. OECD,

Paris, pp. 17 – 64 248

Turner, R. D. 1976 Some Factors Involved in the Settlement and Metamorphosis of Marine

Bivalve Larvae. In: J. M., Sharpley and A. M. Kaplan, eds., Proceedings of the 3rd

International Biodegradation Symposium. Applied Science, Barkin, UK. 409 - 416

Unsworth, R. K. F., De Leon, P. S., Garrard, S. L., Jompa, J., Smith, D. J., and Bell. J. J.

2008 High connectivity of Indo-Pacific seagrass fish assemblages with mangrove

and coral reef habitats. Marine Ecology Progress Series. 353, 213 – 224

Valiela, I., Bowen, J. L., and York, J. K. 2001 Mangrove forests: one of the world’s

threatened major tropical environments. Bioscience. 51, 807 – 815

Verweij, M. C., Nagelkerken, I., de Graaff, D., Peeters, M., Bakker, E. J., and van der Velde,

G. 2006 Structure, food and shade attract juvenile coral reef fish to mangrove and

seagrass habitats: a field experiment. Marine Ecology Progress Series. 306, 257 –

268

Vittal, B. P. R., and Sarma, V. V. 2006 Diversity and ecology of fungi on mangroves of Bay

of Bengal region – an overview. Indian Journal of Marine Sciences. 35, 308 - 317

Walters, B. B. 2005a Patterns of local wood use and cutting of Philippine mangrove forests.

Economic Botany. 59, 66 – 76

Walters, B. B. 2005b Ecological effects of small-scale cutting of Philippine mangrove

forests. Forest Ecology and Management. 206, 331 - 348

Walters, B. B., Rounback, P., Kovacs, J. M., Crona, B., Hussain, S. A., Badola, R.,

Primavera, J. H., Barbier, E., and Dahdouh-Guebas, F. 2008 Ethnobiology, socio-

economics and management of mangrove forests: A review. Aquatic Botany. 89, 220

- 236

Wang, M., Huang, Z., Shi, F., and Wang, W. 2009 Are vegetated areas of mangroves

attractive to juvenile and small fish? The case of Dongzhaigang Bay, Hainan Island,

China. Estuarine, Coastal and Shelf Science. 85, 208 – 216 249

Warfe, D. M., and Barmuta, L. A. 2004 Habitat complexity mediates the foraging success of

multiple predator species. Oecologia. 141, 171 – 178

Weedon, J. T., Cornwell, W.K., Cornelisson, H. C., Zanne, A. E., Wirth, C., and Coomes, D.

A. 2009 Global meta-analysis of wood decomposition rates: a role for trait variation

among tree species. Ecology Letters. 12, 45 - 56

Wier, A. M., Schnitzler, M. A., Tattar, T. A., Klekowski, JR, E. J., and Stern, A. I. 1996

Wound periderm development in red mangrove, Rhizophora mangle L. International

Journal of Plant Science. 157, 63 - 70

Wilkie, M. L., and Fortuna, S. 2003 Status and trends in mangrove area extent Worldwide,

Indonesia. http://www.fao.org/docrep/007/j1533e/j1533E46.htm. Accessed July 2010

Willis, S. C., Winemiller, K. O., and Lopez-Fernandez, H. 2005 Habitat structural

complexity and morphological diversity of fish assemblages in a Neo-tropical

floodplain river. Oecologia. 142, 284 - 295

Woodroffe, C. D. 1982 Geomorphology and development of mangrove swamps, Grand

Caymen Island, West Indies. Bulletin of Marine Science. 32, 381 – 398

Worrall, J. J., Anagnost, S. E., and Zabel, R. A. 1997 Comparison of wood decay among

diverse lignicolous fungi. Mycologia. 89, 199 - 219

Wright, J. P., Jones, C. G., and Flecker, A. S. 2002 An ecosystem engineer, the beaver,

increases species richness at the landscape scale. Oecologia. 132, 96 – 101

Wright, J. P., and Flecker, A. S. 2004 Deforesting the riverscape: the effects of wood on fish

diversity in a Venezuelan piedmont stream. Biological Conservation. 120, 439 – 447

250

Xavier, J. H. D. A., Cordeiro, C. A. M. M., Tenorio, G. D., Diniz, A. D. F., Junior, E. P. N.

P., Rosa, R. S., and Rosa, I. L. 2012 Fish assemblages of the Mamanguape

Environmental Protection Area, NE Brazil: abundance, composition and

microhabitat availability along the mangrove-reef gradient. Neotropical Ichthyology.

10, 109 - 122

Yulianto, E., Sukapti, W. S., Rahardjo, A. T., Noeradi, D., Siregar, D. A., Suparan, P., and

Hirakawa, K. 2004 Mangrove shoreline responses to Holocene environmental

change, Makassar Strait, Indonesia. Review of Palaeobotany and Palynology. 131,

251 – 268

Yusef, A. A., and Francisco, H. 2009 Climate Change Vulnerability Mapping for Southeast

Asia. Economy and Environment Program for Southeast Asia (EEPSEA). Singapore