AN EMPIRICAL EVALUATION of Posidonia Australis (R

AN EMPIRICAL EVALUATION OF Posidonia australis (R. Br.) Hook f. RESTORATION IN WESTERN AUSTRALIA: DEVELOPMENT OF A DECISION-BASED RESTORATION FRAMEWORK

Marnie Lyn Campbell

This thesis is presented for the degree of Doctor of Philosophy School of Biological and Environmental Science Murdoch University, Western Australia. 2000

Declaration

This dissertation is my own account of research I carried out from 1993 to 1999 and has not previously been submitted for the award of any degree at a tertiary institution.

Marnie Lyn Campbell

ACKNOWLEDGEMENTS

There are a number of people and organisations that I would like to thank.

Firstly, I wish to thank my supervisor, Dr Eric Paling, whose guidance has made this thesis what it is today. I would also like to thank my funding company Cockburn Cement Ltd, who made this project possible through a generous scholarship and the provision of valuable resources. At Cockburn Cement Ltd, I’m particularly indebted to Mr Roger Wilson, Mr Richard Peters, Mr Pieter Tencate, Eric, Alan and all the guys at the Woodman Point plant who helped to maintain the laboratory.

I also wish to thank the Great Barrier Reef Marine Park Authority for a grant, which was awarded to me in 1993. To the Royal Australian Navy Clearance Diving Team IV, I extend my thanks for the seagrass transplanting work they gladly helped by providing man and vessel power.

Thanks to all my dive buddies, Ian, Troy, Jannette, Helen, Blair, Malcolm, Richard and Miriam, who’ve always been there no matter what the weather was like. To Troy Sinclair, special thanks for keeping the mesocosms going when I was in Tasmania. I am extremely grateful to Dr Chad Hewitt (CSIRO, CRIMP) and Dr Louise Goggin (CSIRO, CRIMP) for providing useful criticism of my dissertation.

Last but not least, my sincerest thanks to all the people who have kept me sane: The “Campbell’s” (Jenny, Doug, Jodi, Keong, Tiggy, Nathan, Bret, Julie, Scott, Jaimi and Kate), Jannette Nowell, Helen Astill, Kim Benjamin, Troy Sinclair and the CSIRO CRIMP team. A special thanks to Ian Nelson, for all the constant encouragement, helpful suggestions, tireless help in the field, and most importantly for laughing when I needed it most.

ii ABSTRACT

The loss of biodiversity is currently recognised as one of the greatest threats to continued ecosystem function. In terrestrial habitats this has been well researched and publicised resulting in active restoration and mitigation efforts. However, in marine environments the current efforts are less effective. Seagrasses are widely recognised as fundamental species in forming the basis of marine trophic food webs, binding sediments, providing habitat structure, shelter, and nurseries for fish and crustaceans and increasing the activity and movement of several active molecules and nutrients. Despite this crucial role, seagrass losses world-wide continue due to land reclamation, building of marinas and port facilities, eutrophication due to rural and urban runoff, inshore dumping of pollutants and dredging. While nations have legislated seagrass mitigation, no effective means of establishing new seagrass or restoring damaged meadows exist at present.

This dissertation examines current efforts world-wide to elucidate a common framework for identifying the crucial elements of a restoration plan, which include site selection, transplant unit and technique and habitat enhancement. Posidonia australis was identified as one of the dominant meadow forming species in Western Australia and therefore was selected to investigate the utility of this framework. An empirical examination of site selection was undertaken to determine a potential transplant site for Posidonia australis. Critical factors examined were light requirements, burial and handling disturbance and substrate preference. Based upon this evidence, Success Bank was found to be optimal, with high light levels (> 5% surface irradiance), fair water quality, no burial period, low- mid water movement and a sand substrate. Sites at Carnac Island and Woodman Point were rejected because they did not meet these fundamental criteria.

Transplant unit and technique were evaluated for Posidonia australis. This species produces large numbers of seed that have a high viability (91%) but few seedlings actually establish (< 3%). During the course of this project, natural vegetative recruitment was observed in the field with 31% of natural vegetative propagules settling and growing (0.78 mm d-1). Field rhizomes were also observed to extend at rates of 1.04 mm d-1. Based upon these findings P. australis vegetative propagules (plugs) were selected as the most appropriate transplant unit.

Habitat enhancement techniques are an optional component of a restoration activity and may significantly increase transplant success. In order to reduce water movement at the selected transplant site, the use of artificial seagrass mats was experimentally evaluated. Artificial seagrass mats were found to increase plug survival and rhizome elongation. In addition artificial seagrass mats reduced the variability in accretion and erosion of sediments. In the presence of habitat enhancement, up to 50% of seagrass plugs survived, with 39% exhibiting rhizome extension. Based on these findings a decision-based framework for seagrass restoration is presented with a discussion of future applications.

iii TABLE OF CONTENTS

Chapter Page

Declaration i Acknowledgments ii Abstract iii Table of Contents iv List of Figures vi List of Tables viii

Preface Biodiversity and human impacts on seagrass communities 1

Chapter 1 A review of seagrass restoration efforts and the development of a 5 restoration framework 1.1 The marine environment – seagrasses 5 1.2 Mitigation and mitigation banking 8 1.2.1 Mitigation 8 1.2.2 Mitigation banking 9 1.3 Seagrass transplantation: examples 11 1.3.1 Worldwide examples 11 1.4 Why do seagrass transplants fail? 15 1.5 Stage I: site selection 17 1.6 Stage II: transplant unit 18 1.7 Stage III: habitat enhancement methodology 20 1.7.1 Anchors 20 1.7.2 Artificial barriers 21 1.7.3 Mesh 21 1.7.4 Artificial seagrass mats 22 1.8 Conclusions and Aims 24

Chapter 2 Optimising the process of site selection for Posidonia australis 26 transplantation in Western Australia 2.1 Introduction 26 2.1.1 Light levels 26 2.1.2 Disturbance: erosion and accretion 28 2.1.3 Substrate type 29 2.1.4 Optimal and sub-optimal sites 30 2.1.5 Aims 31 2.2 Materials and methods 32 2.2.1 Mesocosm establishment 32 2.2.2 Fruit and rhizome collection 32 2.2.3 Pilot studies 36 2.2.4 Manipulation experiments 37 2.2.5 Statistical analyses 40 2.3 Results 40 2.3.1 Pilot studies 40 2.3.2 Manipulation experiments 40 2.4 Discussion 47 2.4.1 Pilot studies 48 2.4.2 Manipulation experiments 48 2.4.3 Variation between sites 54 2.4.4 Mesocosms as an experimental tool 55 2.4.5 Conclusions 56

Chapter 3 Posidonia australis propagation: selection of transplant unit and technique 58 for restoration 3.1 Introduction 58 3.1.1 Seagrass reproductive strategies 58 3.1.2 Dispersal mechanisms of seagrasses 60

iv 3.1.3 Life history strategies 61 3.1.4 Posidonia colonisation and meadow expansion 62 3.1.5 Aims 63 3.2 Materials and methods 64 3.2.1 Site descriptions 64 3.2.2 Meadow measurements 69 3.2.3 Seed viability and fruit longevity 70 3.2.4 Seedling growth 72 3.2.5 Rhizome elongation over depth and between species on Success Bank 72 3.2.6 Vegetative propagule recruitment 72 3.2.7 Statistical analyses 72 3.3 Results 73 3.3.1 Sexual reproduction 73 3.3.2 Vegetative reproduction 81 3.4 Discussion 82 3.4.1 Sexual reproduction 82 3.4.2 Vegetative reproduction 88 3.4.3 Meadow maintenance and choice of transplant unit 92

Chapter 4 Evaluating vegetative transplant success in Posidonia australis: a field trial 94 with habitat enhancement 4.1 Introduction 94 4.1.1 Restoration and mitigation 94 4.1.2 Aims 99 4.2 Materials and methods 99 4.2.1 Site selection 99 4.2.2 Habitat enhancement 100 4.2.3 Seagrass transplantation 100 4.2.4 Statistical analyses 102 4.3 Results 103 4.3.1 Habitat enhancement 103 4.3.2 Seagrass transplantation 109 4.4 Discussion 112 4.4.1 Habitat enhancement 113 4.4.2 Seagrass transplantation 115 4.4.3 Conclusions: success of enhancement and transplantation 119

Chapter 5 General discussion: the development of a decision based framework for 122 restoration planning 5.1 Introduction 122 5.2 Decision-based framework for restoration planning 125 5.3 Summary of findings 129 5.4 Future considerations 132

References 136

Appendix A 162 Appendix B 168

v LIST OF FIGURES

Figure Page

1.1 IUCN world bioregions from Kelleher et al (1995) (1 Antarctic; 2 Arctic; 3 7 Mediterranean; 4 North West Atlantic; 5 North East Atlantic; 6 Baltic; 7 Wider Caribbean; 8 West Africa; 9 South Atlantic; 10 Central Indian Ocean; 11 Arabian Sea; 12 East Africa; 13 East Asian Seas; 14 South Pacific; 15 North East Pacific; 16 North West Atlantic; 17 South East Pacific; 18 Australia/New Zealand) and regions (shaded grey) where published seagrass restoration efforts have occurred (based on Appendix A). 1.2 A framework developed to aid seagrass transplantation 17 2.1 Mesocosm laboratory at Woodman Point, Western Australia. 32 2.2 A tagged Posidonia australis rhizome. A cable tie marks the last shoot on the rhizome 34 prior to the apical meristem. 2.3 Sand gravity filters used to stop excessive silt and biological debris from entering the 35 mesocosms. 2.4 Ozomatic® ozone generator used to sterilise the water before it entered the 36 mesocosms. 2.5 Rhizome growth measured as a) increase in rhizome length beyond the tagged region; 37 b) necrosis encroaching past the tagged region resulting in a decrease in growth and; c) necrosis at the apical end of the rhizome resulting in a decrease in growth. Necrosis is shown by the shaded ! region. 2.6 Posidonia australis mean net rhizome growth (± SE) in free flowing seawater 42 mesocosms subjected to reduced irradiance: a) 25% SI treatment, b) 5% SI treatment. Symbols are represented by: treatment ()) and control ("); a plot of in situ rhizome ---&--- growth is included for comparison. Rhizomes were collected from two sites, Wreck Rock (closed !) and Success Bank (open "), whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2). 2.7 Posidonia australis mean net rhizome growth (± SE) in free flowing seawater 44 mesocosms subjected to: a) cyclical burial disturbance (days 0–21 and 43-63 are where rhizomes were buried, whilst days 22-42 is where rhizomes are uncovered); and b) complete burial disturbance, where shaded symbols represent Wreck Rock and open symbols represent Success bank. Symbols represented treatment ()) and control ("); a plot of in situ rhizome ---&--- growth is included for comparison. Rhizomes were collected from two sites, Wreck Rock and Success Bank, whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2). 2.8 The effect of handling disturbance on Posidonia australis mean net rhizome growth 46 (± S.E.) in free flowing seawater mesocosms. Handling disturbance treatments: frequent handling ); infrequent handling "; and no handling 4. A plot of in situ rhizome growth ---&--- is included for a comparison of field and mesocosm rhizome growth. Mesocosm rhizomes were collected from two sites, Wreck Rock (closed!) and Success Bank (open"), Western Australia, whilst in situ rhizomes were measured on Success Bank (see Chapter 3, Figure 3.2). 2.9 Posidonia australis mean rhizome growth (± SE) in free flowing seawater mesocosms 47 on limestone rubble ) and a control substrate ". A plot of in situ rhizome ---&--- growth is included for a comparison between treatments and field growth. The rhizomes were collected from two sites, Wreck Rock (closed!) and Success Bank (open"), whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2). 3.1 The Australian distribution of Posidonia australis. 62 3.2 Site locations for the study of Posidonia seed viability, longevity, shoot density and 66 biomass, fruit density, seedling density and rhizome expansion in Western Australia. Site 1) Carnac Island; site 2) Success Bank; site 3) Woodman Point; site 4) Wreck Rock; site 5) Peak Island; site 6) Cosy Corner and; site 7) Cockburn Sound. Meadow fruit and seedling densities and biomass was measured at all sites. Fruit viability, longevity, and seedling growth measurements at sites 2, 3 and 7.

vi 3.3 Transect A and B orientation and site placement on Success Bank (refer to Figure 3.2 67 for site locations). Numbers in brackets refer to depth. Rhizome extension was measured on transect A. Grey areas represent vegetative propagule recruitment transects (not drawn to scale). 3.4 Depth of the Success Bank meadow sites (1-5) on transect B (refer to Figure 3.2 for 67 location). 3.5 Twenty four-hour temperature trial to determine optimum reactivity of 1% 71 tetrazolium with Posidonia australis seeds. Seeds were kept at 18°C, 25°C or 30°C. Seeds were derived from fruits collected at Woodman Point, Success Bank, Wreck Rock and Cockburn Sound (see Figure 3.2). Viable tissue was determined by the distribution of stained live tissue as described by Lakon (1948). 3.6 Correlation between fruit density and seedling density, in the seagrass Posidonia 77 australis (r2 = 0.69, p > 0.05). 3.7 Posidonia australis seed viability of a) fresh fruits at Woodman Point !, Wreck 79 Rock !and Success Bank !, and b) wrackline fruits at Woodman Point ! and Cockburn Sound ! (see Figure 3.2). Time elapsed for wrackline fruits began from when the fruit was collected from wrack, not when it was released from the plant. 4.1 Posidonia australis transplant site ( situated on Success Bank, Western Australia. 96 4.2a The amount of substrate erosion and accretion occurring before in ASG mats (() and 104 non-ASG areas ()) at site 1 on the recipient transect, Success Bank in Western Australia. 4.2b The amount of substrate erosion and accretion occurring before in ASG mats (() and 105 non-ASG areas ()) at site 2 on the recipient transect, Success Bank in Western Australia. 4.2c The amount of substrate erosion and accretion occurring before in ASG mats (() and 106 non-ASG areas ()) at site 3 on the recipient transect, Success Bank in Western Australia. 4.2d The amount of substrate erosion and accretion occurring before in ASG mats (() and 107 non-ASG areas ()) at site 4 on the recipient transect, Success Bank in Western Australia. 4.2e The amount of substrate erosion and accretion occurring before in ASG mats (() and 108 non-ASG areas ()) at site 5 on the recipient transect, Success Bank in Western Australia. 4.3 Percentage survival of Posidonia australis seagrass transplants in ASG mats (() and 110 non-ASG areas (") on the 5 sites on Success Bank, Western Australia: a) site 1, b) site 2, c) site 3, d) site 4, and e) site 5. 4.4 Posidonia australis transplant leaf growth on Success Bank in Western Australia at: 111 a) site 1 ()), site 2 ("), site 3 (h), site 4 (’) and site 5 (v); and b) during October 1994 ()), December 1994 ("), February 1995 (h), April 1995 (’), June 1995 (J), August 1995 (() and October 1995 (!). 4.5 Primary leaf production of Posidonia australis seagrass transplants on Success Bank 112 in Western Australia at: a) site 1 ()), site 2 ("), site 3 (h), site 4 (’) and site 5 (v); and b) during May 1994 ()), June 1994 ("), August 1994 (h), October 1994 (’), January 1995 (J), March 1994 ((), April 1995 (!), June 1995 (f), August 1995 (&) and October 1995 (q). 5.1 Broad decision based framework for restoration planning. 126 5.2 Step 1, objective setting, in the planning process. 127 5.3 Step 2, site selection (Stage I), in the planning process. 130 5.4 Step 3, transplant unit and technique selection (Stage II), in the planning process. 131

vii LIST OF TABLES

Table Page

1.1 Success (%) of transplant attempts reported within the literature with sufficient 16 information to determine environmental stability and transplant success (identified (*) in Appendix A) using sexual or vegetative propagules. Three comparisons were made: a) Posidonia transplants in Western Australia; b) Posidonia transplants from around the globe and; c) all seagrass species on a global scale. -2 -1 -2 -1 2.1 Light saturation (Isat μmol m s ), light compensation (Ic μmol m s ) levels and life 27 history strategies (r- colonisers and K- climax species) for seagrass species. Data was extracted from the following literature; 1) Masini et al. 1990; 2) Beer and Waisel 1982; 3) Bulthuis 1983; 4) Dennison and Alberte 1985 and; 5) Drew 1979. 2.2 Rhizome growth (mm d-1) for each replicated (x3) experiment and the mean of all 41 three replicates (± SE) under conditions of frequent disturbance, infrequent disturbance, no disturbance, burial, 25% SI, 5% SI, a limestone rubble substrate and a sand substrate. 2.3 The mean rhizome growth rates for treatments of frequently, infrequently and not 43 handled. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 2, Figure 2.2). A negative growth rate indicates necrosis. 2.4 The mean net rhizome growth rates (mm d-1 ± SE) in complete burial disturbance and 43 control treatments. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). ). Negative net growth rates indicate necrosis exceeded growth. 2.5 The mean net rhizome growth rates (mm d-1 ± SE) for treatments of frequently, 45 infrequently and not handled (control). Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth. 2.6 The mean net rhizome growth rates (mm d-1 ± SE) on limestone rubble and sand 47 (control) substrates. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth. 2.7 Evaluation of Carnac Island, Woodman Point and Success Bank by the Stage I 57 Criteria listed in the text. ! denotes preferred choice. 3.1 Field measurements taken at the seven sampling sites. Site 1) Carnac Island; site 2) 65 Success Bank; site 3) Woodman Point; site 4) Wreck Rock; site 5) Peak Island; site 6) Cosy Corner; and site 7) Cockburn Sound. 3.2 Site, depth (m) and seagrass species examined on Success Bank. The seagrass species 68 are: Pa = Posidonia australis; Pc = Posidonia coriacea; Sf = Syringodium filiforme; Ho = Halophila ovalis; Ht = Heterozostera tasmanica and; Ag = Amphibolis griffithii. Grey denotes rhizome recruits. 3.3a Mean shoot densities of a range of seagrasses at seven sites in Western Australia (see 74 Figure 3.2). Sites are as in Table 3.1. Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera. 3.3b Student-Newman-Keuls multiple comparison results illustrating significant 74 differences (S) and no significant differences (NS) between shoot density of different species. 3.4a Mean aboveground biomass of a range of seagrasses at seven sites in Western 75 Australia (see Figure 3.2). Sites are as in Table 3.1. Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera. 3.4b Mean belowground biomass of a range of seagrasses at seven sites in Western 75 Australia (see Figure 3.2, Table 3.1). Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera.

viii 3.5 Above and belowground biomass testing the hypothesis that above or belowground 76 biomass was similar between sites, for seagrass species from sites in Western Australia (see Figure 3.2, Table 3.1). 3.6 Average Posidonia australis densities (number m-2) of fruits, seedlings, floating fruits 76 and flower spikes at sites in Western Australia (see Figure 3.2, Table 3.1); n = 30 for all sample sites. 3.7 Student-Newman-Keuls multiple comparison test results, indicating significant 77 differences (S) and no significant (NS) differences between seagrass shoot densities at different sites in Western Australia (see Figure 3.2; Table 3.1). 3.8 Percentage growth of Posidonia seedlings collected from fresh meadow fruits (Wreck 80 Rock, Success Bank and Woodman Point) and wrackline fruits (combined Woodman Point and Cockburn Sound). Growth is divided into three phases (1) emergence of a cotyledon, (2) production of roots and emergent leaves and (3) lengthening of roots. Time is measured in days. 3.9 In situ mean (± SE) rhizome growth rate (mm d-1), from fastest too slowest, for six 81 species of seagrass at three sites on Success Bank, with increasing depth. Site 1 is 4.9 m deep, site 2: 5.8 m, and site 3 is 8 .6 m depth (see Figure 3.4). NA – The species were not present at these sites. 3.10 Four types of seed banks described by Baskin and Baskin (1998). 86

ix Preface Biodiversity and human impacts on seagrass communities

In his book, Essay on the principles of population (cited in Krebs 1985; Tietenberg 1992), Thomas Malthus warned that a time would come when world population growth would be so great that food supply would not sustain the burgeoning population, resulting in an inevitable crash of human society through starvation and death. What he didn’t foresee was that the mechanism for this downfall would not come through the urge to reproduce (although this has played a large part), but through our other associated activities.

The key threats to ecosystem sustainability include: human overpopulation; excessive consumption and inequitable distribution of resources; institutions that diminish biological diversity; insufficient understanding of nature, primarily biodiversity; the failure of economic systems and policies to value the environment; and the failure of conservation measures outside of conservation parks and reserves (Norse 1993; State of the Environment Advisory Council 1996). Numerous impediments to overcoming these threats include insufficient scientific data, poor transfer of information and inadequate knowledge of cultural and biological diversity. To overcome these we need to strengthen our knowledge base; efficiently protect, manage and restore marine and terrestrial ecosystems; provide environmentally sound financial and technical assistance; and involve citizens in decision making. In so doing, human impact will be reduced and managed in a better, ‘greener’ fashion. It would be an indictment of society if the statement by Senegalese ecologist Baba Dioum proves to be true:

“In the end we will conserve only what we love; we will love only what we understand; and we will understand only what we are taught.” (cited in Norse 1993).

The world’s population is increasing exponentially and by the end of next century it should stabilise at around 12 billion people, twice the current (1998) number (Tietenberg 1992; Pearce and Moran 1998). To date the human population and their domesticated livestock (flora & fauna) have spread from small isolated populations to cover most of the earth’s surface. This has been achieved in only 1100 years (Crosby 1986). More arable land is needed to feed this growing population as well as land for urbanisation and communications infrastructure (Tietenberg 1992; Pearce and Moran 1998; Ruckelshaus and Hays 1998). Typically, this land is gained by destroying rainforests, clearing woodlands and savannas, and reclaiming land from both the sea and wetlands (Crosby 1986; Bürger and Lynch 1997; Dobson et al. 1997; Vitousek et al. 1997; Pearce and Moran 1998). The consequences are net loss of arable land through land degradation, increased salinity, topsoil depletion, water pollution and poisoning of the soil with chemicals, which simultaneously act to reduce biological diversity. Presently, it is estimated that 39% of terrestrial primary production is consumed, diverted or destroyed (Norse 1993).

1 Not only have we spread at great rates but our activities have depleted renewable and non-renewable resources, created pollution, caused the extinction of plant and animal species, depleted the ozone layer and caused climate change (Crosby 1986; Tietenberg 1992; McMichael 1993; State of the Environment Advisory Council 1996; Pearce and Moran 1998). This destruction has been heeded with the establishment of many conservation organisations around the world. Such organisations include: the IUCN – World Conservation Union, founded in 1948; the World Wildlife Fund; the United Nations Environment Programme, created in 1972; and the World Bank (which became environmentally conscious in 1987 and began to implement environmental assessment of its projects in 1989). The general aim of these organisations is to help stop the depletion of the world’s fauna and flora and hence protect the diversity of terrestrial and marine environments. In 1989, a large group of international organisations began the difficult task of developing a Global Diversity Strategy (WRI et al. 1992), with the objective of saving life on earth. Similarly in 1996, Australia adopted The National Strategy for the Conservation of Australia’s Biological Diversity with the aim “to protect biological diversity and maintain ecological processes and systems” (Hopper 1998).

Four important reasons exist for implementing a system to maintain Australia’s biodiversity. First, healthy ecosystems are necessary for maintaining life on earth. Second, there is the ethical issue of maintaining resources for our children and the generations to come. Third, the aesthetic and cultural qualities of biodiversity are value adding to the economy. Finally, Australia’s flora and fauna represent a resource that is used in tourism, medicine, energy and building industries (State of the Environment Advisory Council 1996). A range of economic instruments are being developed in Australia, to help conserve biodiversity. They include: 1) Environmental pricing: Funding the conservation of biodiversity through the setting of charges, levies and prices. Examples include park and reserve fees and environmental levies used by councils to buy and protect environmentally sensitive habitats. There is a potential to use environmental levies on industries and businesses that diminish biodiversity, however this is underutilised in Australia. To date, environmental pricing is rare in Australia (State of the Environment Advisory Council 1996); 2) Conservation easements: These bind owners to a set of conditions that maintain diversity. For example some housing developments restrict land clearance or cropping. Usually these acts come under heritage agreements, but they may also be covered through fire protection agreements and hilltop reserves. 3) Funding arrangements: This is a form of mitigation banking that involves the purchase of land and placement of permanent covenants on it to manage part or all of it for conservation. If the land is sold the buyer must abide by the covenant and the money used to buy the land is used to purchase more land for conservation. This method is used by the Victorian Conservation Trust, in the form of a revolving fund; 4) Taxation: Income-tax deductions exist for control of land degradation but are not concerned with conservation of biodiversity. The potential for a ‘green’ tax exists, which as yet has not been implemented.

2 5) Transferable development rights: This mechanism limits development in conservation areas without affecting the underlying value of individual assets. In Australia this mechanism has not been implemented however it has proved effective in New Zealand. 6) Performance bonds: Environmental performance bonds are best suited to situations where the potential environmental damage can be estimated. Such bonds exist for the mining and aquaculture industries. 7) Financial assistance: This usually takes the form of payment to assist with the associated costs purchasing material associated with the required conservation work. Often it is helpful for community action groups. Unfortunately the majority of these instruments are ineffective and ill designed to manage biodiversity. Instead they act to compensate for losses and fund research into sustainable usage of resources. Therefore, they are rarely effective mechanisms to stop the continuing loss of biodiversity.

The best way to maintain biological diversity is to alter or stop the current loss and to activate manipulation techniques such as restoration and mitigation (Norse 1993). Presently, terrestrial and freshwater habitats are considered easier to restore and hence have been effectively manipulated for many years (Norse 1993). Terrestrial community restoration largely relies on stabilising the site, which initiates soil development and facilitates re-establishment by native vegetation (Wyant et al. 1995). ALCOA, a company that mines bauxite in the Darling Ranges, Western Australia, initiates mine site restoration before mining begins by removing a thin (50 mm) surface layer of topsoil, which is stockpiled or transferred directly to another mine pit awaiting restoration (Nichols et al. 1985). Hence, the topsoil is replaced after mining allowing the seeds in the topsoil seedbank to germinate, using the soil’s nutrients to grow. ALCOA has used this technique for many years with moderate success (Nichols et al. 1985).

Community groups are also actively involved in restoration/mitigation work. Such groups often help reestablish areas that have been cleared of native vegetation. Examples of such restoration work are evident in many Australian rural areas where both Landcare and Greening Australia projects have been successful in planting many trees, with the specific aim to control erosion and soil salinisation. The side effect of these projects is that native habitats are re-established. Integration of landscape ecology and bioregional planning is now being practiced by many community-based groups and has seen the development of projects like Greening Australia’s ‘Corridors of Green’ (State of the Environment Advisory Council 1996). Both community projects and mine rehabilitation are helping to correct the process and consequences of habitat fragmentation, which affect terrestrial habitats.

Coastal environments have had limited and often localised success with restoration efforts. Typical restoration areas include coastal heathlands, beach dunes, salt marshes and mangroves. Local councils are involved in dune restoration work, which involves the education of beach users, the erection of fences and erosion walls, and the replanting of native vegetation on denuded dunes and in beach swales. Woodside, a petroleum mining company in Western Australia, undertakes extensive monitoring of mangroves near their on-shore facilities at

3 Dampier. The iron ore mining company, BHP, also monitors and funds research into mangroves near their loading facilities in Port Hedland, Western Australia (M. Piggott. pers. comm.). Both companies monitor mangrove stands to determine damage from their associated activities, however as yet, they are not required to restore damaged mangrove areas. Mangroves have also been cleared to provide land in urban regions such as Sydney, Newcastle, Brisbane, Cairns and Adelaide. In these situations no restoration work has occurred and natural recruitment is hampered by increased sedimentation due to upland clearing or coastal modification.

In aquatic environments, the majority of restoration work occurs in wetlands. Mining companies such as Capel Sands in Western Australia are now responsible for restoring wetlands damaged by their actions. Much of this work is voluntary and methods of monitoring success do not exist. In the U.S., where restoration research is more advanced, wetlands are restored with varying success. For example, wetlands were damaged in San Diego during construction of a freeway and flood control channel. Compensation for these wetlands included efforts to reestablish the endangered plant Cordylanthus maritimus maritimus and separately to create habitat for the endangered bird, Rallus longirostris levipes (Zedler 1998). Although mitigation criteria for C. maritimus maritimus was met, complicating factors (such as lack of pollinators and insufficient winter rainfall) may still induce local extinction of this species (Zedler 1998). In contrast, the created habitat for R. longirostris levipes has been unsuccessful (Zedler 1998). An optimistic goal of restoring 10 million acres of wetlands, 400, 000 miles of riparian ecosystems and 2 million acres of lakes in the U.S. within the next 20 years has been set by the National Research Council (Zedler 1998).

Biodiversity in the marine environment is potentially as high as in the terrestrial environment yet it is poorly known. Marine biodiversity influences commercial and recreational fishery values, recreational activities and aesthetic qualities. Marine restoration and biodiversity evaluation lag behind terrestrial research and are considered to be relatively ‘new’ fields of science. Yet, if equivalency between degraded and restored habitat is to be achieved by marine restoration then biodiversity needs to be thoroughly evaluated and understood. As yet mitigation in marine ecosystems have had few trials and fewer successes.

Chapter 1 A review of seagrass restoration efforts and the development of a restoration framework

1.1 The marine environment – seagrasses The majority of the world’s population lives in coastal regions. For example, in southwest California 16 million people live on the narrow coastal strip representing 8% of the State’s total area (Davis et al. 1995). Similarly, in Australia the majority of the population lives in the capital cities located on the southeast coast. The increasing population density coupled with the propensity for humans to live in coastal regions has placed enormous pressure on coastal ecosystems (Norse 1993; Short and Wyllie-Echeverria 1996; Ruckelshaus and Hays 1998). Marine impacts are not as visible as terrestrial destruction, hence it has only been in recent decades that impacts in the marine environment are being reported, documented and acted upon. Activities that impact upon marine systems include the building of marinas and port facilities, eutrophication due to rural and urban runoff, inshore and offshore dumping of pollutants, over-utilisation of fish stocks, dredging and dredge spoil dumping, reclamation, benthic trawl fishing and the introduction of exotic marine species (e.g. Lubchenco et al. 1991; Norse 1993; State of the Environment Advisory Council 1996).

From an ecological perspective, one of the most detrimental consequences of human impacts in the marine ecosystem is the loss of primary producers, especially seagrasses. Seagrasses are vigorous marine plants (Thayer et al. 1975; Phillips 1982, 1990a; Walker et al. 1991) that form the basis of marine trophic food webs providing numerous ecosystem functions. Rhizomes of established seagrass communities form thick mats that bind sediments. The leaves of seagrasses act as a baffle that slows water movement allowing sediment to come out of suspension and settle, binding sediment to the rhizome mat and reducing erosion (Raupach and Thom 1981; Fonseca and Fischer 1986; Fonseca and Kenworthy 1987; Short 1987). Leaves of seagrasses provide substrates for flora and fauna epiphytes, which provide food for fauna in the meadows and release nutrients that enter the food chain, enhancing both floral and faunal abundance (Kikuchi 1980; Zieman and Wetzel 1980; Scott et al. 1986). Seagrass leaves, stems and rhizomes shelter and protect juvenile fish and crustaceans (Summerson and Peterson 1984; Zimmerman et al. 1991), serving a crucial function as nursery habitat and leading to enhancement of commercial and recreational fishing industries. Seagrasses increase the activity and movement of several active molecules and nutrients: oxygen is produced; sulfur enters the sulfur cycle from seagrass detritus; and nitrogen and phosphorous are returned to the water column through detrital cycling (Carigan and Kalff 1980; Zieman and Wetzel 1980; Phillips 1982, 1990a; Zimmerman et al. 1991). These various functions highlight the irreplaceable nature of seagrasses.

Seagrasses are a geologically old taxonomic group that diverged from the angiosperms in the Cretaceous Period, approximately 145 - 65 myBP (Larkum and den Hartog 1989). Speciation occurred relatively early in their evolution, with fossil evidence indicating that relatively few new species arose after the Eocene Period

(Larkum and den Hartog 1989). Consequently, this provided evidence that seagrasses have remained unchanged for a long period of time and are considered relict species (Larkum and den Hartog 1989). This is supported by a lack of speciation when compared to terrestrial angiosperms: there are 55 species of seagrasses as opposed to ~ 235,000 species of terrestrial angiosperms (Raven et al 1986; Larkum and den Hartog 1989).

The fossil record also provides evidence for limited seagrass speciation. Seagrasses were originally restricted to shallow nearshore environments, possibly because they were wind pollinated (anemophily). Due to the distribution of genera it is probable that high speciation occurred during the period from the Cretaceous to the Eocene (145 to approximately 55 myBP) (Larkum and den Hartog 1989). Subsequent evolution of water pollination (hydrophily) led to expansion into subtidal environments. Larkum and den Hartog (1989) suggest that this shift to water pollination may have also been a basis for a slower rate of speciation due to reduced pollen dispersal distances and lower competition in subtidal environments.

It is believed that seagrasses are well adapted to their environments and form stable communities, however observations indicate that large fluctuations in meadow area occur through time which indicate a fluctuating environment (e.g. salinity, temperature, light). Evidence of many paleo-extinctions both of species and populations leading to patchy geographical distributions suggest that seagrasses are only moderately fit for the environment in which they live and that they can be best categorised as stress tolerant plants (Larkum and den Hartog 1989).

Speciation and endemism is prominent in Australian seagrasses. One third of all seagrass species are endemic to Australia and have relict distributions. This persistence is thought to have evolved because of Australia’s long separation from other landmasses and the stable temperatures that have existed for much of this time. It is postulated by Larkum and den Hartog (1989) that stress factors, such as fluctuating temperatures, herbivory and competition were absent from Australia until relatively recently, leading to fewer paleo-extinctions. Hence, Australian seagrasses are considered successful because they have persisted for a long period of time, but they are poorly adapted to the present environmental conditions, which is clear by their displacement and lack of recovery after certain natural and artificial disturbances. The lack of successful recovery in many instances is a function of poor seagrass growth and reproduction.

The consequences of seagrass loss are well-established and includes community shifts to phytoplankton dominated systems which lack a three dimensional structure; decreases in benthic invertebrate, fish and crustacean numbers; and resuspension of sediments reducing available light (Thayer et al. 1975; Walker and McComb 1992). Seagrass losses are primarily due to eutrophication causing light attenuation, or outbreaks of wasting disease. Although the consequences are recognised, seagrass losses continue to occur, with numerous studies documenting the degree and extent of loss. In Australia alone, it is estimated that > 45,000 ha of seagrasses have disappeared, with few gains (Kirkman 1992, 1995; Walker and McComb 1992). Other

countries and regions such as the United States, Canada, the Mediterranean and the Caribbean have also lost vast amounts of seagrass meadows (Short and Wyllie-Echeverria 1996; Ruckelshaus and Hays 1998), yet estimates on the extent of these losses is not reported or unknown.

Scientists have been studying seagrass communities for many years yet information about the biological diversity (infaunal, epifaunal and epiphytic) in these communities is poor. It is unclear how much seagrass can be lost before this diversity is compromised. Given these observations it would be better to adopt an approach that prevents seagrass loss, damage or disturbance.

Continual loss of seagrasses has resulted in research progressively focussed on restoration and mitigation. Globally, in the past 35 years there have been at least 114 published efforts to restore and/or mitigate seagrass losses with numerous unpublished attempts. The published efforts have used at least 14 different species occurring in six IUCN bioregions (Australia and New Zealand; the wider Caribbean; the northeast Pacific; the northeast Atlantic; the northwest Atlantic; and the Mediterranean; Kelleher et al 1995) (Figure 1.1; see also Table 1.2, Appendix A). In all these efforts, the site, transplant unit and transplantation method have differed but the general reason for doing the work has been the same: to overcome habitat loss. Unfortunately, net loss of habitat has not been prevented through these efforts (Kirkman 1992; Short and Wyllie-Echeverria 1996; Williams and Davis 1996; Ruckelshaus and Hays 1998).

6 5 15 4 16 7 3 7 11 14 10 13 8 17 14 12 9 18

Figure 1.1: IUCN world bioregions from Kelleher et al (1995) (1 Antarctic; 2 Arctic; 3 Mediterranean; 4 Northwest Atlantic; 5 Northeast Atlantic; 6 Baltic; 7 Wider Caribbean; 8 West Africa; 9 South Atlantic; 10 Central Indian Ocean; 11 Arabian Sea; 12 East Africa; 13 East Asian Seas; 14 South Pacific; 15 Northeast Pacific; 16 Northwest Atlantic; 17 Southeast Pacific; 18 Australia/New Zealand) and regions (shaded grey) where published seagrass restoration efforts have occurred (based on Appendix A).

1.2 Mitigation and mitigation banking Mitigation and mitigation banking have both been used to counter seagrass loss. Seagrass mitigation is defined as making seagrass loss less severe; two options are to either restore an area or to create a new meadow. Seagrass restoration on the other hand, is defined here as re-establishing seagrass in an area where disturbance and/or loss of an existing meadow has occurred (Lewis 1987; Kirkman 1989). Seagrass restoration aims to return the area to its pre-existing composition, structure and function. Mitigation occurs either as a colonisation event (e.g. scattering sexual propagules within an area) or by artificially creating a “new” meadow that has been disturbed (e.g. using plugs). Creation of a meadow involves establishing seagrass at a site that has not been documented to have supported a meadow in the recent past (Kirkman 1989).

If sexual propagules are used, the plants must establish and spread in a similar fashion to a colonisation event with recruitment, establishment and vegetative growth/spread of propagules. The probability of success in a seagrass colonisation mitigation effort (using sexual propagules) in Western Australia is relatively low, as demonstrated by LeProvost Environmental Consultants (1990) and Kirkman (1992, 1995). Both these researchers recorded 0% survival of seagrass seedlings used in “colonisation” style transplantation efforts. The advantage to using sexual propagules is that it is low cost, large areas can be covered and genetic diversity is maintained.

With monoecious plants, vegetative propagules ensure that the transplant unit is rapidly capable of both vegetative and sexual reproduction to spread and restore the area (Bird et al. 1994). However, if the plant is dioecious then care must be taken to ensure that you have both male and female plants in the new population (Bird et al. 1994). Paling et al. (1997) has demonstrated that using large (0.5 m2) plugs of seagrass is very successful (75-97%), although costly. Special mention must also be made that using vegetative propagules to restore a meadow, does not necessarily maintain genetic diversity. Maintenance of genetic diversity is highly dependent on transplant units representing the entire donor population.

1.2.1 Mitigation Mitigation is one method used to maintain biological diversity when areas have been damaged or destroyed. The mitigation process involves three steps: according to Guerrant and Pavlik (1998) planning, implementation and evaluation. They state that planning considers the legal issues, project context, resource management policies and existing scientific literature, leading to the development and documentation of a mitigation plan. Each mitigation plan is unique and considers aspects such as the site, transplant unit, number of transplants, transplant technique and who is responsible for implementing the work. Once this is complete, implementation of the plan can occur. Evaluation involves goal and objective setting (defining success and failure), setting performance indicators of success or failure and monitoring the outcomes. Intrinsic to this

process is objective setting: will mitigation replace the damaged (or to be damaged) community with a similar community or can other more easily restored or created communities be used? Generally, it is considered that once the objective is determined, long-term monitoring (a minimum of five years) is needed to document the outcomes (Fonseca et al. 1996; Ruckelshaus and Hays 1998). To determine success, a strict definition that encompasses the predetermined objectives of the mitigation effort must be used and all parties involved must be familiar with this definition.

Mitigation in the United States is commonly defined as the replacement of a lost, disturbed or damaged population or ecosystem at locations where they will be safe from development (Nitsos 1988; Harrison 1990; Norse 1993). Generally, it deals with wetlands and terrestrial communities (e.g. USNRC 1992; Norse 1993; National Research Council 1995; Guerrant and Pavlik 1998; Maunder et al. 1998; Zedler 1998). United States mitigation legislation for seagrass loss is under the shared jurisdiction of the Army Corps of Engineers (COE), the National Marine Fisheries Service (NMFS), the U.S. Fish and Wildlife Service (administering the Endangered Species Act), Stae and Federal Environmental Protection Agencies (EPAs) (administering the Coastal Zone Management Act), the U.S. Clean Water Act and State-level Natural Resource Agencies (USNRC 1992; Williams and Davis 1996; Nitsos 1988; Ruckelshaus and Hays 1998). The COE has oversight on the majority of restoration efforts (Ruckelshaus and Hays 1998).

Determining success of a mitigation effort is often difficult, involving long term planning, monitoring of affected marine areas and record keeping of such efforts. Success as defined above, is often impossible because many areas are mitigated as an afterthought, leaving no opportunity to record pre-disturbance community level information. The COE has limited funding dedicated to post-mitigation monitoring and evaluation of project success or maintaining information on the acreage of affected habitat and mitigation permits issued. Wetlands are an exception and have been heavily monitored and reported upon due to independent legislation and international treaties such as the RAMSAR convention (Ruckelshaus and Hays 1998).

1.2.2 Mitigation Banking Mitigation banking is now used in the United States (USNRC 1992) and can be accomplished in a number of ways, three of which are: 1) The body that is causing the damage (for example a mining company) buys an equivalent area of land (e.g. pastureland) that they restore to mitigate for the area (aquatic or terrestrial) to be damaged or already damaged; 2) The body that is causing the damage finances a mitigation effort to be performed by a third party such as a consultant, a state or federal agency or an non-government organisation (NGO); 3) The body pays an amount of money equivalent to the cost of performing a mitigation effort to an environmental organisation (generally an NGO). This organisation may use the funds either to buy non-

impacted land to be preserved, to fund other preservation campaigns, or to perform a restoration activity, none of which necessarily have to do with the habitat impacted by the actions of the body. All three methods ensure that efforts are made to alleviate the loss of a habitat when an area is damaged or destroyed. However, an equivalent area and/or community is not always replaced. The first two imply some valuation of impact mitigation through 1:1 replacement. The third allows the industry to abrogate responsibility to a third party (NGO) which may act in an entirely different habitat (e.g substitute terrestrial mitigation for marine destruction).

Many presume that marine mitigation offers little chance of success and hence mitigation banking is seen as the only viable option. This assumption is based on the lack of success in wetland restoration projects (Ruckleshaus and Hays 1998; Zedler 1998). In recent times, this lack of confidence in marine mitigation has caused NMFS to stop accepting seagrass mitigation projects in exchange for habitat destruction (Ruckleshaus and Hays 1998). Because of this decision, marine habitat loss has continued with the restoration of a terrestrial habitat as mitigation. This by definition fails to comply with the concept of mitigation because the impacts on marine habitats are not alleviated, made less severe or moderated through the restoration of terrestrial habitat.

Some countries, such as France, have legislated to protect seagrasses, whilst others protect seagrasses through other non-legislated means (Gordon 1995). In Australia, marine reserves and parks protect seagrass meadows that fall within their boundaries. Yet, reserves and parks do not stop human impacts and in some cases can increase pressure upon meadows via increased recreational activities (e.g. recreational boating activities, moorings, SCUBA diving and snorkelling) and in some parks limited commercial activity (e.g. Great Barrier Reef Marine Park).

Laws in Queensland require seagrass mitigation where the impacted area is required to be replaced with an equivalent area of seagrass in another location (Gordon 1995). A successful method of transplanting seagrass species must exist in Queensland to achieve this but at present, no successful method has been documented. C. Conacher (pers. comm.) has had success with Zostera capricorni plugs however, long-term success could not be evaluated due to habitat loss: a housing development was established by reclaiming land over the transplant site.

In Western Australia, mitigation is hindered because no large-scale transplantation, restoration, or rehabilitation of seagrasses has been attempted with any success prior to 1996 (Kirkman 1992, 1995; Paling 1995). One such effort is occurring at present: A Cement Company (Cockburn Cement Limited) is financing researchers to reestablish seagrass meadows to mitigate the damage that results from seasand dredging operations. To date the project has been moderately successful with 75-95% of transplants surviving and a number of plugs spreading (Paling et al. 1997; E. Paling pers. comm.).

1.3 Seagrass transplantation: examples A number of transplantation projects that have occurred around the world are documented in Appendix A. In summary, the majority of restoration efforts have focussed on transplanting Zostera marina, using plugs. All efforts have had variable results, with no one method being successful for all species and vice versa. A number of examples from Appendix A are divided into their bioregions (refer to Figure 1.1) and discussed in more detail below.

1.3.1 Worldwide examples Northeast Pacific Bioregion To date, the United States has expended a significantly greater effort in transplanting seagrasses than any other nation, with approximately 680, 000 documented plantings of various seagrass species nationwide (Fonseca 1997). A recent review of 138 transplantation projects in the U.S.A. found that 42% of projects were successful, however only 5% of projects achieved 100% seagrass cover (Fonseca et al. 1996). Generally, the overall percent cover obtained was ~ 42%. An earlier review by Thorhaug (1986) recorded 165 worldwide seagrass restoration efforts with 75 efforts (45%) being successful. Thorhaug (1986) did not standardise “success” in her review, instead, she relied on the definition of success that each researcher had used for each effort. In both reviews, the majority of transplant efforts have occurred in Florida and California. For example, Derrenbecker and Lewis (1983) anchored seedlings and sprigs of Halodule wrightii and Syringodium filiforme with 44-98% surviving. In California, Ware (1993) transplanted Zostera marina using both plugs and anchored leaf bundles. The majority (80%) of anchored leaf bundles survived and plugs had low survival (17%). Hoffman (1990) did not state percent survival however reported that 12 months after a Z. marina bed had been transplanted in Mission Bay, California, the diversity of the meadow had not reached the same value as adjacent natural beds. Thus, the transplantation was effective but equivalency had not been achieved.

Zostera marina transplants in British Columbia, have achieved inconclusive results due to the early reporting of the transplant effort (Harrison 1990; Nomme and Harrison 1991). However, the short-term monitoring of some projects have indicated that anchored vegetative Z. marina transplants have a good potential for success in large-scale efforts (Harrison 1990). For example, in Gibsons Harbour, British Columbia, Z. marina was transplanted to compensate for losses after a rip-rap breakwater was constructed to protect a marina. Anchored and unanchored sods and sprigs were used as the transplant unit (Harrison 1990). After 39 weeks, summer transplants had faired well, although it was too early to obtain conclusive results.

Wider Caribbean Bioregion Florida has had a large number of mitigation projects implemented to compensate for the destruction of seagrass habitats (Thorhaug 1974, 1983, 1985; Ruckelshaus and Hays 1998; Fonseca et al. 1996). Thorhaug

(1983) reported a successful habitat restoration attempt north of Key Largo, Florida Keys. The construction of a pipeline had destroyed portions of Halodule wrightii and Thalassia testudinum seagrass meadows. After ten months Halodule shoots covered 31% of the pipeline trench and Thalassia seedlings had recruited to the area (Thorhaug 1983). Thorhaug concluded by saying that habitat restoration with seagrass in tropical estuaries appeared feasible. A second example is the Florida Keys Bridge replacement project, which also aimed to mitigate seagrass loss, however the mitigation effort was less effective. In this example 93.3 acres of seagrasses (T. testudinum, Syringodium filiforme and H. wrightii) had been lost or damaged by activities such as dredging, infilling, shading, boat propeller cuts and propeller wash, during the replacement of 37 bridges. Mitigation was required through direct transplantation and re-opening/enhancement of coastal lagoons. Only 62.8 acres were deemed suitable to restore and approximately 47.5 acres of seagrasses were planted. Initially, success was low (34.6 acres survived, although many sites lost all transplants), yet long-term (10 year) success was achieved. However, it was not the transplant efforts that were successful, but natural recolonisation, presumably enhanced by mangrove restoration projects, that increased the acreage of seagrasses (Ruckelshaus and Hays 1998).

Results from such efforts have often been minimal, with vegetative propagules being more successful (31 – 71%) than sexual propagules (2.5%) (Thorhaug 1974, 1983, 1985). Many of the efforts have been hampered by high turbidity, with high-energy sites requiring heavy anchoring. Also, success was species and propagule dependent: sprigs of Thalassia testudinum were far more successful than seedlings of any other species, such as Halodule wrightii and Syringodium filiforme (Thorhaug 1974, 1983, 1985). Hence, in this bioregion, mitigation success is variable and often natural recolonisation has out-performed restoration efforts (Thorhaug 1974, 1983, 1985; Williams 1988; Ruckelshaus and Hays 1998).

Northwest Atlantic Bioregion Zostera marina is the usual transplant species used in this region. Results are variable, for example Addy (1947a, 1947b) used seeds and plugs of Z. marina, with plugs being noted as successful but no results were given for the seeds. Foraging cow-nosed rays affected transplants in Virginia (Roberts cited in Phillips 1974). A review by Phillips (1974) concluded that transplantation work performed in this and other bioregions was influenced by ambient salinities causing low germination rates of Z. marina; the small size of Z. marina seeds resulted in them appearing to be ‘lost’ in the field; and seedlings were difficult to locate. Vegetative propagules were viewed more favourably because they gave an instant meadow however, the logistics of moving large plants, over large distances is problematic. Recently, in vitro propagated Ruppia maritima has been used successfully to create new beds and restore seagrass meadows in North Carolina (Bird et al. 1994). From this and previous studies, it is apparent that appropriate site selection and the ability of transplants to take root is essential for success (Phillips 1974; Roberts cited in Phillips 1974; Bird et al. 1994).

Northeast Atlantic Bioregion Zostera noltii and Z. marina var. angustifolia have been transplanted using plugs in Norfolk and Suffolk, Great Britain (Ranwell et al. 1974). After one year, survival of plugs was 100%, which fell to 35% after two years but plugs spread up to four times the original transplant size (330 cm2). The reduction in survival was attributed to an accretion event that had buried many of the transplants. Of interest is that the donor meadow recovered within 5 months of the plugs being removed (Ranwell et al. 1974).

Mediterranean Bioregion In the Mediterranean, early transplant efforts by Cooper (1976, 1977 and 1978) using anchored Posidonia oceanica plugs were successful. Later studies have also successfully transplanted P. oceanica leaf bundles and determined that autumn was the optimum transplant time, that transplants fair better if taken from deeper donor sites and plagiotropic rhizomes should be at least 10-15 cm long (Meinesz et al. 1992; Molenaar and Meinesz 1992; Molenaar et al. 1993; Meinesz et al. 1993; Genot et al. 1994). Thus, vegetative propagules of P. oceanica are a viable transplant option.

Australia/New Zealand Bioregion In Australia, seagrasses have generally recovered slowly from disturbances. An infamous example is in Jervis Bay, New South Wales where seismic blast holes have yet to recover 36 years after the blasting (Kirkman 1989; West et al. 1989). Similarly, detonation holes offshore of Edithburgh, Yorke Peninsula, South Australia, recovered slowly, however when the substrate was stabilised recovery occurred in less than one year (Johnson 1988; Clarke and Kirkman 1989). Kirkman (1989) has described how Posidonia australis meadows in New South Wales, damaged by amphibious vehicles during World War II, have failed to recover. Similarly, areas mined for Posidonia fibre in 1911 have yet to recover in Spencer Gulf, South Australia (Clarke and Kirkman 1989). Hence, in Australia natural recovery is extremely slow and it is considered beneficial to promote recovery of meadows by restoration. To prevent habitat loss a number of seagrass transplant efforts have been attempted in Australia, however these efforts number far fewer than those in the United States or the Mediterranean.

Zostera capricorni has been transplanted in tropical and temperate Australia with varying degrees of success. In Queensland, large plugs of Z. capricorni were transplanted on mud flats, where they grew successfully until a housing development was built on the transplant site (C. Conacher pers. comm.). A restoration project with Z. capricorni in Botany Bay, New South Wales, has been successful to date where the sites are protected and Zostera is spreading to cover the transplant area (P. Gibbs pers. comm.). A previous attempt in Botany Bay, using anchored Posidonia australis and Z. capricorni plugs failed due to storm erosion (West et al. 1990). During 1983 and 1984, Bulthuis (cited in Clarke and Kirkman 1989; S. Campbell pers. comm.) attempted intertidal transplants of Heterozostera tasmanica and Z. muelleri in Westernport, Victoria. The H. tasmanica transplants became desiccated and suffocated under a fine layer of mud, whilst the Z. muelleri

plugs survived. Currently the status of Bulthuis’s transplants is unknown, however it is suspected that they failed to survive (S. Campbell pers. comm.). To date, no restoration work has been performed in Tasmania or the Northern Territory.

A maximum of 12, 000 total units (of varying size and type) of seagrass, in at least 13 transplant efforts, have been made in Western Australia (Paling 1995; summarised in Appendix A), with varying success. The majority (61%) of these transplant efforts used seedlings, followed by plugs (25.5%) and sprigs (13.5%). They all occurred within the Perth metropolitan area (the metropolitan area begins at Yanchep, north of Perth, and extends to Rockingham, south of Perth) or at Rottnest Island. Cambridge (1980) did the earliest transplant work in 1975 using Posidonia australis and P. sinuosa seedlings in tubes. These transplants grew well when water clarity was high and nutrient levels were low, however there was no long term monitoring effort (Cambridge 1980). The vast majority of transplants have been performed by Kirkman, using Amphibolis antarctica and A. griffithii. Rarely have more than 50% of his seedling transplants survived the initial post- transplant period and long term monitoring efforts have shown 100% mortality in all instances (Paling 1995; Kirkman 1989). LeProvost Environmental Consultants (1990) have used P. australis seedlings and had no success, losing all transplants within eight months. Thus, it appears that seedlings would be a poor transplant unit choice given the low success in Western Australia.

Sprigs appear to have been slightly more successful than seedlings, although no long term monitoring data is available for these results and in some cases, as part of postgraduate research, the transplants were removed for biomass analysis. Fifty-five to seventy-five percent of the Amphibolis griffithii sprig transplants Hancock (1992) pegged survived after six months although survival decreased to 0% after the initial period. In a similar experiment, Walker (1994) repeated Hancock’s work using Amphibolis and Posidonia sprigs. On both occasions, success was poor due to the failure of the anchoring system.

In contrast, monoculture plugs of Amphibolis and Posidonia transplanted at the same sites survived (50-70% and > 80% respectively) and spread (Walker 1994). Paling et al. (1997) have also had success (97% survival) using plugs of A. griffithii. Paling’s work is ongoing and has expanded to use plugs of P. coriacea (E. Paling pers. comm.). Mechanical plug transplants by Paling have also been successful since November 1996, with 85% of 0.5 m2 plugs surviving. Few other transplant efforts have used plugs with the exception of D. Walker’s project using A. antarctica in Cockburn and Warnbro Sounds (Paling 1995). Her transplants supported the hypothesis that water quality was a prime-limiting factor in transplant survival. No plugs survived the polluted waters of Cockburn Sound and 100% survival of plugs in Warnbro Sound, where water quality is higher. The failure of other methods in Western Australia combined with the successes of plugs listed above indicate that plugs are an ideal choice of transplant unit in Western Australia.

1.4 Why do seagrass transplants fail? Many mitigation attempts are not iterative and are performed solely as a legal requirement, not as science. Although success and failure are monitored, few attempts are made to understand the mitigation result. Thus, mitigators might know why the attempt failed but the next step in determining how to overcome failure is not reached. This ‘lack of learning’ needs to be re-addressed.

When reviewing transplantation experiments, each researcher has their own intrinsic concept of success. Indicators of transplant success include survival of planting units and percent cover of the target area (Fonseca et al. 1996; Ruckelshaus and Hays 1998). By using transplant survival Fonseca et al. (1996) has summarised the success of transplants in the U.S. Other researchers (e.g. Thorhaug 1974, 1986; Lewis 1987) have relied on previously published criteria of success but do not explicitly state which criteria were used. Biological equivalency, in terms of seagrass-associated invertebrates and fish species composition is rarely used as a measure of success, in part due to the variable nature of such associated communities and the taxonomic overhead required. Kirkman (1992) stated that none of the large-scale seagrass restoration efforts’ undertaken in Australia before 1992 were a complete success. He defined success as transplants colonising and rapidly spreading to provide a new community or resemble a previous community. Failure to meet this objective means that an area is not colonised (i.e. the transplants die) or extension beyond the original transplant unit doesn’t occur.

Initially the objective of this review was to standardise these concepts by using a strict definition of success and failure. Success was measured as the survival and extension of more than 50% of transplant units. So, if more than 50% of the transplant units survive but only 30% spread then the effort has failed. This definition doesn’t consider biological equivalence of the restored and natural habitats, because information on functional equivalency and biodiversity is still lacking. However, each reviewed paper used a different measure of success, often without defining their measure. Thus, I have accepted on ‘face-value’ individual researcher’s definitions of success. It would be useful to have a standard definition for seagrass mitigation success that incorporates the concept of replacing similar cover. For example, the lost meadow had a percentage cover of 30% and the mitigated meadow would have 30% cover or more to be successful.

Several (134) transplant efforts were reviewed examining the type of transplant unit used (sexual or vegetative) and the type of transplant environment (stable or variable) (Table 1.1). Reasons for success and failure of attempts were also reviewed. Statistical analysis was not possible due to the varied methodologies, species and habitats each transplantation effort used and the differing definitions of success between efforts. However, to standardise Table 1.1, environmental conditions were classified as stable if they occurred in deep (littoral >8 m) regions (i.e. a low-energy regime and low turbidity). Typically stable regions were unaffected by storm events and dredging activities. Variable environments were classified as occurring in shallow regions (<8 m to intertidal), with high-energy regimes or capable of having high-energy regimes (e.g. affected

by storms) and capable of high turbidity. Variable environments can be either naturally variable (shallow water affected by wind, waves, storms etc.), anthropogenically variable (affected by human activities including dredge activities, run-off, eutrophication) or both.

A number of factors are evident in this evaluation. First, the low number of studies in Western Australia biases the information presented. Second, Posidonia transplants in Western Australia and on a global scale are superficially successful when sexual and vegetative propagules are used in stable environments. Lastly, all transplanted seagrass species have an increased success rate when using either vegetative or sexual propagules if the conditions are stable. It appears that environmental conditions may be the controlling factor in transplant attempts. Based on these results, a decision framework can be developed to aid in planning seagrass transplant efforts (Figure 1.2). This framework guides site selection, transplant unit selection and transplant methodology, although the latter is species dependent. Thus, the framework can potentially be used to increase the chance of success when transplanting seagrasses.

Table 1.1: Success (%) of transplant attempts reported within the literature with sufficient information to determine environmental stability and transplant success (identified (*) in Appendix A) using sexual or vegetative propagules. Three comparisons were made: a) Posidonia transplants in Western Australia; b) Posidonia transplants from around the globe and; c) all seagrass species on a global scale. (a) Sexual propagule Vegetative propagule

Stable environment 66% (n = 6) 66% (n = 6)

Variable environment 20% (n = 5) 100% (n = 1)

(b) Sexual propagule Vegetative propagule

Stable environment 57% (n = 7) 73% (n = 15)

Variable environment 20% (n = 5) 100% (n = 5)

Stable environment 50% (n = 16) 64% (n = 15)

Variable environment 22% (n = 38) 17% (n = 66)

A number of factors control the fate of transplant success. The most important factors are site selection (stable versus variable environments), transplant unit (sexual versus vegetative propagule) and habitat enhancement, which are often interrelated. To succeed in transplanting seagrasses all three factors must be carefully considered. The framework outlined here incorporates these factors into three stages. Stage I involves

evaluation of a suitable transplant site. Stage II is the selection of a transplant unit that is species and site specific with some consideration of cost:benefit but relying heavily on the likelihood of success. Stage IIb is an expansion of Stage IIa and involves the refinement of the transplant unit, through consideration of transplant unit type and transplant methodology. Stage III is optional and consists of manipulation of the habitat by artificial means. These stages are reviewed below.

Stage I Site selection (water motion and quality (depth))

Stage II a Sexual propagule Vegetative propagule

Stage IIb Seeds Seedlings Sprigs Plugs

Stage III (optional) Anchors, artificial barriers, mats and ASG mats

Figure 1.2: Decision framework developed to aid seagrass transplantation.

1.5 Stage I: site selection Site selection is critical (see Fonseca et al 1996; Ruckelshaus and Hays 1998) and must consider the health of both the donor meadow and the recipient area. To date there is no standardised method for selecting a site. Factors to consider when selecting a site are numerous and include light, water quality and clarity; relative depths of donor and recipient meadow; physical stability; potential for bioturbation by burrowing animals; substrate type; transplant unit spacing and human disturbance.

Water quality has demonstrable impacts on seagrasses. Silberstein et al. (1986) demonstrated a link between seagrass loss and highly eutrophied water, when they hypothesised that increased epiphyte growth was responsible for a reduction in seagrass meadows in Cockburn Sound, Western Australia, during the discharge of nutrient-rich effluent. The eutrophication increased epiphyte growth and phytoplankton blooms, which reduced light levels, killing 3,300 ha of seagrass meadows (Cambridge and McComb 1984). Similarly, seagrass meadows in Princess Royal and Oyster Harbours in Albany, Western Australia, declined because of eutrophication that caused excessive epiphyte growth and phytoplankton blooms, increasing light attenuation (Bastyan 1986). Other researchers have also demonstrated this link between reduced light and decrease in seagrass growth and seagrass depth limitations (Bulthuis and Woelkerling 1981; Bulthuis 1983; Orth and Moore 1983; Dennison 1987; Shepherd et al. 1989; Masini et al. 1990; Duarte 1991; Zimmerman et al. 1991; Dennison et al. 1993; Zimmerman et al. 1994; Zimmerman et al. 1995; Ruckelshaus and Hays 1998).

Current velocity can drastically affect seagrass meadows by erosion. Calm waters generally have continuous meadows and almost continuous annual seedling recruitment (Fonseca et al. 1983). As current velocity increases meadows become patchy, substrate becomes more mounded and recruitment is erratic, until the area becomes too rough for seagrasses to survive (Fonseca et al. 1983; Fonseca and Kenworthy 1987; Duarte and Sand-Jensen 1990b). Fauna at the site can also influence physical stability. Merkel (1990a) successfully used mesh to exclude foraging round stingrays (Urolophus halleri) from transplant areas. Snapping shrimp (Alpheus sp.) and ghost shrimp (Callianassa sp.) are capable of suspending considerable amounts of silt capable of suffocating Heterozostera tasmanica and Thalassia plants (Clarke and Kirkman 1989). Similarly, high water velocity, erosion and sediment loads can directly remove, damage or suffocate transplant units (Merkel 1990a, 1990b; Hancock 1992; Walker 1994). Thayer et al. (1975) noted that seagrasses condition the substrate and become an integral part of the substrate.

Transplants are more likely to survive, grow and develop if they are taken from a donor meadow that is deeper than the recipient meadow (Molenaar and Meinesz 1992). Survival is usually greater when transplants are made at shallow sites where light levels are higher (West et al. 1990; Molenaar and Meinesz 1992). Increased survival of transplants originally from deeper water is hypothesised to be due to increased chlorophyll levels triggered by low light environments (Pirc 1984). The opposite is also assumed to be true: plants taken from shallow water, which are not adapted to low light levels, struggle and often fail when transplanted into deeper depths.

For ease of comparison, the experimental control meadow should be in close proximity to the restored site. This reduces transplant handling and ensures that similar disturbances occur at both sites. The historical presence of plants in the transplant area guarantees that conditions were once suitable for seagrass growth, survival at the sites and that the sediment is appropriate.

Finally, transplant sites should be safe from human disturbance such as dredging, eutrophication, or reclamation. For example, Conacher’s (pers. comm.) transplants were very successful in their early stages, but the area was developed and hence the restored habitat was lost. Thus, restoration efforts will amount to nothing if the restored area continues to be disturbed by activities such as dredging and marina/port developments.

1.6 Stage II: transplant unit Seeds, seedlings, sprigs and plugs of seagrasses are often used as transplant units (Thorhaug 1986; Kirkman 1989; Phillips 1990; Paling 1995). All transplant units (TUs) other than seeds, must have a meristem to expand (Tomlinson 1974). The meristem is usually apical, however when the apex is damaged or destroyed

the next two closest meristems on lateral axes take over and assume a plagiotropic growth mode (Caye and Meinesz 1985a; Meinesz et al. 1991).

Australian researchers (Kirkman and Kuo 1990; Hancock 1992; Kirkman 1992, 1995; Paling 1995; Paling et al. 1997) have frequently transplanted seeds and seedlings. This has been the most common transplant unit for Posidonia species in Western Australia (Cambridge 1977, 1978, 1980; Kirkman 1989, 1995; LeProvost Environmental Consultants 1990; Hancock 1992). To collect seeds or seedlings the fruits are gathered from donor meadows and are held in free flowing aquaria until seed germination and dehiscence has occurred. After dehiscence, the seedlings are grown to a suitable size and then transplanted into meadows. Seedlings can also be collected in situ and either planted in biodegradable pots (Phillips 1990b) or directly into the substrate, with or without an anchoring device. Thorhaug (1986) believes that seedlings of Thalassia testudinum have the greatest potential of success. In Australia, anchoring and mesh/matting are popular methods ensuring sprigs and seedlings remain in place (Kirkman 1989, 1990; Hancock 1992; Walker 1994). Unfortunately seedlings are fragile and are easily damaged when collected or placed into anchoring devices (pers. obs.).

Sprigs (“turions”, “shoots”, or “leaf bundles”) are vegetative shoots with intact roots, rhizomes and leaves, with no sediment attached. The lack of sediment aids in handling but they are easily damaged because the rhizome is exposed. The removal of sprigs from a donor bed is less destructive than plug removal and hence is favoured by some American seagrass researchers (Derrenbecker and Lewis 1983; Fonseca et al. 1985; Fonseca et al. 1987; Tomasko and LaPointe 1991; Durako et al. 1992). Posidonia oceanica researchers, in recent years, have used sprigs (Meinesz et al. 1991; Molenaar and Meinesz 1992; Molenaar 1992; Meinesz et al. 1992; Molenaar et al. 1993). In Western Australia, the success of sprig transplants has been mixed. For example, Hancock (1992) had no success with sprigs of Posidonia and Amphibolis because they were inadequately anchored, the anchor restricted rhizome growth and the sites were inappropriately selected. Alternatively, Gibbs (pers. comm.) has used sprigs of Zostera muelleri, with initial promising results at sites where the water quality was reasonable and the energy regime was calm.

As reviewed previously, plugs are the most successful transplant unit in Western Australia. A plug (also known as a “turf” or a “sod”) is a plant core comprising roots, rhizomes and leaves, which is removed from the substrate with the sediment intact (Phillips 1990b). A hole is excavated to plant the plug, the plug is put in, and the hole is backfilled (Phillips 1990b). Plugs are very efficient at trapping debris, which helps to establish and bind the plug to the recipient site (West et al. 1990). Because of these characteristics, Phillips (1980) and Thorhaug (1986) consider plugs to be ideal transplant units (TUs).

The greatest shortcoming of plugs is that a healthy donor bed is required and their removal damages the bed, contributing to additional net loss of seagrass. If a meadow is about to be dredged or destroyed however,

plugs can be removed and replanted elsewhere, negating the net loss of habitat. Another disadvantage includes an increase in drag and scour upon the plants if the plug is not buried deep enough. Similarly, if an apical meristem is not present then expansion is not guaranteed, especially if the plug is small, reducing the number of available apices (pers. obs.). The use of Posidonia plugs by Western Australian researchers has resulted in mixed success (Hancock 1992; Walker 1994; Paling et al. 1997). Posidonia plugs failed for both Walker (1994) and Hancock (1992) (0-20% and 0-70% survival respectively, with 0% extension for either). However, a recent project by Paling et al. (1997) using mixed P. coriacea and Amphibolis antarctica plugs has had 75-97% survival and spreading rates that are stated to be comparable with control meadows (E. Paling pers. comm.).

Transplant unit spacing is also an important factor to consider (Duarte and Sand-Jensen 1990b; Molenaar and Meinesz 1992; Meinesz et al. 1992; Molenaar 1992). The closer the TUs are to one another the less time it will take for expansion to result in coverage of the transplantation area. Unit spacing becomes increasingly important if the site is not optimal. A further benefit from placing the TUs close together is that they positively influence the hydrodynamics of neighbouring plants.

1.7 Stage III: habitat enhancement methodology Excessive water motion is acknowledged as an impediment to seagrass transplant survival. To overcome this impediment researchers have anchored the TUs to prevent dislodgment by waves, currents, surge (Hoffman 1990; Hancock 1992) and foraging fauna (Merkel 1990a). There are numerous methods that stabilise substrates and anchor TUs. Methods applicable to Australia include: sediment stabilisation using chicken wire and mesh (Carangelo et al. 1979; Merkel 1990a; Walker 1994), biodegradable mesh (Fonseca et al. 1979; Fonseca et al. 1985; Kirkman 1989), artificial seagrass (ASG) mats (Nelson 1992), wire anchors (Thorhaug 1974; West et al. 1990), spun rock fibre anchors (Kirkman 1990, 1995; LeProvost Environmental Consultants 1990; Hancock 1992) and using protective artificial barriers (Cooper 1979; Harrison 1990; Kirkman 1989; Walker 1994). A brief review of these methods follows.

1.7.1 Anchors Plastic and wire pegs, bricks and spun rock fibre are structures used to anchor transplants (Fuss and Kelly 1969; Kelly et al. 1971; Eleuterius 1975; Derrenbecker and Lewis 1983; Thorhaug 1987; Hoffman 1990; Kirkman 1989; LeProvost Environmental Consultants 1990; Phillips 1990b; West et al. 1990; Paling 1992; Hancock 1992; Ware 1993; Kirkman 1995). Success using anchoring is variable: Zostera marina, Amphibolis and Posidonia transplant survival has ranged between 0 – 100% (Merkel 1990a; Nitsos 1990; Thom 1990; Ware 1993; Paling 1995). Generally, anchored Posidonia survival is poor (0 - 50% (LeProvost Environmental Consultants 1990; West et al. 1990; Paling 1992; Hancock 1992; see Appendix A). However, Thalassia testudinum, Halodule wrightii and Syringodium filiforme have all grown successfully when anchored (Fuss and Kelly 1969; Kelly et al. 1971; Thorhaug 1974; Derrenbecker and Lewis 1983; Thorhaug 1987).

Ruppia maritima is one species that has been grown successfully using biodegradable anchoring (Bird et al. 1994). Spun rock fibre (or Growool™) is a biodegradable anchor commonly used in Western Australian experiments (Kirkman 1989, 1990, 1995; Hancock 1992). It is only marginally effective against preventing erosion from displacing the TUs (Hancock 1992). The common failures of a biodegradable anchor are that it may be positively buoyant; it increases surface area, which produces greater scouring and drag on the transplant (Hoffman 1990; Hancock 1992); it hinders seedling primary roots (Hancock 1992) which pull a seedling into the sediment (Kirkman and Kuo 1990); and can damage the rhizomes and roots when placing them into the anchor (Hancock 1992). The type of TUs and sediment stability also influences efficiency of anchoring. Sprigs are the best TUs when using anchors, yet sprigs take longer to stabilise an area because they do not trap sediments effectively. Unlike mesh and artificial seagrass (ASG) mats, sediment stability is not controlled or reduced when using anchors. The overall outcome of using sprigs is that survival, rhizome elongation and establishment potential are reduced. Furthermore, using anchors is time consuming and costly as each transplant must be individually anchored.

1.7.2 Artificial barriers Artificial barriers alter the wave and energy regime and create a protected area with stable sediments where seagrass can be transplanted. However, cost can prohibit building artificial barriers. Artificial barriers can also destroy one habitat (e.g. soft sediment) whilst creating another (e.g. artificial reef). They alter erosion and deposition in an area, as well as the prevailing sea conditions, which may affect other areas. Few artificial barriers have been built to protect seagrass meadows, however; their use needs to be investigated further (Cooper 1979; Kirkman 1989; Harrison 1990; Walker 1994).

1.7.3 Mesh Mesh is another form of anchoring. Wire and plastic mesh has been used with varying degrees of success. Mesh was successful with Zostera marina (Fonseca et al. 1979; Merkel 1990a) and Amphibolis (Walker 1992; Walker 1994). Fonseca’s et al. (1979) meshed Z. marina transplants were very successful, resembling control meadows one year after transplantation. Merkel (1990a) used mesh to protect seagrasses from foraging stingrays. Mesh can be less effective because it causes excessive sediment deposition (Merkel 1990a; Walker 1994), which Walker (1994) demonstrated causes transplant suffocation in Posidonia. Other studies have used the excessive sedimentation when using mesh to control erosion (E. Paling pers. comm.). Paling had high levels of erosion between mixed P. coriacea and Amphibolis antarctica transplant plugs. This was overcome by placing mesh between the plugs, which allowed sediment accretion and, when sediment reached a required level, the mesh was removed avoiding transplant suffocation.

A disadvantage of wire mesh is that it oxidises and is weakened. Plastic mesh does not degrade and hence can cause problems for marine animals, particularly if consumed. Biodegradable mesh is an alternative to wire

and plastic mesh however, it has had limited success. Kirkman (1989 and 1990) and Fonseca et al. (1985) both used this method with 0-47% survival. The greatest disadvantage (which is also an advantage) to biodegradable mesh is that it degrades rapidly, leaving the transplant unprotected. It is also weaker than wire or plastic and for that reason, is less able to withstand storms and other disturbances. Mesh lacks an interacting faunal assemblage that produces detritus, enhancing nutrients, nutrient circulation and aeration of the sediment.

1.7.4 Artificial seagrass mats ASG mats have artificial leaves which slow water currents (Madsen and Warnke 1983; Fonseca and Fisher 1986; Thorhaug 1986; Fonseca and Kenworthy 1987; Short 1987; Merkel 1990a; Nelson 1992; Walker 1994), enhance deposition of suspended material (Ginsberg and Lowenstan 1958; Raupach and Thom 1981; Ackerman 1983; Fonseca et al. 1983; Gambia et al. 1990; Nelson 1992; Nielsen 1992) and provides a three dimensional substrate for benthic settlement (Bell and Hicks 1991; Nelson 1992). ASG mats avoid suffocation of seagrass and, unlike mesh, allow the development of seagrass faunal assemblages that produce detritus.

Few studies have used ASG mats. West et al. (1990) used ASG mats to stabilise a habitat when transplanting Posidonia australis in Botany Bay, New South Wales. They found ASG mats to be expensive and ineffective in improving transplant survival. This experiment however, was hampered by violent storms that led to the termination of the experiment after four months and subsequent results were inconclusive. Gibbs (pers. comm.) has also used ASG mats with transplants in Botany Bay. In this instance the shallow environment was highly eutrophied and by the second year of monitoring epiphyte growth on the ASG leaves had reduced available light to the transplants, resulting in high mortality (P. Gibbs pers. comm.). Thus, it is difficult to establish if ASG mats are viable. Further longer-term studies of the use of ASG mats would determine their utility in seagrass transplantation.

At present, ASG mats are relatively expensive because they are new, experimental and have only been applied on a small scale. Costs would potentially be reduced if ASG mats were used on a larger scale. Further refinement of the placement of ASG mats could also reduce labour and therefore reduce costs and risks associated with diving.

A benefit to using ASG mats is that they are versatile and can be used with a variety of TUs (seeds, seedlings, sprigs and plugs). The placement of mats is simple and quick compared to other anchoring methods. Transplanting seagrass into ASG mats is also relatively quick and easy and involves similar effort to transplanting into mesh. Transplanting using anchors requires slightly less physical exertion, but is more time consuming and requires more ‘bottom-time’ for divers. Thus, ASG mats have the potential to be a good

transplant method in Australian waters, when using Posidonia species. Other reviewers support this conclusion (Phillips 1974, 1980; Thorhaug 1986; Paling 1995).

For each stage of the framework shown in Figure 1.2, there are a number of hypotheses that have been reported in the literature (many of which have become dogma) and others are implicit in various publications. These hypotheses are as follow:

Stage I (site selection):

HIa = Light is a limiting factor for seagrass growth (e.g. Cambridge and McComb 1984; Pirc 1984; Bastyan 1986; Silberstein et al. 1986).

HIb = Water quality and clarity affects seagrass growth (e.g. Bulthuis 1983; Dennison 1987; Cambridge and McComb 1984; Silberstein et al. 1986).

HIc = Seagrasses should be transplanted into similar habitats to where they already exist (Thayer at al. 1975; Duarte and Sand-Jensen 1990a, 1990b; Fonseca 1997).

HId = Physically stable sites have better recruitment and are less patchy (C. Conacher pers. comm; Fonseca et al. 1983).

HIe = Excessive water motion impedes seagrass transplant survival (Hancock 1992; Walker 1994).

HIf = Burial disturbance limits horizontal rhizome growth but stimulates vertical growth (Patriquin 1973; Boudouresque et al. 1984; Meinesz et al. 1992; Gallegos et al. 1993; Molenaar et al. 1993a, 1994b; Marbá and Duarte 1994; Duarte et al. 1997).

HIg = Bioturbation affects seagrass growth (Fonseca et al. 1983; Fonseca and Kenworthy 1987; Duarte and Sand-Jensen 1990b; Merkel 1990a, 1990b).

HIh = Substrate type affects seagrass growth. Stage II (transplant unit and technique):

HIIa = Vegetative propagules are better transplant units than sexual propagules (Phillips 1980; Thorhaug 1986; Paling 1995).

HIIb = Transplant unit spacing is important for rapid patch coalescence and protection of neighbouring transplant units (Duarte and Sand-Jensen 1990b; Molenaar and Meinesz 1992; Meinesz et al. 1992; Molenaar 1992).

HIIc = Transplant handling affects success.

HIId = Transplant size affects success (Walker 1994).

HIIe = Transplants should be made from deeper to shallower environments (Molenaar and Meinesz 1992; West et al. 1990). Stage III (habitat enhancement methodology):

HIIIa = Anchors enhance seagrass transplant survival (e.g. Nitsos 1990; Thom 1990; Ware 1993).

HIIIb = Artificial barriers enhance transplant survival (e.g. Cooper 1979; Kirkman 1989; Harrison 1990).

HIIIc = Mesh enhance transplant habitats (e.g. Fonseca et al 1979; Merkel 1990a).

HIIId = Artificial seagrass mats enhance accretion.

HIIIe = Artificial seagrass mats used in conjunction with transplant units enhance transplant survival.

1.8 Conclusions and aims Seagrasses are an important component of the marine environment. Their distribution and associated biodiversity have been reduced by anthropogenic disturbances. While a variety of mitigation techniques have been attempted, they have met with limited success. The reason for this appears to be a scarcity of basic knowledge concerning site selection and choice of transplant unit and technique. Other limitations relate to biological characteristics of seagrasses, such as meadow dynamics, genetic diversity and seedling characteristics.

It is clear that little attention has been placed upon improving transplant survivorship, the advantages of sexual versus vegetative propagules and the role of colonising seagrass species in seagrass dynamics. Reviews of seagrass transplantation in Australia (Kirkman 1989; Kirkman 1995; Paling 1995; Lord et al. 1999) and internationally (Thorhaug 1986; Lewis 1987; Thom 1990; USNRC 1992; Fonseca et al. 1996) indicate that sexual propagules are the most common transplant choice as they maintain genetic diversity and are less costly than vegetative propagules but success rates are low, making them less cost-effective. Vegetative propagules are still a favoured transplant unit in the published literature (Appendix A) in part because success is higher despite increased cost.

As shown in the field of conservation biology (see Guerrant and Pavlik 1998; Ruckelshaus and Hays 1998), without theoretical knowledge there is little chance of restoration success. Therefore, the major aim of this thesis was to examine various hypotheses put forth by researchers and evaluate the lesser known aspects of seagrass transplantation in Western Australia. Posidonia australis was chosen as the experimental species because it is the dominant seagrass in the study area; it has a ready supply of sexual propagules; its vegetative growth is easy to measure; and its rhizomes are more robust than other species of Posidonia.

With this in mind, this thesis specifically aims to examine Posidonia transplantation in Western Australia, and provide an evaluation of transplant considerations. This chapter reviewed mitigation methods, restoration and transplant attempts and analysed transplant failures. It has resulted in a framework of considerations for seagrass transplantation attempts outlining three stages. These stages form the basis for the following chapters.

Chapter two quantifies the process of site selection (Stage I) in coastal waters of metropolitan Western Australia, by investigating the role of light, burial and substrate type on Posidonia australis seedlings and rhizomes. Additionally the effect of handling disturbance was evaluated (HIIc) to determine the best practice for transplant technique. Mesocosms were used to mimic the natural marine environment, providing an easily

manipulated unit that resembled the Success Bank physical characteristics. Three hypotheses from the literature were evaluated: HIa = light is a limiting factor of seagrass growth; HId = physically stable sites have better recruitment and are less patchy; and HIf = burial disturbance limits horizontal rhizome growth but stimulates vertical growth. Three working hypotheses were developed 1) light limits P. australis propagule growth; 2) disturbance (defined as both burial and handling), limits P. australis propagule growth; and 3) substrate type (as a correlate of water depth; e.g. limestone rubble or sand) limits P. australis propagule growth.

Chapter three examines the question of transplant unit (Stage II) specifically evaluating the hypothesis (HIIa) from the list that vegetative propagules are better transplant units than sexual ones. Sexual propagules of Posidonia australis have long been used within Western Australian transplantation attempts because they are inexpensive, easy to handle and maintain genetic diversity. However, poor success in these attempts suggests that it is not the best transplant unit. Thus, my working hypothesis will be that sexual propagules sustain P. australis communities (through expansion and maintenance) and are suitable seagrass transplant units.

The fourth chapter presents the results of a transplant experiment that was performed over a depth gradient and in an artificially enhanced environment. This chapter examines hypotheses in Stage I (site selection) and Stage III (habitat enhancement methodology). A depth gradient allowed comparison of transplant survival in differing light regimes and energy regimes. This chapter explores the following hypotheses from the literature

HIc = seagrasses should be transplanted into similar habitats to where they already exist; HIIe = transplants should be made from deeper to shallower environments. Erosion in the study region was highlighted as a problem in previous studies (Nelson 1992; Walker 1994; Paling et al. 1997) and hence the site was enhanced using artificial seagrass (ASG) mats. In order to evaluate the effect of erosion and erosion stabilisation three

hypotheses were examined: HIe = excessive water motion impedes seagrass transplant survival; HIIId = artificial seagrass mats enhance accretion and HIIIe = artificial seagrass mats enhance transplant survival.

In the final chapter, the results from the preceding chapters are used as a specific Western Australian case to evaluate the framework developed in Chapter 1. This in turn provides managers with the ability to appropriately and consistently address the issues of seagrass transplantation and hence meet with greatest success.

25 Chapter 2 Optimising the process of site selection for Posidonia australis transplantation in Western Australia

2.1 Introduction Identifying the appropriate site to place seagrass propagules is perhaps the single most crucial factor in a transplantation program. Theoretically, niche concepts can be used to identify transplant locations. The fundamental niche is where a seagrass can establish, survive and reproduce, although the species may not occur in this location at present. Limits to the fundamental niche are generally physiological (Hutchinson 1958; MacArthur 1968) although it also includes regions outside of a species natural distribution and is thus limited by dispersal and historic factors. In contrast, the realised niche is where a species is currently found and includes habitats where, for example, physical parameters such as depth, salinity, wave action and available light are similar. The realised niche is a subset of the fundamental niche and in general its limits are synecological processes (such as competition, predation or herbivory) although autecological limits, such as dispersal, can be significant factors (Hutchinson 1957; MacArthur 1968). Examples of available areas in the realised niche include gaps in seagrass meadows, areas adjacent to meadows and areas where disturbance has directly reduced seagrass cover (e.g. Duarte and Sand Jensen 1990; Fonseca 1997). Thus, a seagrass may only be present in shallow water (its realised niche), yet if transplanted to deeper water it can survive because of the breadth of its fundamental niche.

Based on the concepts outlined above, areas available for seagrass meadow expansion or colonisation would also be suitable for seagrass transplantation. Site selection begins with the process of determining the key parameters that affect seagrass growth and survival. The process of site selection per se is not often reported in the literature. Historically, measured parameters that affect seagrass growth include water depth, current strength, burial, substrate type, water temperature, salinity levels, predation, disturbance and nutrient levels (Fonseca et al. 1983; Littler et al. 1983; Fonseca and Fisher 1986; Fonseca and Kenworthy 1987; Gambi et al. 1990; Nelson 1992; Baron et al. 1993; den Hartog 1994; Marbà and Duarte 1994; Duarte et al. 1997; Fonseca and Bell 1998). In terms of importance, the following factors appear to affect all seagrass species distributions and are paramount when selecting an appropriate transplant site: light levels, disturbance (burial) and substrate type. These are reviewed below and a brief description of optimal and sub-optimal sites follows.

2.1.1 Light levels Light levels are a primary limiting factor in seagrass growth, distribution, abundance and biomass (Drew 1978, 1979; Bulthuis and Woelkerling 1981; Bulthuis, 1983; Orth and Moore 1983; Cambridge and McComb 1984; Dennison and Alberte 1985; Dennison 1987; Dawes and Tomasko 1988; Williams 1988; Larkum and West 1990; Masini et al. 1990; Duarte 1991; Zimmerman et al. 1991; Zimmerman et al. 1994). Light levels decrease as a function of depth and water quality or turbidity. Species light requirements act to restrict both depth and regional distribution.

26 It is apparent from the literature, that saturation (Isat) and compensation (Ic) values for different species vary widely (Table 2.1). Despite this, the average minimum light requirement for seagrass species is close to 11% of surface irradiance (SI) but differs seasonally (Drew 1978; Duarte 1991), and would be expected to vary with latitude. For example, Zostera marina, in Denmark, requires a minimum of 8-11% SI (Olesen & Sand Jensen 1993), while Halodule wrightii, in Texas, requires ~18% SI (Duarte 1991; Dunton 1994). Posidonia oceanica and Cymodocea nodosa, in the Mediterranean, each require ~10% SI (Bulthuis 1983). In Western Port and Port Phillip Bay, Victoria, Heterozostera tasmanica may only need 5% SI to establish and survive (Bulthuis 1983). Despite the fact that no correlation between species life history strategies (r- and K-selected) and their reported saturation and compensation values is readily apparent (Table 2.1), it is hypothesised that r-selected species such as H. tasmanica are low light adapted. This would act to maximise their light harvesting ability while reducing respiratory demands and self shading (Bulthuis 1983, Josselyn et al. 1986; West 1990; Duarte 1991). A latitudinal gradient in measured saturation values is observed in Table 2.1 for both r- and K-selected species.

-2 -1 -2 -1 Table 2.1: Light saturation (Isat μmol m s ), light compensation (Ic μmol m s ) levels and life history strategies (r-colonisers and K-climax species) for seagrass species. Data was extracted from the following: 1) Masini et al. 1990; 2) Bulthuis 1983; 3) Dennison and Alberte 1985; 4) Drew 1978; and 5) Drew 1979. * Measurements were made on the introduced population of H. stipulacea in the Mediterranean.

Species Life History Strategy Study Site Isat Ic (r or K-selected) Latitude

Halophila stipulacea 5* r 40º N* 92 9

Heterozostera tasmanica 2 r 39º S 95 26

Zostera marina 3 r 40º N 100 15-25

Zostera angustifolia 5 r 40º N 147 14

Cymodocea nodosa 4 r 40º N 175 18

Amphibolis antarctica 1 K 31º S 55 20

Posidonia sinuosa 1 K 31º S 73 25

Posidonia australis 1 K 31º S 92 23

Posidonia oceanica 4 K 40º N 120 18

Phyllospadix torreyi 5 K 40º N 166 23

Light levels in the marine environment are reduced by anthropogenic activity. For example, eutrophication reduces available light through increased epiphyte growth on the seagrass and phytoplankton growth. In Cockburn Sound (Cambridge et al. 1986; Silberstein et al. 1986; Shepherd et al. 1989), Princess Royal Harbour in Western Australia (Bastyan 1986; Mills 1987; Hillman et al. 1990; Hillman et al. 1991) and Werribee, Port Phillip Bay in Victoria (S. Campbell pers. comm.; unpub. data)

27 reductions of seagrass cover have occurred due to eutrophication. Increased urbanisation and poor management of catchment areas have further reduced available light through gelbstoff, resulting in seagrass meadows threatened by reduced light availability.

High light irradiance may also limit seagrass growth by inhibiting photosynthesis through harmful photo- oxidation reactions, termed photoinhibition (Valiela 1991; Salisbury and Ross 1992). Photoinhibiton is a controversial topic due to different prevailing thoughts on what causes the inhibition. Current theory suggests either the loss of the herbicide-binding DI polypeptide of photosystem II or damage to a functional group in the reaction centre (see Critchley 1988). In Australia, mangrove (Avicennia marina) responses to excessive light levels reflect a regulatory and protective response rather than damage to the reaction centre complex of photosystem II (Björkman et al. 1988). If a seagrass is optically dense it may not be affected by photoinhibition, as is the case in the adults of the marine alga Codium fragile ssp. tomentosoides (Ramus et al. 1976a, 1976b; Ramus 1978; Sealey et al. 1990; Trowbridge 1998). Photoinhibition has yet to be examined in seagrasses, with Hillman et al. (1989) suggesting that it has not been demonstrated in any native Australian species.

2.1.2 Disturbance: erosion and accretion Movement of sand directly threatens seagrasses by smothering, scouring and eroding the immediate substrate (Littler et al. 1983; Marbà and Duarte 1994; Duarte et al. 1997). Constantly moving sediments necessitate continual recolonisation (Littler et al. 1983; Baron et al. 1993; Marbà and Duarte 1994; Duarte et al. 1997). Smothering reduces light availability and changes the nutrients and dissolved gasses surrounding the rhizomes (Littler et al. 1983; Barko and Smart 1986). Scouring physically damages seagrasses and erodes the substrate. Erosion of the immediate substrate removes sediment from under the growing edge of the meadow, creating a hole or blowout. This erosion can be as great as 20 cm in depth, leaving the rhizome exposed. Seagrasses growing on the edge of erosional scarps have reduced vertical growth; the opposite is true for seagrasses growing in accreting environments (den Hartog 1994; Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997).

When sediment levels fluctuate and excessive accretion leads to burial, orthotropic (vertical growth) shoot growth is stimulated enabling plant survival (Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997). Increasing orthotropic shoot growth ensures that shoots quickly reach light, guaranteeing photosynthesis and relocating the apical meristem to the surface (Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997). For example, Thalassia testudinum vertical growth can increase to 30 mm y-1 when plants are buried (Marbà et al. 1994). Stimulation of orthotropic shoot growth also occurs in coastal dune plants where, in similar fashion to seagrasses, plant growth fluctuates with sediment depth (Marshall 1965; Wallen 1980; Eldered and Maun 1982; Disraeli 1984; Maun and Lapierre 1984; Zhang and Maun 1990).

Seedlings also respond to burial via stimulated growth, although deep or prolonged burial can cause high mortality (Marbà and Duarte 1994). For example, many Cymodocea nodosa seedlings can tolerate burial

28 of < 7 cm, but not deeper (Marbà and Duarte 1994). Prolonged burial increases mortality in adult plants and hence not all seagrasses can survive burial (den Hartog 1994; Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997). For example, Thalassia hemprichii and C. rotundata respond rapidly to burial (increased internodal growth), however both species fail to recover from even moderate burial (Duarte et al. 1997). Zostera marina also deals well with short-term burial but prolonged hypoxia results in high mortality rates (den Hartog 1994). The immediate reaction of seagrasses to burial is to increase vertical growth, regardless of a species’ capability to survive (Duarte et al. 1997). Thus, seagrass meadow cover fluctuates with burial.

Plants that can withstand scouring and smothering are stress-adapted. Littler et al. (1983) found that species most affected by sand-stress tend to have high productivity, low biomass and opportunistic life histories. This description fits the characteristics of r-selected species (Krebs 1985; see Chapter 3). Conversely, sand-stress tolerant species are long-lived, have slow growth rates, which results in low rates of recolonisation (Littler et al. 1983); characteristics exhibited by K-selected species (see Chapter 3). This is seen in the field, where r-selected species such as Halophila ovalis and Heterozostera tasmanica are easily displaced after disturbance (e.g. winter storms). Alternatively, K-selected species such as Posidonia australis, Thalassia testudinum (Gallegos et al. 1993; den Hartog 1994; Marbà and Duarte 1994; Duarte et al. 1997) and Phyllospadix scouleri (Littler et al. 1983), can withstand disturbance and persist after such events. However, this is not always the case, as r-selected species such as Zostera marina and Cymodocea nodosa can tolerate moderate burial (Gallegos et al. 1993; den Hartog 1994; Marbà and Duarte 1994; Duarte et al. 1997).

Erosion and accretion are common phenomena in coastal Western Australian waters. Seagrass meadows in these regions are buried by accreting sediments during winter and early spring (Nelson 1992; Walker, 1994). Burial is usually temporary however; it can last for extended periods. For example, mixed Posidonia sinuosa, and P. australis meadows at Carnac Island survived burial lasting for 3-4 months (Nelson 1992). No published studies have yet determined how P. australis rhizomes and seedlings react to burial.

2.1.3 Substrate type Seagrasses grow on a range of substrates and utilise them for a number of functions such as anchoring plants, seed banks (Pirc et al. 1980; Churchill 1992; Harrison 1993; Moore et al. 1993; Baskin and Baskin 1998), nutrient sinks and recycling areas for decomposing detritus and nutrients, and as habitats for micro-organisms (Moriarty and Boon 1989). Substrate type also influences benthic recruitment; areas devoid of ‘barriers’ (e.g. anthropogenic debris, reef, blowout edges, worm tubes) recruit fewer seeds and seedlings than areas with barriers (Eckman 1983). For example, Amphibolis species and Phyllospadix scouleri use a “grappling hook” apparatus near the rhizomes to attach and anchor to substrates, allowing seedlings to establish on meadow edges, reefs, old seagrass rhizome mattes and rocky substrates (Ducker et al. 1977; Kuo et al. 1990).

29 Water depth and climate are the principal controlling factors that influence substrate composition. In general, seagrass substrates are mostly unconsolidated and anoxic. In deeper water, substrates tend to consist of fine homogeneous particles whereas in shallow water they consist of coarse particles that are non-uniform. Substrates can either be carbonate-based and lacking phosphorous, or terrigenous and lacking nitrogen (Short 1987). Typically, tropical regions have carbonate substrates resulting in seagrasses from these regions tending to be phosphorous limited, while temperate regions have terrigenous substrates and the seagrasses tend to be nitrogen limited (Short 1987). Substrate and leaf uptake supply nutrients to the plants, with leaf slough and detrital matter replenishing nutrients in the habitat via the nutrient cycle (Short 1987; Hemminga et al. 1991; Hemminga et al. 1994). The nutrient cycle and nutrient limitation in seagrass meadows and macrophytes has been thoroughly examined (e.g. Iizumi et al. 1980; Bulthuis and Woelkering 1981; Short 1981, 1983a, 1983b, 1987; Iizumi et al. 1982; Short and McRoy 1984; Roberts et al. 1984; Short et al. 1985; Barko and Smart 1986; Barko et al. 1988; Barko et al. 1991; Hemminga et al. 1991; Paling 1992; Hemminga et al. 1994) and will not be expanded upon here.

Barko et al. (1991) has demonstrated that the physical and chemical properties of sediments are a product of macrophyte growth as well as potential delimiters of growth. Although plants may exhibit a preference, through better growth in certain substrates, plants are usually able to grow on a selection of substrates by adapting their morphology (Denny 1972; Barko et al. 1991). Rhizome growth in less fertile sediments is maximised and plants have high root:shoot (R:S) ratios (Sand-Jensen and Søndergaard 1979; Barko and Smart 1983). Plants in fertile environments maximise shoot production and have low R:S ratios (Barko et al. 1991). Another morphological adaptation is seen in Phyllospadix scouleri, which has adapted to grow on rocky littoral substrates (Turner 1985; Cooper and McRoy 1988). It survives the oxic, high water motion of the littoral range by thickening its leaves and rhizome epidermal tissue (protection from abrasion), reducing the number of lacunae and developing more root-hairs and hypodermal fibres (Cooper and McRoy 1988). Another example of modified morphology is in the seagrass Posidonia oceanica, which lives in high water motion areas of the Mediterranean and has extensively developed hypodermal fibres (Molinier and Picard 1952).

In Western Australia, Posidonia australis occurs on sand, silt and rock. The effects of these substrate types on P. australis anatomy and morphology are not fully known. However, if a specific substrate offering better rhizome and/or seedling establishment and growth potential were known then the success of seagrass transplantation efforts could be improved through appropriate site selection.

2.1.4 Optimal and sub-optimal sites In terrestrial communities early revegetation work was often attempted to stabilise eroding soil or to produce economically important areas (Jordan et al. 1987). For example, the beginning of ecological restoration technology began in the 1930s in the USA, in response to the ‘Dust Bowl’ and deforestation of the Great Lakes (Jordan et al. 1987). Presently, terrestrial and aquatic restoration ecology has progressed to such a point where it is used to restore disturbed ecosystems, halt habitat fragmentation, conserve

30 habitats, maintain species diversity and to protect rare and endangered species (Jordan et al. 1987; Norse 1993; State of the Environment Advisory Council 1996; see Chapter 1). For ecological restoration to succeed, the selected site must be optimal for the species and the restoration objectives. In terrestrial and freshwater habitats site selection is often relatively easy, however marine habitats have been difficult to manipulate and hence site selection becomes crucial (Norse 1993).

As discussed above, several characteristics influence seagrass growth and need to be considered when selecting transplantation sites. If the site has factors best for the target seagrass species’ growth, then the site is ideal and therefore “optimal”. However, the site will often not be ideal; instead the site factors are more likely to fall somewhere between ideal and inferior, or sub-optimal. The identification of such optimal sites is necessary for seagrass transplant growth, survival and spread. Yet, often sub-optimal sites are used in restoration work. This occurs because economics is often a pre-determining factor when selecting restoration habitats. Economic factors are of prime concern in mitigation work (see Chapter 1). For example, if a company damages or destroys an area, it may be obligated, legally, to restore this area, even though the area is sub-optimal and cannot currently be restored. Hence, we need to strike a balance between choosing optimal and sub-optimal sites when performing transplantations and ecological restoration.

2.1.5 Aims The process of site selection for seagrass transplantation in Western Australia is of prime importance, due to its continual loss from the W.A environment and its commercial and recreational fisheries links, and thus needs further experimental study to determine the effects that reduced light levels, different substrate types, burial and disturbance has on Posidonia australis rhizomes and seedlings. This chapter aims to examine P. australis rhizome and seedling growth in reduced light levels, different substrate types, burial and disturbance, with the purpose of establishing a protocol for selecting transplant sites for P. australis in Western Australia. As stated in Chapter 1, the working hypotheses tested in experimental mesocosms are:

HIa = Light is a limiting factor for seagrass growth;

HIe = Excessive water motion impedes seagrass transplant survival;

HIh = Substrate type affects seagrass growth and;

HIIc = Transplant handling affects success.

31 2.2 Materials and methods 2.2.1 Mesocosm establishment An outdoor, free-flowing seawater mesocosm laboratory was established at Woodman Point for experimental manipulations (Figure 2.1). Initially 12 mesocosm were built (for pilot studies), however an additional six were added for increased replication making 18 for the experiments. Each mesocosm had a capacity of 120 L and an average flow rate of 240 L h-1. Acid washed sand was used as a substrate in each mesocosm (with the exception of substrate type experiments). Mesocosms were cleaned twice weekly to remove the build up of diatoms and filamentous algae. Water was filtered through two sand filters to remove excessive silt and biological debris (Figure 2.3). An Ozomatic® ozone generator was installed inline to kill excessive bacterial and fungal growth (Figure 2.4).

Figure 2.1: Mesocosm laboratory at Woodman Point, Western Australia.

2.2.2 Fruit and rhizome collection Fruit collection A total of 2339 fresh Posidonia australis meadow fruits were collected directly from Success Bank (n = 582), Wreck Rock (n = 372) and Woodman Point (n = 1385) meadows. Comprehensive descriptions of these sites are provided in Chapter 3. Fruits were collected by divers in a haphazard fashion with multiple fruits collected from the same plant. The density of fruits varied between sites, limiting the total number of fruits collected. After collection, all fruits were washed in a disinfectant (1 part Miltons® baby bottle wash to 80 parts seawater) and randomly placed into free-flowing seawater mesocosms, maintaining sites separate. Many Western Australian seagrass researchers use Miltons® because, like chlorox (the preferred disinfectant in the United States; S. Wyllie-Echeveriia pers. comm) it prevents plant infection

32 (C. Simm pers. comm.). Fruits were allowed to dehisce and release their seedlings. Seedling growth was also recorded and described in Chapter 3 (section 3.2.4).

Rhizome collection Posidonia australis rhizomes were removed from a depth between 5.0 and 9.0 m at both Success Bank and Wreck Rock meadows. Rhizomes were washed in disinfectant (as above), labelled (showing collection site and rhizome identification number) and then six randomly selected rhizomes from each site were placed into each mesocosm, resulting in a total of 12 rhizomes in each mesocosm. Each rhizome consisted of at least five shoots and an apical meristem to ensure rhizome growth could occur (Kirkman 1989). At the beginning of each experiment (and for replicate experiments), fresh rhizomes were collected.

Rhizome elongation Apical meristems were tagged using methods similar to those of Caye and Meinesz (1985a), Dennison (1990c), Tomasko and Lapointe (1991) and Erftemeijer et al. (1993) with a small plastic “cable tie” marking the last shoot on the rhizome prior to the apical meristem (Figure 2.2). Thus, in situ rhizome elongation beyond the tagged area was used as an indication of growth while reductions in rhizome length was used as an indication of necrosis. In situ rhizome measurements were made at Success Bank on Transect A (see Chapter 3, Table 3.1). This method is appropriate for in situ measurements, however it is unsuitable for measuring growth of rhizomes removed from their natural environment (such as mesocosms, vegetative propagules or transplant plugs) because the sum of two opposing factors (growth by the terminal meristem and necrosis at the basal region of the rhizome) could make apparent length seem unchanged. Molenaar and Meinesz (1992) and Meinesz et al. (1992) discuss the difference between real length and apparent length of rhizomes and have demonstrated how to overcome the canceling effects of growth and necrosis by using one of four methods, described below.

Direct measurement of elongation by measuring growth increments, not total length of the rhizome (i.e. growth in front of the tagged area). This method was selected in this study: rhizome elongation in both mesocosm and transplant Posidonia australis plants was measured by growth increments. This method was selected because it has been effective for the Mediterranean cogener P. oceanica (Molenaar and Meinesz 1992; Meinesz et al. 1992).

The second method standardises the rhizome into a known number of segments (which can be based on internode length) and growth beyond the original number of segments is recorded. Like in situ measurement, the apparent length of the rhizome and number of segments may seem unchanged because of growth and necrosis. Standardising rhizome segments has been used by Schwarzschild et al. (1994) to measure the effects of atrazine on the growth of Zostera marina.

The third method measures the forward movement of a patch of seagrass. This has been used by Duarte and Sand-Jensen (1990a) and Olesen and Sand-Jensen (1994) to measure Cymodocea nodosa patch

33 growth in the Mediterranean and Zostera marina patch growth in Denmark. However, this method does not rationalise that a patch of seagrass can have an extending edge and a receding edge, much like a blowout and hence, patch growth can be overestimated. Also, since direct measurement of rhizome growth is not made it is impossible to know if rhizomes are elongating or patch recruitment is occurring.

The final method indirectly measures rhizome growth through reconstruction techniques of past growth history, such as plastochrone interval. This method is favoured by many researchers (Brouns 1985, 1987; Pergent and Pergent-Martini 1990; Gallegos et al. 1993; Molenaar et al. 1993; Duarte et al. 1994; Gallegos et al. 1994; Marba et al. 1994) and is used if time or financial constraints exist or if a large database of species plastochrone information is available. Reconstruction techniques have rarely been used for Australian Posidonia species (West 1990), although Pedersen (pers. comm.) has attempted to demonstrate seasonal patterns in the rhizome length-sequence in some Australian species.

Rhizome tag

Figure 2.2: A tagged Posidonia australis rhizome. A cable tie marks the last shoot on the rhizome prior to the apical meristem.

Figure 2.3: Sand gravity filters used to stop excessive silt and biological debris from entering the mesocosms. (“Quobba” the dog guarding the sand filters from entry of silt and biological debris).

Figure 2.4: Ozomatic® ozone generator used to sterilise the water before it entered the mesocosms.

2.2.3 Pilot Studies An initialisation period (where the mesocosms flowed but no seagrasses were present) ran for five months and determined that baseline differences in light and temperature between mesocosms were negligible. Temperature and light were measured using data-loggers manufactured by Dataflow Systems Pty Ltd. Data during this period were compared with field transect data (Chapter 3) to note if mesocosms were similar to the field environment. There was no statistical difference between mesocosm and field transect temperature (t[36] = 1.32; p > 0.05). However, light levels were significantly higher in the mesocosms

(t[36] = 0.41, p < 0.05). To compensate for the higher light levels, the mesocosm laboratory was shaded, which reduced light levels to within statistically similar levels to those found at 5 m on Success Bank

(t[36] = 1.26; p > 0.05).

Caye and Meinesz (1985a) found that seagrasses are susceptible to fungal and bacterial infections. In order to establish if such infections posed a threat in the mesocosm environment, a short pilot experiment was run. Twenty seedlings were randomly planted in 6 mesocosms (n = 120) and 20 rhizomes were randomly planted in the remaining 6 mesocosms (n = 120). Of these seedlings and rhizomes, 10 each were pre-treated with a disinfectant solution (1:80 ratio of disinfectant to seawater), while the rest were left untreated. The results demonstrated that fungal and bacterial infection occurred in 100% of rhizomes and seedlings if left untreated, whilst few (< 2%) disinfected seedlings or rhizomes succumbed to infection. This indicated that a disinfectant regime was required when using the mesocosms. Thus,

36 subsequent mesocosm experiments were pre-treated with a disinfectant solution to reduce the risk of infection.

A second plot study was conducted to evaluate net rhizome growth. Collected Posidonia australis rhizomes were tagged (as above) and added to the mesocosms. The average of net growth rates, measured every 14 days for a period of 98 days, of the 20 rhizomes in each mesocosm was used as the datum for analyses (n = 12). Net rhizome growth was defined as an increment in length beyond the tagged area (as described above), not the total length of the rhizome. This method attempts to compensate for losses at the terminal end of the rhizome and account for true growth in front of the tagged region (Figure 2.5a). It must be emphasised that negative growth values can occur using this method, if losses at the terminal end encroach into the tagged region (Figure 2.5b), or if the apical end becomes necrotic (Figure 2.5c). If the necrotic region of the terminal end encroaches beyond the tagged region, then a negative net growth value can still result from a positive growth value if growth is exceeded by necrosis (see Figure 2.5b).

a) b) c)

Tag Tag Tag Apical end Apical end Apical end

Growth Growth Growth

Terminal Terminal Terminal end end end

Figure 2.5: Rhizome growth measured as a) increase in rhizome length beyond the tagged region; b) necrosis encroaching past the tagged region resulting in a decrease in growth and; c) necrosis at the apical end of the rhizome resulting in a decrease in growth. Necrosis is shown by the shaded ! region.

Rhizomes were replaced with experimental propagules (rhizomes and seedlings) at the completion of the pilot studies. Experimental propagules were left to stabilise for 30 days before the manipulation experiments began (described below). Propagules were replaced with fresh material after each manipulation experiment and a stabilisation period of 30 days occurred. No measurements were made during the stabilisation periods.

2.2.4 Manipulation experiments All manipulative experiments were performed three times (on three separate occasions). No significant differences between the results in each of the three experimental replicates were detected and all subsequent data is a mean of all three experiments.

In each of the experiments (light, burial disturbance, handling disturbance and substrate), four mesocosms -2 -1 were used as controls. These were uncovered and remained in full light (average 407.35 µmol m s ).

37 Rhizomes were allocated to mesocosms randomly, ensuring that a range of rhizome sizes was represented in each mesocosm. Similarly, allocation of treatment to mesocosms was done randomly. The measurement of rhizomes and seedlings is described below for each experimental.

The following sampling strategy was employed in all experiments except disturbance: rhizome and seedling growth in seven randomly selected mesocosms were measured weekly; in an additional seven randomly selected mesocosm measurements of rhizome and seedling growth were made at the beginning and the end of the experimental period; the remaining four mesocosms were used as controls and measured weekly. In the disturbance treatments, propagule growth was measured weekly, fortnightly and at the beginning and the end of the experimental period.

It is difficult to know what affects elongation rates without direct measurements of a range of responses such as carbohydrate stores and photosynthesis. In these experiments net elongation rate and hence growth was implied via increases in net rhizome length. An increase in net rhizome length beyond the tagged area is considered to be growth, whilst a decrease in net length beyond the tagged area is necrosis.

Seagrass rhizome and fruit collection Rhizome and fruit collection was performed as described above. Twelve Posidonia australis rhizomes were placed in each of the 18 mesocosms with six rhizomes from Success Bank and six rhizomes from Wreck Rock. Fruits soon dehisced and releasing seedlings. The split pericarps were removed from the mesocosms as soon as dehiscence occurred. Ten seedlings per mesocosm were used in each of the experimental treatments. In the event that not enough seedlings were available (due to mortality), dives (n = 3) were made to collect fresh in situ seedlings, from Woodman Point and Wreck Rock. These seedlings were then tagged for locality and allocated randomly to the mesocosms. The rhizomes and seedlings used in the treatments are referred to as propagules. Propagules were anchored into the sediment by covering their rhizomes with sediment.

Light experiment Two light treatments were shaded (25% SI) and heavily shaded (5% SI). Light was measured using data- loggers manufactured by Dataflow Systems Pty Ltd. The 25% SI treatment ran for 36 days (n = 7), while the 5% SI treatment ran for 44 days (n = 7). The 25% SI treatments had surface irradiance reduced by using green shade cloth manufactured by Jaylon Pty. Ltd. The shade cloth reduced irradiance by almost 75% SI (407 µmol m-2s-1 to 102 µmol m-2s-1). However, irradiance levels remained above saturation (73 µmol m-2 s-1) and compensation levels (25 µmol m-2 s-1) recorded for Posidonia australis in Albany, Western Australia (Masini et al. 1990; Table 2.1). The 5% SI treatments were covered by black plastic reducing irradiance to below compensation levels reported for P. australis in Albany (Masini et al. 1990). Actual irradiance levels within the 5% SI mesocosms averaged 22 µmol m-2s-1. Measurements of rhizome elongation occurred approximately weekly.

Burial disturbance The irradiance levels in all treatments averaged 407 µmol m-2s-1. The initial burial treatment was cyclical, involving the burial and uncovering of propagules. Burial consisted of covering the propagules with acid washed sand, so that no leaves were left exposed. Such burial occurs frequently within the field transects examined in this study (Nelson 1992; Walker 1994; pers. obsv.). After a burial period of 21 days, the propagules were uncovered (sand was fanned away), allowing leaves to be exposed to light. The propagules were left uncovered for a further 21 days, allowing them to ‘recover’ from the burial. At the end of this period, the propagules were then re-covered with acid washed sand and left for a further 21 days. At the end of this period, the treatment was halted and final measurements were made. The experiment lasted for a total of 63 days.

Complete burial treatment of propagules was performed in the same manner as cyclical burial, with the exception that the propagules were left buried for 43 days, the complete experimental period. Controls and treatment propagule growth was measured approximately weekly.

Handling disturbance The irradiance levels in all treatments averaged 407 µmol m-2s-1. To examine handling effects (disturbance when measurements were made) propagules in four mesocosms were only measured at the beginning and the end of the treatment period. These are referred to as the “no handling” treatment. A further five mesocosms were deemed “infrequent handling” and had their propagules measured every second measurement period (i.e. measurements occurred fortnightly). The propagules in a further five mesocosms were measured weekly and are referred to as the “constant handling” treatment. The final four mesocosms acted as controls. The disturbance experiment ran for 141 days. No significant difference in growth rate was detected between treatments indicating that the frequency of measurement would not alter growth rates of the rhizomes or seedlings. Hence, measurements for subsequent experiments were made approximately weekly.

Substrate type Two substrate treatments, acid washed sand and limestone rubble of a similar size as that available in the Woodman Point area, were examined. Propagules were anchored into the sediment by burying the rhizomes in the sand or covering the rhizomes with rubble. This experiment ran for 48 days, with seven mesocosms having sand and seven having rubble substrates. A silt substrate could not be tested, as the water flow into the mesocosms was too turbulent resulting in constant resuspension of the substrate. Propagule growth measurements occurred approximately weekly.

39 2.2.5 Statistical Analyses Total rhizome growth rates were compared between field and experimental controls using student t-tests (t statistic). Temporal pseudoreplication is a necessary component of the mesocosm experiments and therefore one way RM ANOVA was used to test for similarity between treatments. Repeated measure one way ANOVA non-parametric analyses were conducted using a Friedman’s RM ANOVA test (Zar 1996).

Significant effects were analysed using a Dunn’s test (when the number of observations per treatment group was not equal) or a Student-Newman-Kuels (SNK) test (when the number of observations per treatment group were equal) (Winer 1971; Glantz 1992).

2.3 Results 2.3.1 Pilot studies Results from this pilot study demonstrated that Posidonia australis rhizomes recovered from the stress of removal from the field meadow within 20-40 days. Also, net rhizome growth was not significantly different between mesocosms. During the pilot study net rhizome elongation occurred at a rate of 0.96 mm d-1 in spring and 0.70 mm d-1 in summer, which was comparable to net growth in the field environment (1.0 mm d-1; see Chapter 3). Thus, it was determined that P. australis rhizomes grew within the mesocosms after an initial period of adjustment to the mesocosm environment and that rhizome

growth in all the mesocosms was similar (t[69]= 1.14; p > 0.05) and comparable to the field environment. Net growth rates between controls in the pilot studies and controls in the experimental treatments were significantly different (t[69] = 9.32; p < 0.05), with higher net growth occurring in the pilot studies.

2.3.2 Manipulation experiments Light Net rhizome growth fluctuated in the reduced (25% SI) light environment (Figure 2.6a). Field rhizomes grew at significantly greater rates than control rhizomes (t[9] = 6.19, p < 0.05; Figure 2.6a). The treatment 2 and control net growth rates were similar, irrespective of the origin of rhizomes (χ [9] = 0.22, p > 0.05; Figure 2.6a, Table 2.2). Seedlings survived for 15 days. Some seedlings produced rhizomes, which grew to an average length of 6 ± 0.20 mm (0.4 mm d-1) and did not produce root hairs. Seedlings had on averaged 3 emergent leaves. All seedlings died after 15 days in 25% SI conditions.

Reducing the light levels to 5% SI caused necrosis in all of the rhizomes. Rhizome net growth rates in the field were significantly greater than growth in the control treatment (t[9] = 4.39, p < 0.05; Figure 2.6b).

Rhizome growth and necrosis rates were similar between Wreck Rock and Success Bank (t[5] = 0.195; p > 0.05). Generally, rhizome growth fluctuated, with necrosis being recorded on all but two occasions and 5% SI rhizome necrosis was significant up until day 15 (t[7] = 0.012; p < 0.05; Figure 2.6b). Overall, the net growth rate was similar irrespective of site where the rhizomes came from and did not vary 2 significantly between treatment and control (χ [9] = 2.12, p > 0.05; Table 2.2).

40 A high percentage (79%) of seedlings survived in the reduced light conditions. The seedlings produced rhizomes that grew at a rate of 1.2 ± 0.04 mm d-1 and had root hairs. Seedlings had produced on averaged three emergent leaves in eight days and 4 leaves by 15 days of the experiment starting. At the end of the experiment, 52% of seedlings had three shoots.

Table 2.2: The mean net rhizome growth rates (mm d-1 ± SE) in 25%, 5% SI and control treatments. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth.

Mean net rhizome growth (mm d-1 ± SE) Site 25% SI 5% SI Control Wreck Rock -0.83 ± 0.41 -13.3 ± 1.9 -2.8 ± 0.2 Success Bank -1.5 ± 0.48 -14 ± 2.5 -3.8 ± 0.1

41 a) 10 8 6 4 2 0 -2 -4 -6 -8

Mean net rhizome growth (mm ± SE) (mm growth net rhizome Mean -10 0 1020304050 Experimental period (days)

-20

-40

-60

Mean net rhizome growth (mm ± SE) growth (mm rhizome net Mean -80 0 1020304050 Experimental period (days)

Figure 2.6: Posidonia australis mean net rhizome growth (± SE) in free flowing seawater mesocosms subjected to reduced irradiance: a) 25% SI treatment, b) 5% SI treatment. Symbols are represented by: treatment ()) and control ("); a plot of in situ rhizome ---&--- growth is included for comparison. Rhizomes were collected from two sites, Wreck Rock (closed !) and Success Bank (open "), whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2).

42 Burial disturbance

Rhizome growth in the controls was significantly less than that observed in the field environment (t[6] = 346, p < 0.05, Table 3.3). Rhizomes subjected to a cycle of burial, uncovering and re-burial exhibited steady growth throughout the experimental period. Burial stimulated rhizome growth, which was 2 significantly greater than control rhizome growth (χ [14] < 0.001, p < 0.05; Figure 2.7a). Growth rates were similar throughout the experimental period, with growth decreasing (not significantly) during the 2 last 21 days (χ [14] = 1.18; p > 0.05; Figure 2.7a). No seedlings survived beyond the first burial period.

Table 2.3: Mean net rhizome growth rates (mm d-1 ± SE) for rhizomes from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2) treatment and control, with the treatment and control combined means, during each stage of the cyclical burial disturbance. Rhizomes were buried for a period of 21 days, then uncovered for 21 days and then buried again for a further 21 days.

Mean net rhizome growth (mm d-1 ± SE) Site Cyclical burial Control Wreck Rock treatment 16.85 ± 2.3 1.2 ± 0.1 Success Bank treatment 16.9 ± 3.6 1.3 ± 0.2

When buried for a prolonged period of time rhizome growth fluctuated (Figure 2.7b). Field rhizome growth was significantly greater than control rhizomes (t[11] = 2.46, p < 0.05). Rhizome growth rate was 2 not significantly different between treatment and control (χ [14] = 0.16, p > 0.05; see Figure 2.7b, Table 2.4). Seedlings became necrotic and died within seven days of beginning this treatment. Most seedlings (89%) had produced a small rhizome, however it did not grow appreciably or produce root hairs and fewer than 3 leaves were present on 98% of the seedlings.

Table 2.4: The mean net rhizome growth rates (mm d-1 ± SE) in complete burial disturbance and control treatments. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth.

Mean net rhizome growth (mm d-1 ± SE) Site Burial disturbance Control Wreck Rock 0.5 ± 0.8 -0.4 ± 0.1 Success Bank -0.2 ± 1.0 2.1 ± 0.3

43 40 35 30 25 20 15 10 5 0 -5 -10 Mean net rhizome growth rate (mm ± SE) (mm growth rate rhizome net Mean 0 10203040506070 Experimental period (days)

-5

-10

-15

-20

Mean net rhizome growth (mm ± SE) (mm growth rhizome net Mean -25 0 10203040506070 Experimental period (days)

Figure 2.7: Posidonia australis mean net rhizome growth (± SE) in free flowing seawater mesocosms subjected to: a) cyclical burial disturbance (days 0–21 and 43-63 are where rhizomes were buried, whilst days 22-42 is where rhizomes are uncovered); and b) complete burial disturbance, where shaded symbols represent Wreck Rock and open symbols represent Success bank. Symbols represented treatment ()) and control ("); a plot of in situ rhizome ---&--- growth is included for comparison. Rhizomes were collected from two sites, Wreck Rock and Success Bank, whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2).

44 Handling disturbance Mean net rhizome growth rates in the control were significantly less than rhizome growth rates in the field (t[20] = 15.3, p < 0.05; Figure 2.8). Mean net growth rates for treatments (frequently, infrequently and

not handled rhizomes) were statistically similar independent of the treatment (F[3,40] = 1.12; p > 0.05) or

initial site from which they were taken (i.e. Wreck Rock F[3,40] = 1.10; p > 0.05; or Success Bank;

F[3,40] = 0.29; p > 0.05; Table 2.5). Generally, net rhizome growth rates in the treatments (handled and infrequently handled) were significantly greater than the control (not handled) (F[1, 21] = 7.8, p < 0.05).

Frequently and infrequently handled seedlings died between day 48 and 63. Seedlings that were not handled survived throughout the experiment and averaged 14.4 cm in length, had roots and root hairs, produced two to three shoots and appeared healthy and robust.

Table 2.5: The mean net rhizome growth rates (mm d-1 ± SE) for treatments of frequently, infrequently and not handled (control). Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth.

Mean net rhizome growth (mm d-1 ± SE) Site Frequently handled Infrequently handled Not handled (control) Wreck Rock -3.1 ± 1.0 -9.4 ± 1.5 -3.5 ± 0.2 Success Bank -3.7 ± 1.1 -6.8 ± 1.0 -1.9 ± 0.1

45 60

-20

Mean net rhizome growth (mm ± SE) ± growth net rhizome (mm Mean -40 0 20 40 60 80 100 120 140 160 Experimental period (days)

Figure 2.8: The effect of handling disturbance on Posidonia australis mean net rhizome growth (± S.E.) in free flowing seawater mesocosms. Handling disturbance treatments: frequent handling ); infrequent handling "; and no handling 4. A plot of in situ rhizome growth ---&--- is included for a comparison of field and mesocosm rhizome growth. Mesocosm rhizomes were collected from two sites, Wreck Rock (closed!) and Success Bank (open"), Western Australia, whilst in situ rhizomes were measured on Success Bank (see Chapter 3, Figure 3.2).

Substrate type

Seagrass rhizomes growing in the sand treatment (control) grew at significantly different rates to seagrasses growing in the field (t[9] = 2.68, p < 0.05; Figure 2.9, Table 2.6). Net growth rates in sand and

limestone rubble treatments were similar (F[1,14] = 0.52, p > 0.05). Net growth rates fluctuated during the experimental period, with no influence of rhizome origin. A high percentage (74%) of seedlings survived on limestone rubble for 20 days. During this time, seedling rhizomes grew quickly, produced root hairs and emergent leaves much like the rhizomes in the shade and dark experiments. After 21 days, seedlings became necrotic and died. Rhizomes growing on limestone rubble produced adventitious roots. Seedlings growing on the control substrate appeared healthier, with more leaves, longer rhizomes and more roots.

46 Table 2.6: The mean net rhizome growth rates (mm d-1 ± SE) on limestone rubble and sand (control) substrates. Rhizomes were collected from Wreck Rock and Success Bank (see Chapter 3, Figure 3.2). Negative net growth rates indicate necrosis exceeded growth.

Mean net rhizome growth (mm d-1 ± SE) Site Limestone rubble Sand (control) Wreck Rock -1.1 ± 0.8 -1.5 ± 0.9 Success Bank 4.0 ± 1.5 -3.0 ± 1.6

-10

-20

-30

Mean net rhizome growth (mm ± SE) (mm growth net rhizome Mean -40 0 102030405060 Experimental period (days)

Figure 2.9: Posidonia australis mean rhizome growth (± SE) in free flowing seawater mesocosms on limestone rubble ) and a control substrate ". A plot of in situ rhizome ---&--- growth is included for a comparison between treatments and field growth. The rhizomes were collected from two sites, Wreck Rock (closed!) and Success Bank (open"), whilst in situ rhizomes were measured on Success Bank in Western Australia (see Chapter 3, Figure 3.2).

2.4 Discussion The work in this chapter was designed to examine four factors (reduced irradiance, burial disturbance, handling disturbance, and substrate type) that influence Posidonia australis seedling and rhizome growth, with the aim of determining criteria for selecting appropriate transplant sites for P. australis in Western

Australia. Four working hypotheses are derived from Chapter 1: HIa = light limits P. australis propagule growth; HIf = burial disturbance (as a correlate of water movement) limits P. australis propagule growth;

and HIh = the effect of substrate type on propagule growth (all for Stage I site selection); and also HIIc = handling disturbance limits P. australis propagule growth (Stage II transplant unit). The evaluation of these hypotheses will help develop criteria for selecting suitable sites for seagrass transplants (particularly P. australis).

2.4.1 Pilot studies The rinsing of rhizomes and seedlings in disinfectant solution, to kill bacteria and fungus, reduced initial mortality by 98%. Bacterial infection is often the cause of high mortality rates, through poor rhizome healing, within the first six months of experimental aquarium work (Meinesz et al. 1991; Meinesz et al. 1993). However, healed rhizome survival can be up to 95% (Meinesz et al. 1993) implying that when using seagrass seedlings or sprigs reducing bacterial and fungal infection is important for transplanting success.

The growth rates in the pilot study controls were not statistically different from field measurements. Growth in the pilot study, however, was significantly higher than the experimental controls. All subsequent experimental controls were significantly lower than field growth suggesting that the subsequent experiments were not representative of field conditions. Despite this, the experiments were self-controlled and consequently the influence of various factors on seagrass growth can still be assessed.

2.4.2 Manipulation experiments Light The consequences of reducing light levels to 25% and 5% SI on Posidonia australis growth has not been extensively examined. Light levels are an important limiting factor for seagrass growth (e.g. Bulthuis and Woelkerling 1981; Bulthuis 1983; Orth and Moore 1983; Dennison and Alberte 1985; Dennison 1987; Duarte 1991; Masini et al. 1990; Olesen and Sand-Jensen 1993; Zimmerman et al. 1994). To define and establish suitable sites to transplant P. australis requires accurate knowledge of the SI needed for this species to grow. Thus, the 25% and 5% SI experiments will determine how P. australis propagules grow in reduced light levels.

Reducing SI to 25% (102 μmol m-2 s-1; see Figure 2.6a) did not affect rhizome growth (growth in the treatment was similar to the control), irrespective of the site of rhizomes. Irradiance in the 25% SI treatment was higher than the reported saturation levels (73 μmol m-2 s-1) for Posidonia australis and hence, growth in these treatments should have been similar to the control unless photoinhibition occurred. In contrast, 5% SI (22 μmol m-2 s-1) is below the reported compensation levels (25 μmol m-2 s-1; Masini et al. 1990).

Seagrass metabolic rate versus light intensity follow the classic PI curve, where growth plateaus upon reaching saturation irradiance levels (Bulthuis 1983; Masini et al. 1990). Thus in Posidonia australis, growth increases, plateauing when light levels reach 73 µmol m-2s-1 (Masini et al. 1990). In this study, growth in the 25% SI treatment (102 µmol m-2s-1) and control (407 µmol m-2s-1) was similar because both these treatments had light levels above reported saturation levels. Yet, growth in these treatments was significantly lower than field rhizomes. In the field, light levels range between 321 µmol m-2 s-1 (during summer) to 144 µmol m-2s-1, which is higher than the treatment. If light levels in the field were lower than the mesocosms (where growth rates are lower) then the results may be indicative of photoinhibition.

48 As mentioned in the introduction, photoinhibition has not been satisfactorily documented in Australian seagrasses (Hillman et al. 1989) this study provides evidence suggesting that at these light levels photoinhibition did not occur.

Posidonia australis rhizomes placed in 5% SI irrespective of origin, had negative net growth (Figure 2.6b). The mortality observed was not just limited to the 5% SI treatment, it also occurred to a lesser extent in the control (Figure 2.6b). Statistical analysis showed no significant difference between treatment and control, although a decrease in necrosis occurred in the first 15 days in the 5% SI treatment rhizomes, which did not occur in the control rhizomes (Figure 2.6b). Based on statistical analyses the 5% SI treatment had no significant effect on rhizome growth.

Previous studies have shown that reducing in situ light levels results in altered meadow structure and productivity (Bulthuis 1983; Neverauskas 1988; Giesen et al. 1990; Dunton 1994; Gordon et al. 1994). Typically, declines in in situ light level result in reduced leaf and shoot density, and leaf length (Bulthuis 1983; Neverauskas 1988; Vermaat et al. 1993; Dunton 1994; Gordon et al. 1994). Changes in rhizome growth have not been measured, making comparisons with this present study difficult. Conversely, leaf and shoot density were not measured in this study because the stress of propagule removal from a meadow and placement into a mesocosm often resulted in a decrease in leaf and shoot density. Although, propagules were allowed a 30 day period (stabilisation period) to adjust to the mesocosm environment, it was difficult to distinguish between leaf and shoot density reductions bought about by removal stress or treatment (in this case light) stress.

The one unifying factor in these studies is that when SI levels are below a critical point, seagrass growth is affected and this critical SI level varies between species (Duarte 1991). For example, Heterozostera tasmanica remains relatively unaffected by reductions to 5% SI yet, once SI falls below this level growth is reduced (Bulthuis 1983). Similarly, Halodule wrigthii can tolerate a reduction to 18% SI (Dunton 1994), whilst both Posidonia oceanica and Cymodocea nodosa can tolerate ~10% SI (Drew and Jupp 1976), but below these thresholds, growth is reduced. Posidonia sinuosa is known to have reduced growth when SI is decreased between 80 and 99% (Gordon et al. 1994) and Zostera marina is reported to have reduced growth when SI is reduced by between 37 and 80% (Burkholder and Doheny 1968; Backman and Barilotti 1976). However the critical SI level is not documented or established for these and many other species. In the present study, 25% SI is above the critical value that reduces growth in P. australis rhizomes.

Reducing light to 5% SI (which is below reported irradiance compensation levels) should have caused a decline in Posidonia australis growth (Masini et al. 1990) and if the SI reduction occurred for a prolonged period of time necrosis and death would likely have ensued. Even moderate reductions in SI, if prolonged, can result in seagrass loss. For example, in Cockburn Sound SI reaching seagrass meadows averaged 78.3%, with excessive epiphyte shading, initiated by highly eutrophied water, reducing SI to 54%. This resulted in the decline of 3,300 hectares of mixed Posidonia meadows (Cambridge and

49 McComb 1984; Cambridge et al. 1986; Silberstein et al. 1986; Shepherd et al. 1989). Similarly, Posidonia meadows in Princess Royal and Oyster Harbours, Western Australia declined with declining SI due to excessive epiphyte loading caused by high eutrophication (Bastyan 1986; Mills 1987; Masini et al. 1990; Hillman et al. 1991). Increasing eutrophication, which reduces irradiance, resulting in seagrass die- off is common in many regions such as the United States (Sand-Jensen 1977; Orth and Moore 1983; Short and Wyllie-Echeverria 1996; Fonseca et al. 1996; Ruckelshaus and Hays 1998) and Europe (den Hartog and Polderman 1975; Giesen et al. 1990). Interestingly, nitrogen and phosphate are scarce in the Mediterranean (Drew 1978) and hence, seagrass species, such as P. oceanica, can exist in deeper water because light levels are higher (Dennison 1987; Duarte 1991) due to a lack of phytoplankton blooms as experienced in eutrophied waters. Dennison and Kirkman (1996) have used this fact to develop a seagrass survival model based on light attenuation as measured with a secchi disc.

Some species, such as Heterozostera tasmanica (Bulthuis 1983), maximise their light harvesting abilities and reduce respiratory demands to cope with reduced irradiance. These species have a decreased leaf/shoot density, a small rhizome and reduced leaf clustering, which helps to prevent self-shading and to maintain a positive carbon balance in low light conditions (Bulthuis 1983; Josselyn et al. 1986; West 1990). Posidonia species do not display these characteristics; they have a large rhizome, high leaf clustering and high leaf density, which reduces light harvesting and increases respiratory demands. These factors make Posidonia unsuitable to extremely low light environments, which is reflected in its Australian distribution and depth limitations (Bulthuis 1983; Coles et al. 1989; Kirkman and Walker 1989; Poiner et al. 1989; Shepherd et al. 1989; Walker 1989; West et al. 1989).

Posidonia australis seedlings in 25% SI survived for 15 days, while the control seedlings survived until the 22nd day. Some seedling rhizomes (control and treatment growth was similar) grew 0.4 mm d-1. The early mortality of the treatment seedlings may have been due to the reduced light, however irradiance was above compensation levels, hence it seems unlikely that 25% SI caused the mortality. Also, as discussed in the burial treatment, seedlings have stored carbohydrates that should allow survival for at least nine months (Hocking et al. 1980, 1981). Typically, this nine month period is long enough for seedlings to establish photosynthetic processes and replenish carbohydrate stores (Hocking et al. 1980, 1981) and thus mortality would not be expected from reductions in light within the first nine months after settling because carbohydrates stores are available. Moreover, it appears that handling may have affected the seedling survival in this experiment because control seedlings also perished within a short period of time. This is not unexpected, as the handling disturbance treatments indicate that handling may cause the demise of seedlings.

In this study 5% SI caused necrosis in Posidonia australis rhizomes, but statistically necrosis levels were similar to those occurring in the control. Posidonia growth in the field was statistically similar to the 5% SI and control. This is statistically accurate, however the trend was that 5% SI and the control caused necrosis, while the field rhizomes grew: these results are diametrically opposed. The drastic level of necrosis up till day 15 may have affected the statistical analysis, yet, in all three replicate experiments the

50 same pattern occurred: necrosis was high in the first 15 days of the treatment and it continued but to a lesser extent after 15 days (Figure 2.6b). This result suggests that 5% SI is below the critical SI value, yet it is not substantiated because similar levels of necrosis occurred in the control.

Seedlings were relatively unaffected by the 5% SI treatment. This is possibly due to stored carbohydrates upon which they can draw to survive (Hocking et al. 1980, 1981). In contrast to previous treatments (disturbance, burial and 25% SI), the constant measurement of seedling growth appeared not to affect survival. Why this occurred is questionable and suggests that reducing SI to 5% has less effect on seedlings than burial (which will scour and smother) and reductions to 25% SI.

Thus, it appears that 5% SI produced necrosis in Posidonia australis rhizomes, though controls also suffered from necrosis. Posidonia australis seedlings coped well with the decline in light irradiance, suggesting that they are a suitable transplant unit in light environments that are below compensation levels. Reducing SI to 25% did not affect P. australis growth, and reductions to 5% SI appeared to cause necrosis, but is inconclusive. Previous studies suggest that moderate reductions in irradiance, if prolonged, will affect growth. From this work, it is not possible to state whether P. australis needs SI levels between 5% and 25%, yet this may be true. Perhaps, instead P. australis is more affected by prolonged reductions in SI. Further studies are required to determine the SI levels required for P. australis to grow and hence, determine transplant sites.

Burial disturbance The marine environment along the Perth metropolitan coastline is an accreting system (Paling et al. 1997) and burial events occur frequently (Hancock 1992; Nelson 1992; Walker 1994). Burial of seagrasses has the potential to affect growth and survival and hence was examined to determine the influence such factors have on Posidonia australis propagules (rhizomes and seedling). This in turn will help determine suitable site and propagule choice when attempting to transplant P. australis.

When burial disturbance was cyclical, rhizome growth was stimulated (Figure 2.7a). However, when burial was constant, growth of Posidonia australis rhizomes was not affected (Figure 2.7b). In the constant burial experiment, net rhizome growth rates fluctuated, with rates being similar to field and control conditions (Figure 2.7b).

Stimulation of growth in response to cyclical burial was expected, largely because many seagrass species respond to burial by increasing vertical shoot (orthotropic) growth (Patriquin 1973; Boudouresque et al. 1984; Meinesz et al. 1992; Gallegos et al. 1993; Molenaar et al. 1993; Marbá 1994a, 1994b; Marbá and Duarte 1994; Duarte et al. 1997). Cymodocea nodosa seedlings tolerate burial < 7cm, with both its seedlings and rhizomes reacting to burial by increasing vertical growth (Marbá 1994a, 1994b; Marbá and Duarte 1994). Similarly, fast growing species like C. rotundata, Enhalus acoroides, Syringodium isoetifolium and Thalassia hemprichii all significantly increase vertical growth when buried, however this response is species-specific (Duarte et al. 1997).

Posidonia australis did not respond to constant burial, instead it appeared to tolerate it (Figure 2.7b). Some dominant, slower-growing species, such as Phyllospadix scouleri and Posidonia oceanica can withstand burial because of their rhizomatous root system (Littler et al. 1983; Molenaar et al. 1993). Posidonia oceanica tolerates burial because its orthotropic rhizomes are able to reach the surface and once at the surface they can alter their growth form to become plagiotropic (Molenaar et al. 1993). This allows the rhizomes to reach the light and relocate the meristem to the surface. It is important to note that plagiotropic rhizomes cannot alter their growth form and become orthotropic (Molenaar et al. 1993), hence are unable to relocate the meristem to the surface when buried. Furthermore, dominant, slow- growing species, such as Cymodocea serrulata (Duarte et al. 1997) and Zostera species (den Hartog 1994), cannot respond rapidly to burial and hence, cannot survive prolonged burial. “Slower growing” species are not opportunistic relative to burial or disturbance; they can be proximally stress-tolerant (Littler et al. 1983) but if burial is prolonged they may succumb (den Hartog 1994; Duarte et al. 1997). This seems to be the case for P. australis in this experiment, which increased growth to similar levels as observed in the field environment when burial was cyclical, and tolerated burial when it was constant (Figure 2.7). Generally, in multispecific seagrass meadows burial creates an environment where some dominant species die and are replaced by faster growing, opportunistic species that are capable of altering and increasing growth to withstand burial (Kirkman and Walker 1989; Duarte et al. 1997).

Posidonia australis seedlings were susceptible to burial, with seedlings subjected to cyclical burial and erosion dying by the 21st day of the experimental period and seedlings subjected to constant burial dying within 7 days. Reduced light levels should not affect Posidonia seedlings as they have enough stored carbohydrates to survive on average for nine months (Hocking et al. 1980, 1981). If, within nine months, a seedling had not grown leaves, then mortality is expected, because the seedling would be unable to replenish its used carbohydrate stores. The experimental period was much less than nine months and hence it is unlikely that carbohydrate stores had been depleted. Instead, it seems likely that handling and abrasion of seedlings, during measurement, may have detrimentally affected seedling survival. This deduction is supported by the results from the handling disturbance experiment discussed below.

Thus, Posidonia australis growth was stimulated when burial was cyclical, but growth was unaffected by constant burial, indicating that this species is stress tolerant. Conversely, P. australis seedlings were susceptible and did not cope with burial and died within a short period of time after burial. Therefore, P. australis rhizomes are more resistant to burial than P. australis seedlings.

Handling disturbance Much evidence supports the concept that seagrasses are displaced by natural (see Patriquin 1975; den Hartog 1987; Robblee et al. 1991) and anthropogenic disturbance (see Orth and Moore 1983; Bulthuis 1983; West 1983; Cambridge and McComb 1984; Bastyan 1986; Cambridge et al. 1986; King and Hodgson 1986; Silberstein et al. 1986; Shepherd et al. 1989; Short and Wyllie-Echeverria 1996; Walker and McComb 1992; Ruckelshaus and Hays 1998). For example, Zostera marina in Europe and Thalassia

52 testudinum in Florida have suffered large reductions in cover bought about by the ‘wasting disease’ and pathogens (den Hartog 1987; Robblee et al. 1991). Similarly, direct disturbance of meadows via dredging and reclamation (Short and Wyllie-Echeverria 1996; Ruckelshaus and Hays 1998) has reduced coverage in the Florida Keys and Australia (Larkum and West 1990). Indirect disturbance through eutrophication (West 1983; Bastyan 1986; King and Hodgson 1986), such as in Cockburn Sound, Western Australia (Cambridge and McComb 1984; Cambridge et al. 1986; Silberstein et al. 1986), has also reduced meadow cover.

The very act of transplanting seagrasses however can cause a high level of disturbance, placing stress on transplant propagules. Because of this stress, the affect of disturbance on Posidonia australis propagules was examined to determine how propagules cope. If it is known how P. australis propagules react to disturbance then appropriate measures can be taken to alleviate some of this stress when attempting a transplantation effort in order to measure the likelihood of success.

Handling was found to be detrimental to the survival of seedlings, with control (no handling) seedlings surviving for a significantly longer period than either the frequently or infrequently handled seedlings. Seedlings are more delicate than rhizomes and no studies have empirically examined handling disturbance as a cause of the failure of seedling transplant efforts, although some have hypothesised that handling reduced the fitness of the seedling (Hancock 1992). The temporary exposure of rhizomes, during measurement, has also been noted as causing the deterioration of plants (Clarke and Kirkman 1989; D. Walker pers. comm.) Three rationales to why handling affects seedling survival exist. First, disruption of the rhizosphere leads to reduced nutrient uptake. Second, disruption of the rhizosphere leads to an increased occurrence of bacterial and fungal infection. Finally, handling easily damages seedlings because they are small and delicate. All three rationales have merit and are possible, but all result in the reduction of the potential for seedlings to be effective transplant units. Regardless of handling disturbance level, rhizome growth was poor (Figure 2.8). The control (no handling) treatment growth was significantly less than handling disturbance treatments, inferring that handling disturbance may have increased rhizome growth (Figure 2.8). This suggests that seagrass rhizomes respond to disturbance by increasing growth, as seen in the cyclical burial treatment discussed above. However, it is more plausible that the increase in growth is an experimental artifact and thus, handling should be minimised in both seedling and rhizome manipulations.

Comparisons with studies on the effects of artificial handling or disturbance of transplant units and propagules are impossible to make because such studies are rare in the literature. Although, based on bioturbation studies, inferences can be made that animals are capable of disrupting the seagrasses rhizosphere (see Merkel 1990a), however this activity does not detrimentally affect all species. In fact, some species are more resistant to bioturbation (Valentine et al. 1994). Thus, perhaps parallels can be drawn to the affect of handling, with some species being more resistant to handling than others.

53 Generally, it is assumed that transplant stress exists because many transplants die, however the measurement and quantification of such stress has not occurred. Hancock (1992) indicated that handling of propagules decreased seagrass transplant survival. Other researchers have also made statements that handling decreased the success of transplant attempts (Clarke and Kirkman 1989; LeProvost Environmental Consultants 1990; Kirkman 1995; Walker 1994; D. Walker cited in Hancock 1992). This study has not been able to confirm a detrimental influence from handling Posidonia australis rhizomes. It is assumed that handling would damage rhizomes and seedlings and disrupt the rhizosphere, hence influencing experimental results. However, the handling in this study illustrated that rhizomes are robust and unaffected by weekly or fortnightly handling (in fact they benefited from this disturbance), whilst seedlings deteriorated when handled. This infers that P. australis rhizomes are more suitable as transplant units, however such conclusions are tentative and need further experimental quantification (Chapter 3).

Substrate type The final evaluated factor that has the potential to effect seagrass growth was substrate type. Two substrates that are common in the Perth coastal region were examined: sand and limestone rubble. The growth of Posidonia australis propagules on both these substrates was investigated to determine if there is a ‘better’ substrate choice when transplanting P. australis.

The growth of Posidonia australis rhizomes in both limestone and sand treatments fluctuated irrespective of the site the rhizomes originated from and were not significantly different (Figure 2.9) and did not differ to field rhizomes. The implication from the observed similarity between substrate type and field rhizome growth is that P. australis has no preference for substrate type: P. australis grows sufficiently well on both sand and limestone substrate.

Posidonia australis seedlings grew in both sand and on limestone rubble, with no statistical difference occurring between substrate type. This is seen in the field environment, where P. australis seedlings are commonly found on reef top and on limestone rubble (e.g. Woodman Point, pers. obsv.). However, seedlings growing on hard substrate, such as limestone rubble, do not produce a taproot and are unable to sufficiently anchor themselves in the substrate, making them susceptible to dislodgment by currents and foraging fauna. Thus, although it occurs, seedlings gain no benefit from growing on hard substrates and are more likely to be displaced from such environments.

Posidonia australis propagules were able to grow on both sand and limestone rubble, with neither substrate conferring a distinct advantage to rhizomes or seedlings, although sand substrates may present an advantage, because seedlings can more easily anchor and remain anchored in sand.

2.4.3 Variation between sites Interestingly, the examined sites (Wreck Rock and Success Bank), although many nautical miles apart, appear to have Posidonia australis plants that reacted to the tested treatments in a very similar fashion. Variation between sites was expected based on the work Waycott (1995) and Waycott and Les (1996),

54 which indicate that the genetic structure of P. australis populations over a large geographic range (south coast of Western Australia, to Rottnest Island and the Perth metropolitan area) is similar. However, these populations display either r- or K-selected characteristics, even over small spatial scales (populations on Rottnest Island), which may be a reflection of abiotic factors at the sites. Abiotic processes include climate, mineral resources, mortality – causing disturbances and geological processes (M. Huston pers. comm.). Variations in productivity are a product of site index, rate of succession, disturbance and mortality and interactions (dynamic equilibrium between disturbance and recovery). The lack of variability between Wreck Rock and Success Bank suggests that Posidonia propagules from these sites are under similar abiotic influences and hence, react to the test treatments in a similar fashion. Further in situ measurements between these two sites may provide evidence for this.

2.4.4 Mesocosms as a measuring tool Mesocosms are a useful tool if the right questions are being asked (Daehler and Strong 1996; Lawton 1996). The experiments in this study were repeated and replicated with results illustrating that although the mesocosm results were significantly different to field measurements, the experiments were self- controlled allowing the assessment of the influence of light, burial disturbance, handling disturbance and substrate type on seagrass growth. Although mesocosms can be useful, they are a tool to supplement rather than replace field observations and experiments, and analyses of long-term databases in the search for large-scale patterns and processes in seagrass ecology.

Mesocosms are thought to be useful because they allow the study of community dynamics by empirical analysis, which otherwise may take decades for sufficient field studies (Daehler and Strong 1996). However, scientific opinions on the degree of mesocosm usefulness vary widely, with some advocating putting more effort towards large-scale field manipulations (Carpenter 1996). Lawton (1996) and Carpenter (1996) provide good discussion on the pro’s and con’s (respectively) of mesocosm use. The criticisms include: that they are highly artificial and too simple for comparisons with the real world; the community within the mesocosms is an unnatural assemblage; mesocosm experiments are closed to immigration and emigration; the communities are not in equilibrium; there are no seasons, density- independent disturbances or major environmental perturbations in a mesocosm; inappropriate spatial and temporal scales; and they are expensive to build, operate and maintain. Lawton (1996), Carpenter (1996) and Drake et al. (1996) discuss a number of advantages to mesocosm studies, such as the degree of control and replication mesocosms afford, which is impossible in the field; the speed of such experiments; their statistical power; the ability to parse ecological processes; and how they enable a precise repetition of experiments.

Although mesocosms offer control and replication, often by design the replicates are physically interdependent (Hurlbert 1984) and hence pseudoreplication occurs. Independence of replicates was also compromised by the repeated measure of rhizome elongation over time (temporal pseudoreplication). However, measuring rhizomes over time is desirable because it increases the sensitivity of an experiment (Hurlbert 1984). To account for this a repeated-measures experimental design and analysis was used. A

55 disadvantage to this design is ‘carryover’ effects of the previous measurement if the time between measurements is insufficient. Rhizomes were measured at a number of different time frames to account for possible carryover effects and the results showed no statistical difference between time frames. Therefore, it seems improbable that a carryover effect occurred in this work. Statistical Type II error’s (concluding that the treatment had no effect when in reality it did) are also common in repeated-measures analyses (Glantz 1992; Zar 1996). However, since this increases the degree of conservatism it is considered preferable to analyses that encourage Type I error (concluding that the treatment had an effect when in reality it did not).

What is certain from these results is that Posidonia australis rhizomes are more robust to the manipulations than seedlings, suggesting that rhizomes are a better transplant unit (more capable of withstanding handling/disturbance).

2.4.5 Conclusions Site selection is an important process in a seagrass transplantation attempts. A review of literature suggested that light levels, burial and handling disturbance, and substrate type are all factors that had the potential to influence the success of seagrass transplantation and needed to be considered when selecting a potential transplant site. Hence, a set of manipulation experiments was designed to test these factors, with the purpose of determining protocols that could be used when attempting to transplant Posidonia australis in Western Australia.

Posidonia australis seedlings were seemingly poor propagules, succumbing to reduced light, burial and handling disturbance. Seedlings were also unlikely to thrive on hard substrate because anchoring is difficult in such conditions. Thus, vegetative propagules appear to be a better choice of transplant unit.

Based on the present information and a review of pertinent literature, a prediction for a suitable area for Posidonia australis transplants can be made. The Stage I (site selection) criteria are: 1) Stage I (site selection) a) Irradiance levels above 5% SI; b) Water quality needs to be high and not eutrophied; c) Short burial periods with cyclical burial periods preferred, which infers that; d) Water movement should be low or reduced; and e) Substrate type can either be limestone rubble or coarse sand, with sand a better choice due to ease of transplanting and anchoring ability. 2) Stage II (transplant unit and technique) a) Handling should be minimised.

The need for identifying potential mitigation sites exists for the region south of Fremantle due to seagrass loss by sea sand-mining. Three sites have been identified: Carnac Island, Woodman Point and Success Bank. These three sites are compared with the Site Selection Criteria in Table 2.7. Carnac Island is

56 subjected to prolonged burial periods, a high-energy regime (Nelson 1992; Walker 1994) and has high surface irradiance. The prolonged burial periods and high energy regime makes it unsuitable. Woodman Point is inadequate because it receives highly eutrophied water from Cockburn Sound (Nelson 1992) resulting in high epiphyte loads that reduce available light. Although Woodman Point has a coarse, rubble substrate suitable for transplantation, its suitability is reduced because it is near a dredge spoil area at the Cockburn Cement jetty which has the potential to reduce SI to below 5%. Success Bank has a range of substrates (coarse to fine sand), a range of available light as the Bank goes from 5 m to 12 m depth, is in a moderate to low energy regime (Nelson 1992) and the occurrence of long term burial has not been seen or recorded. Success Bank appears to be the best choice for seagrass transplantation in the regions examined.

Table 2.7: Evaluation of Carnac Island, Woodman Point and Success Bank by the Stage I Criteria listed in the text. ! denotes preferred choice.

Stage I Criteria Carnac Island Woodman Point Success Bank 1) Light > 5% SI high low high 2) Water quality fair low fair 3) Burial period long mid-high nil 4) Water movement high low mid-low 5) Substrate type sand limestone rubble sand

Chapter 3 Posidonia australis propagation: selection of transplant unit and technique for restoration

3.1 Introduction A prime consideration for transplant (and therefore mitigation) success is how seagrasses maintain and persist in existing meadows and disperse to colonise new habitats. In this regard the r- and K-selection strategies described by ecological theory (Krebs 1985), can be readily applied to seagrasses. Those considered to be colonising species, such as Heterozostera, Halophila and Syringodium, possess features typical of r-strategists, such as low biomass, high shoot density, fast growing rhizomes and are short lived. Whereas ‘climax’ species such as Posidonia and Amphibolis may be considered to be K-selected and have a high biomass, low shoot density, slow rhizome growth and are long lived. It is generally considered that colonisation success may be based upon r- and K-selected life strategies (Krebs 1985). However, individual life history traits and the characteristics of the recipient community need consideration.

3.1.1 Seagrass reproductive strategies Seagrasses maintain populations and disperse via sexual (seeds) and/or vegetative propagules. Like many other plants, seagrasses are heterogonic and hence capable of using both forms of propagation (Ramage and Schiel 1998). The degree to which each propagule type is employed appears to be species, and to a lesser extent, genus specific. The various sexual and vegetative reproduction methods are considered below in order to place Posidonia australis reproduction into context. The aim is to identify the most effective transplant unit type (sexual or vegetative) and to extend this knowledge to seagrass restoration research.

Sexual reproduction In general, the evolutionary effect of sexual propagation is to accelerate the rate of evolution, decrease extinctions and purge deleterious mutations from the lineage more efficiently (Ridley 1996). It provides a mechanism for dispersal to new habitats and is frequently more successful when conditions are unstable (Krebs 1985; Ridley 1996). Factors important in sexual reproduction include pollination, seed viability, seed production, seed dormancy and fruit/seed dispersal. Pollination has been well studied elsewhere (e.g. McConchie and Knox 1989) with the primary findings indicating that water borne (hydrophillous) pollen dispersal is limited and results in low fertilisation success. Very little study has been carried out on seed viability, seed production or fruit/seed dispersal. Dormancy, however, has been thoroughly studied (McComb et al. 1981; McMillan 1986; Kuo and McComb 1989; Jewett-Smith and McMillan 1990; Kuo et al. 1990). Seed viability and germination of some seagrass species (e.g. Zostera marina) is considered to rarely play an important part in this process, unlike freshwater plants (Rea and Ganf 1994). Instead, the recruitment of propagules is believed to limit colonisation events (Harrison 1993). Success is therefore influenced by dispersal potential, disease, predation, competition (for space, light and nutrients), local population size (density dependence) and disturbance (Brown and Venable 1986; Duarte 1991; Harrison 1993; Orth et al. 1994; Rees 1997; Westoby et al. 1997).

Seagrass sexual propagules are either viviparous (Amphibolis, Thalassodendron and some Phyllospadix species) or oviparous. A number of seagrass species have dormant seeds (for examples see Caye and Meinesz 1985; Pirc et al. 1986; Duarte and Sand-Jensen 1990b; Churchill 1992; Harrison 1993; Moore et al. 1993; Conacher et al. 1994b), which enter a seed bank and await germination cues, such as seasonal changes in water temperatures and low oxygen levels, due to the deposition of sediment onto the seeds creating an anaerobic microclimate. For example, both Cymodocea nodosa and Z. capricorni require an anaerobic microclimate to germinate (Pirc et al. 1986; Conacher et al. 1994b). Dormant seeds delay reproduction, which may benefit the plant (by avoidance of unsuitable conditions), however it may also disadvantage the plant because there is an associated cost to delaying reproduction (i.e. a plant may die before it is able to reproduce). Species without a seed bank (e.g. Posidonia australis) rely on appropriate conditions at the time of release for their propagules to germinate and establish.

Vegetative reproduction Vegetative propagation, an alternative to sexual reproduction, may have evolved in response to evolutionary pressures favouring uniform populations (Ridley 1996). Many plants use vegetative propagation when conditions are stable (Krebs 1985; Ridley 1996; Ramage and Schiel 1998) however it can restrict the ability of a population to adapt to varying conditions and dispersal capabilities are limited (Orth and Moore 1994; Ridley 1996). The benefit to propagating vegetatively is that a plant can invade areas nearby, which are inherently suitable, because the conditions are similar. Another advantage of vegetative propagation is that achieving pollination in an aquatic environment is a high risk activity and a clonal growth form removes this risk (Duarte et al. 1994; van Groenendael et al. 1997).

In seagrasses, rhizomes are the vegetative structures capable of regenerating new plants when detached from the parent plant (through natural or artificial means). This form of reproduction is meristem dependent. When a meristem is present, a vegetative propagule has the ability to take root and colonise an area (Tomlinson 1974). Olesen and Sand-Jensen (1994) have observed that pieces of floating shoots do not often form meadows or spreading patches because their rhizomes lack apical meristems and hence cannot grow. The development of photosynthetic leaf-bearing shoots on each rhizome node is also required for successful growth (Raven et al. 1986; Olesen and Sand-Jensen 1994).

The recruitment of a vegetative propagule alone will not ensure meadow formation but must be coupled with rhizome elongation (Duarte and Sand-Jensen 1990). In Cymodocea nodosa and Posidonia oceanica patch formation is a function of both rhizome elongation (which determines patch growth), seed numbers, and seed longevity which determines the potential for formation of new patches (Duarte and Sand-Jensen 1990a, 1990b). Although, a species that has a slow elongation rate, such as Zostera marina, can still be a colonising species because it has a high sexual reproductive effort (Duarte and Sand-Jensen 1990). Successful vegetative reproduction may be defined therefore as a function of rhizome elongation rates and the number of viable vegetative propagules (i.e. those that have an apical meristem).

Seagrass meadows maintained vegetatively usually occupy stable habitats such as deeper sublittoral regions but will occasionally invade shallow, less stable regions via sexual propagation (Larkum and den Hartog 1989). Genetic evidence from Posidonia meadows in Western Australia substantiates this observation and supports the notion that populations are maintained sexually (i.e. they have a high genetic diversity). Williams and Davis (1996) have demonstrated that seagrasses transplanted as vegetative propagules lack genetic diversity and curiously, fail to reproduce sexually, even when collected from shallow environments.

3.1.2 Dispersal mechanisms of seagrasses Sexual propagules Propagule dispersal is necessary for patch formation and subsequent meadow expansion. Water currents are the vector for seagrass propagules providing a mechanism for out-crossing, thus preventing inbreeding depression (Fonseca and Kenworthy 1987; McConchie and Knox 1989; van Groenendael et al. 1997). If the current is sufficiently strong, propagules are suspended in the water column and transported away from their release site (Anderson and Charters 1982; Fonseca and Kenworthy 1987). Although conditions in an area are often suited to seed dispersal, more dense seeds remain close to parent meadows (Orth et al. 1994). For example, Zostera seeds are dense and when released, they settle quickly near the parent meadow and become rapidly incorporated in the sediment, forming a seed bank (Orth et al. 1994).

In contrast, Amphibolis seedlings are viviparous and form leaves before detaching from the parent. The leaves act to increase drag and aid in dispersal; the plant also has a ‘grappling hook’ apparatus to aid in substrate attachment (McConchie and Knox 1989). Amphibolis species appear to be capable of dispersing far from parent meadows (tens to hundreds of metres) but genetic evidence does not support this (Waycott 1995). Instead, some Amphibolis species (e.g. Amphibolis antarctica) may be clonal, or highly inbred, although it is a dioecious species (Waycott 1995; Waycott and Les 1996). In Australia, Amphibolis seedlings are released in winter and become attached to a substrate by late winter - early spring (Ducker et al. 1977; D. Walker pers. comm.). Seedlings often settle at the edges of Posidonia meadows where the rhizome matte is exposed. This type of anchoring is often reflected by the presence of Amphibolis plants on the lee side of many Posidonia meadows (Nelson 1992).

Posidonia fruits are also positively buoyant allowing a wide distribution (McConchie and Knox 1989). Waycott (1995) found that Posidonia meadows are genetically diverse, suggesting that either sexual reproduction is occurring and fruits disperse far from their parent meadows or vegetative propagules disperse and establish in meadows. No evidence exists at present to support the hypothesis that vegetative propagules disperse and establish (Kuo and Kirkman 1987; Kirkman and Kuo 1990).

Life history strategies associated with long and short dispersal have pros and cons. Long distance dispersal away from parent stock enhances out-crossing thus ensuring genetic diversity, yet significant wastage can occur. Currents can deposit fruits or seeds in unsuitable habitats, such as deep water or beaches) preventing successful establishment. Short distance dispersal (such as immediate settling of heavy seeds) ensures a more suitable environment for seagrass growth, however seeds that settle in a

meadow are often unable to penetrate the thick rhizome matte and therefore remain loosely attached and easily dislodged (pers. obs.). In addition short distance dispersal results in increased inbreeding and hence lower local genetic diversity.

Vegetative propagation Vegetative propagation and dispersal can also result in long and short distance jumps. Long distance dispersal of vegetative propagules is controlled by water movement. Typically, a vegetative portion is dislodged in rough conditions (e.g. a storm event) and dispersed by wave action or currents to an area where a barrier (such as a meadow edge, a blowout wall, a reef or channel) allows settlement. If the vegetative portion has an apical meristem then the recruit may grow and spread much like sexual propagules. Long distance dispersal benefits include increasing local genetic diversity.

In contrast, vegetative propagation may occur by seagrass ‘runners’ forming ramets away from the parent. These ramets can be as far as tens of metres from the parent plant. Ramets are typically connected to the parent plant by rhizomial runners, yet these can breakaway without being dislodged from the substrate. For example, Halophila ovalis produces long ‘runner-like’ rhizomes, which can be broken or excised leaving both the parent plant and the ‘new’ vegetative propagule to continue growing (pers. obs.).

3.1.3 Life history strategies Ramage and Schiel (1998) state that the life history strategies of seagrasses are related to stable and unstable environments, resulting in annual and perennial life histories. Annual seagrasses occupy unstable habitats, represented by fluctuating salinities, high temperatures and/or high sediment movement. Growth typically begins in spring, continues through summer and declines in autumn. During winter, rhizomes may either die-off, leaving seeds in the seed bank to re-establish the meadow the following spring (Harrison 1979; Ramage and Schiel 1998; K. Benjamin pers. comm.) or persist through the winter and set seed prior to die-off. The rhizomes of Heterozostera tasmanica die during the winter rains and decreased salinity in the Swan River estuary, and rely on a seed bank to re-establish in spring. Similarly, when Zostera novazelandica grows in high water temperature and is exposed to large salinity fluctuations, it exhibits an annual life history that relies on recruitment of germinating seeds in spring (Ramage and Schiel 1998).

Typically, perennial seagrasses occupy stable environments (Tomlinson 1974; Phillips et al. 1983; Robertson and Mann 1984; van Lent and Verschuure 1994a, 1994b, 1995). They grow in summer, allocating their biomass to rhizome elongation. During winter, the leaves may decline but the rhizome persists (Tomlinson 1974; Phillips et al. 1983; Robertson and Mann 1984; van Lent and Verschuure 1994a, 1994b, 1995). Seeds are considered to be less important than vegetative growth in the year-to-year survival of populations of these species (Tomlinson 1974).

Only two genera, Halophila and Zostera, have species that exhibit annual life history strategies, these species are not obligate annuals but reflect environment conditions (Keddy and Patriquin 1978; Jacobs

1982; Phillips et al. 1983; Clarke and Kirkman 1989; van Lent and Verschuure 1995). The vast majority of seagrass species are perennial. In the Netherlands, for example, Z. marina has both annual and perennial populations, with germination, morphology and flowering being a genotypic response (Olesen and Sand-Jensen 1994; van Lent and Verschuure 1995).

3.1.4 Posidonia colonisation and meadow expansion Posidonia australis is a perennial seagrass, found in temperate waters from Western Australia to New South Wales (Figure 3.1), which exhibits both vegetative and sexual propagation. It begins to flower in April when day length shortens and water temperatures decrease (Hocking et al. 1980; Kuo and Kirkman 1990). Fruits take 12 weeks to reach maturity from anthesis. The fruits mature in spring when water temperatures and light increase (Hocking et al. 1980; Kuo and Kirkman 1990) and are released in early to mid-summer (November – December; Cambridge 1975). The fruits are positively buoyant and are dispersed by water currents and winds (McConchie and Knox 1989). Within one to two days of release, dehiscence of the pericarp occurs and the negatively buoyant seedling is released. Previous work notes that P. australis and P. sinuosa have 6.50 ± 0.61 and 8.26 ± 0.93 viable fruits per flowering shoot, respectively (Hocking et al. 1980), although these values may be underestimates. Hocking et al. (1980) assumed the viability of the first two fruits produced on a flowering shoot was representative of all fruits on the shoot.

Figure 3.1: The Australian distribution of Posidonia australis (Campbell unpub. data).

It is difficult to determine when germination occurs in Posidonia species because development is continuous and no dormancy period has been observed (Cambridge 1975; McComb et al. 1981; A. McComb pers. comm.). Hence, it can be assumed that germination occurs before the seed falls from the fruit (McComb et al. 1981; A. McComb pers. comm.). The seedling is negatively buoyant and settles soon after dehiscence from the pericarp (Hocking et al. 1980; Kuo and McComb 1989). These seedlings

have been found on bare sand, within established meadows, surrounding meadow edges and on reef tops (Cambridge 1975).

A general perception has existed that successful recruitment (or establishment) is low despite regular sexual reproduction in Australian Posidonia species (Kirkman 1989; West et al. 1990). While this is apparent from observations of unsuccessful seed colonisation over time, it contrasts with the observation of high genetic diversity in Posidonia meadows (Waycott 1995). One possible explanation for this is that successful seedling recruitment is cryptic and occurs in established meadows, but this does not explain initial meadow establishment.

Information on vegetative reproduction in Australian Posidonia species is scarce, but Kuo and Kirkman (1987) and Kirkman and Kuo (1990) consider that Australian Posidonia species cannot colonise from vegetative propagules to establish meadows. They state that the growth rate of Australian Posidonia species is too low to result in patch growth. However, other observations suggest that rhizome growth rates are sufficient for spread and the colonisation of new territory to form new meadows (Bastyan 1986; G. Bastyan pers. comm.; Paling 1992, 1995). Similarly, storm debris has been observed taking root and extending (see section 3.2.6.). This implies that P. australis sustains meadows and colonises new habitat using vegetative growth.

In contrast to Australian Posidonia, the Mediterranean P. oceanica, flowers irregularly and infrequently, limiting sexual reproduction. It is considered therefore, that colonisation is primarily by vegetative reproduction (Giraud 1977; Thelin and Boudouresque 1985; Mazzella et al. 1983, 1984; Caye and Meinesz 1984; Pergent et al. 1989; Meinesz et al. 1993). Posidonia oceanica spreads by horizontal growth of the rhizome and can reproduce vegetatively by two methods: either the rhizome separates from the parent stock by necrosis of the proximal section or a distal section tears away and is transported by water movement (Meinesz et al. 1992). The first method results in local colonisation while the second may result in dispersal (Meinesz et al. 1992). Based upon these observations, the favoured transplant method is vegetative plagiotropic cuttings with at least two leaf bundles (Meinesz et al. 1992; Molenaar and Meinesz 1992; Meinesz et al. 1993; Molenaar et al. 1993).

3.1.5 Aims The current knowledge of Posidonia reproduction and spread is poor with confusing and often contradictory perceptions. Further experimental study is required to determine the roles of sexual and vegetative propagation in P. australis. This chapter aims to examine sexual and vegetative propagation in the seagrass P. australis, with the proximal purpose of establishing how P. australis maintains meadows and extends into new territory. Ultimately this will identify which P. australis propagule type is suitable for establishing and maintaining a P. australis population (or allowing the expansion) and is therefore suitable for use in seagrass transplantation projects. As stated in Chapter 1, my working hypothesis is HIIa = vegetative propagules are better transplant units than sexual propagules.

3.2 Materials and methods 3.2.1 Site descriptions Sites along the coast of Western Australia were selected to represent both ‘pristine’ and ‘eutrophied’ habitats with varying environmental conditions (such as depth, sediment type and infauna). These sites contained a similar suite of seagrass species and had similar commercial and recreational uses and included: 1) Carnac Island; 2) Success Bank; 3) Woodman Point; 4) Wreck Rock, in the Marmion Marine Park; 5) Hamelin Bay, Peak Island, in south Western Australia; 6) Hamelin Bay, Cosy Corner; and 7) Cockburn Sound (Figure 3.2). Tidal movements at the sites were negligible. Differences between sites were used to detect variability in meadow shoot and fruit density, biomass and rhizome elongation and to guide site selection in the seagrass transplantation and restoration experiments (Chapter 4).

Three meadows, Carnac Island (1), Success Bank (2) and Woodman Point (3), were selected because they are considered to be potential seagrass mitigation sites (Cockburn Cement Ltd pers. comm.). Furthermore, due to previous studies, comparative information is available on these sites (LeProvost Environmental Consultants 1990; Hancock 1992; Nelson 1992; Walker 1994; Paling et al. 1997). Cockburn Sound (7) is in close proximity to these potential seagrass mitigation sites (1, 2 and 3) and has been extensively studied (Cambridge 1979; Cambridge 1980; Cambridge and McComb 1984; Cambridge et al. 1986; Scott et al. 1986). These four sites experience multiple use including recreational fishing and diving, and are in close proximity to the Kwinana Industrial Estate. In addition, Woodman Point and Cockburn Sound may suffer from high land-sourced eutrophication (Cambridge et al. 1986).

Wreck Rock (4), a site north of the previous four sites, was selected because it suffers less eutrophication but is used for commercial and recreational purposes (fishing and diving). A protective reef system, along the northern metropolitan coastline, protects this site from oceanic swells. Hamelin Bay (5 and 6) was selected because it is considered a ‘pristine’ environment and is located approximately 225 km south of Cockburn Sound. This region supports a commercial fishery and is used for recreational activities (photography, diving, fishing and boating).

Meadows at each site were monitored using two transects; one transect along a depth gradient (5–12 m) and one transect at 5 m. At Success Bank, a transect was permanently established. Qualitative swims 20 m on each side of each transect were used to monitor overall “health” of the meadow. During these swims the presence of seagrass species was recorded (Table 3.2). In this study, the health of the meadows was defined as a subjective evaluation of epiphyte load, benthic herbivore density and shoot densities, and takes into account seasonal differences. Divers swam the transects collecting quantitative samples of plant biomass, rhizomes, fruits, seedlings and sediments (on three occasions in November 1993, 1994 and 1995) as outlined in Table 3.1 below. These tasks are described in detail following the site descriptions below and in the following chapters.

Table 3.1: Field measurements (indicated by !) taken at the seven sampling sites. Site 1) Carnac Island; site 2) Success Bank; site 3) Woodman Point; site 4) Wreck Rock; site 5) Peak Island; site 6) Cosy Corner; and site 7) Cockburn Sound.

Sites Methods 1 2 3 4 5 6 7 Meadow health Shoot density Seagrass biomass Fruit collection - Fresh - Wrack Rhizome collection Rhizome elongation Fruit density Seedling density Flower spike density Free-floating fruit density Vegetative propagule recruitment

Carnac Island, 32° 07’ 39”S, 115° 39’ 94”E This meadow is located on the leeward side of Carnac Island, which is an “A” class marine reserve. During the monitoring the meadow had a constant density and was comprised of a mixed Posidonia australis and P. sinuosa dominated complex. The understorey species consisted of Halophila ovalis, Heterozostera tasmanica and Syringodium filiforme with a fringing edge of Amphibolis griffithii. Generally, P. sinuosa was found in calmer, shallow waters and P. australis in deeper, rougher water. Currents sweep around the island causing considerable erosion and accretion (Nelson 1992). Depth ranges from 0.5-8.0 m with blowouts scattered throughout the meadow. Sediments are a mixed sand/shell composite.

Success Bank, 32° 05’ 94”S, 115° 43’ 23”E Two transects (A and B) were permanently established on Success Bank in 1993. Nine sites were used along these transects (Figure 3.3). Transect A was 200 m long at the 4.9 m depth contour and had five sites. Transect B ran perpendicular to transect A and ran 200 m long down a depth gradient (Figure 3.4). Site 1 marked transect A’s mid-point and the shallow end of transect B. Site 5 marks transect B’s deepest site (12.0 m).

On Success Bank the distribution and density of seagrass species changes with depth. Dominant overstorey species are Posidonia australis and P. coriacea in shallow sites with Amphibolis griffithii replacing Posidonia species in deeper sites. Understorey species include Syringodium filiforme at shallow sites (< 5 m) and Halophila ovalis and Heterozostera tasmanica. Seagrass density was patchy with depth. The sediment was predominantly coarse, covered by a fine layer of silt in shallow areas (< 6 m) and becoming predominantly silty in deeper water (> 6 m). During the winter months sediments are compacted and silt is scarce due to erosion. The meadow composition is summarised in Table 3.2.

55 22 22 66 11 33 77

Figure 3.2: Site locations for the study of Posidonia seed viability, longevity, shoot density and biomass, fruit density, seedling density and rhizome expansion in Western Australia. Site 1) Carnac Island; site 2) Success Bank; site 3) Woodman Point; site 4) Wreck Rock; site 5) Peak Island; site 6) Cosy Corner and; site 7) Cockburn Sound. Meadow fruit and seedling densities and biomass was measured at all sites. Fruit viability, longevity, and seedling growth measurements at sites 2, 3 and 7.

6 7 8 9 Transect A 1 (4.9 m)

2 (5.8 m)

3 (8.6 m)

4 (10.7 m)

Transect B 5 (12.0 m)

Figure 3.3: Transect A and B orientation and site placement on Success Bank (refer to Figure 3.2 for site locations). Numbers in brackets refer to depth. Rhizome extension was measured on transect A. Grey areas represent vegetative propagule recruitment quadrats (not drawn to scale).

14 Depth profile 12

6 Depth (m) 4

0 12345 Sites

Figure 3.4: Depth of the Success Bank meadow sites (1-5) on transect B (refer to Figure 3.2 for location).

Table 3.2: Site, depth (m) and seagrass species examined on Success Bank. The seagrass species are: Pa = Posidonia australis; Pc = Posidonia coriacea; Sf = Syringodium filiforme; Ho = Halophila ovalis; Ht = Heterozostera tasmanica and; Ag = Amphibolis griffithii. ! denotes rhizome recruits and ! denotes seagrass presence.

Site Depth Seagrass species Pa Pc Sf Ho Ht Ag 1 4.9 2 5.8 3 8.6 4 10.7 5 12.9 6 4.9 7 4.9 8 4.9 9 4.9

Woodman Point, 32° 08’ 03”S, 115° 44’ 47”E The seagrass meadow at Woodman Point was patchy and dominated by Posidonia australis with an understorey of Halophila ovalis and Heterozostera tasmanica. No Amphibolis griffithii was noted. Depth ranged from 0.5 – 6.0 m with meadow density becoming patchy with increased depth. The sediments were characterised by coarse sand, shell and pieces of limestone rubble. Generally, sediments were fine in the meadow and coarse close to the shoreline and breakwall. The meadow is protected from southwesterly and southeasterly weather but open to northwesterly storms.

Wreck Rock, 31° 48’ 04”S, 115° 43’ 03”E Wreck Rock is a small limestone reef island surrounded by a mixed meadow dominated by Posidonia australis and P. sinuosa, with Halophila ovalis, Heterozostera tasmanica and Syringodium filiforme forming an understorey. Amphibolis griffithii forms a fringing edge on the east side of the meadow. Depth ranged from 0.5-7.0 m. Close to the island sediments were coarse, becoming finer further away from the island. The island is located on the landward side of two outer reefs that provide protection from oceanic swells. The study site was on the leeward (eastern) side of the island.

Hamelin Bay Peak Island 34° 13’ 03”S, 115° 00’ 14”E Cosy Corner 34° 15’ 52”S, 115° 01’ 50”E Two sites were selected at Hamelin Bay, which is within the Boranup State Forest:

1) Peak Island was approximately 10.5 m deep with fine to coarse sand that formed deep (20 - 30 cm) troughs. It is protected from southwesterly winds. Posidonia australis and Amphibolis antarctica dominate the seagrass meadow and occur on the top of sand troughs. Halophila ovalis and Heterozostera tasmanica formed the understorey, growing in the troughs.

2) Cosy Corner, just south of Peak Island, is an embayment sheltered from northwesterly winds by many small islands and reefs. The meadow was in 3.0 m depth, on a shell and sand substrate. Posidonia australis dominated the meadow and it was fringed by A. antarctica fringes the meadow. Halophila ovalis was the understorey species in the fringing meadow and was also present in bare sandy patches. Heterozostera tasmanica was the understorey species in the P. australis meadow.

Cockburn Sound, 32° 08’ 38”S, 115° 44’ 50”E The study meadow was located in 0.5 – 12.0 m depth. Posidonia australis dominated the meadow with an understorey of H. ovalis and H. tasmanica. Density of seagrass declined with depth. Sediments became silty with depth, with coarse material and seagrass detritus accumulating in the shallow waters. The area is exposed to southwesterly and southeasterly weather, however it is protected from northwesterly and northeasterly storms. High nutrient load and restricted water flushing has caused the loss of seagrass meadows from within Cockburn Sound (Steedman and Craig 1983; Cambridge and McComb 1984; Cambridge et al. 1986; Silberstein et al. 1986).

3.2 2 Meadow measurements Sexual propagation was examined by focussing on basic attributes such as fruit production in relation to meadow biomass and density, seed viability, seedling establishment and growth. Vegetative propagation was investigated by considering in situ rhizome elongation at meadow edges over a depth range and at similar depths, in situ rhizome elongation in vegetative recruits and rhizome elongation in a controlled environment.

Seagrass density Meadow density was determined by counting the number of shoots in ten 25 x 25 cm quadrats at all sites (see Table 3.1; Dennison 1990a). Quadrats were placed randomly for both direction and distance. Success Bank was an exception, where quadrats were haphazardly placed 2 m to the west of the transect line at sites. Measurements were made on three occasions: November 1993; November 1994; and November 1995, except at both sites in Hamelin Bay, which was only measured in November 1994. November was chosen for measurements because Posidonia australis fruits are present and biomass is high at the experimental sites at this time of year. Mean shoot densities were extrapolated to give a mean shoot density per metre squared.

Seagrass biomass Aboveground and belowground seagrass biomass was measured by removing all the seagrass from within the ten density quadrats at all sites (as described above) and washing them free of sand and detritus (see Ott 1990). In the laboratory, the samples were washed again, epiphytes removed and the seagrasses blotted dry. Total wet weight was recorded. Samples were then divided into species, aboveground and belowground structures separated, dried at 80°C for 24 hours and then weighed to give dry weight (g). Total dry weight was subsequently extrapolated to dry weight per metre square.

Rhizome collection As described in Chapter 2, Posidonia australis rhizomes from Success Bank and Wreck Rock were collected (during summer, or when required) (see Table 3.1), disinfected, labelled and then six rhizomes from each site were randomly allocated into each mesocosms, resulting in a total of 12 rhizomes in each mesocosm (see Chapter 2). Each rhizome consisted of at least five shoots and an apical meristem to ensure rhizome growth (Kirkman 1989).

Rhizome elongation Rhizome elongation was measured as growth increments, as described in Chapter 2, to minimise disturbance to the in situ rhizosphere, rhizomes were covered by sediment after each measurement, which was taken after carefully fanning away the covering substrate. Fanning and recovery disturbance had been previously shown to have no effect on growth (D. Walker and C. Simm pers. comm.).

Fruit, seedling, flower spike and free floating fruit density In situ Posidonia australis fruit, seedling and flower spike (attached or free floating) densities were collected and numbers of fruits recorded, in the ten 25 x 25 cm quadrats at Wreck Rock, Woodman Point, Success Bank, Carnac Island and Hamelin Bay described previously (see Table 3.1). Fruits that were free- floating in the water column, above the quadrats were also collected. Measurements occurred weekly during November and December in 1993, 1994 and 1995, with the exception of Hamelin, which was only measured in November 1994. Mean fruit and seedling densities were extrapolated to provide mean densities per m2.

A total of 1402 Posidonia australis wrackline fruits were collected from Woodman Point (n = 758) and Cockburn Sound (n = 644) on four occasions (November and December 1994 and 1995). Dehydrated or physically damaged fruits (by visual inspection) were not collected. No wrackline seeds from Wreck Rock could be collected as the surrounding beaches had no seeds on them and no free floating seeds were observed in the immediate vicinity of the site. A total of 2339 fresh P. australis meadow fruits were collected directly from Success Bank (n = 582), Wreck Rock (n = 372) and Woodman Point (n = 1385) meadows. Fruits were collected by divers in a haphazard fashion with multiple fruits from the same plant. Throughout this study fruits collected from wrack are referred to as ‘wrackline fruits’ and fruits collected fresh from the meadow are referred to as ‘fresh fruits’. The density of fruits varied between sites, limiting the total number of fruits collected. After collection, all fruits were washed in disinfectant (1 part Miltons® baby bottle wash to 80 parts seawater) and randomly placed into free-flowing seawater mesocosms, maintaining sites separate. Fruits were allowed to dehisce and release their seedlings. Seedling growth was also recorded (section 3.2.4).

3.2.3 Seed viability and fruit longevity Seed viability Tetrazolium solution staining is a reliable method of establishing the viability of dormant and non- dormant seeds (Smith 1951; Conacher et al. 1994b; Baskin and Baskin 1998). Seeds were divided in half

longitudinally and one half was placed on filter paper soaked in a 1.0% tetrazolium solution (2,3,5- tetrazolium chloride (Sigma Chemicals) in a phosphate buffer). Seeds at 18°C, 25°C and 30°C were examined hourly over 24 hours to determine the duration for which seeds should be in contact with tetrazolium (Figure 3.5). The temperature and time regime was selected based on discussions with C. Conacher and P. Hubbard. Following Lakon’s (1948) methods, the distribution of stained live tissue and unstained dead tissue was used to determine the viability of the seeds (see Appendix B). After 10 hours at 30°C the seeds had produced enough triphenyl formazon (identified by the formation of a pink stain) to assess their viability. Shorter duration and lower temperatures were inadequate to determine viability. All subsequent seed viability tests were conducted using 1.0% tetrazolium for 10 h at 30°C.

120 180C 100 250C 0 80 30 C

20 Tetrazolium stain (%) Tetrazolium 0

0 5 10 15 20 25 Time (h)

Figure 3.5: Twenty four-hour temperature trial to determine optimum reactivity of 1% tetrazolium with Posidonia australis seeds. Seeds were kept at 18°C, 25°C or 30°C. Seeds were derived from fruits collected at Woodman Point, Success Bank, Wreck Rock and Cockburn Sound (see Figure 3.2). Viable tissue was determined by the distribution of stained live tissue as described by Lakon (1948).

Viability was tested weekly for five weeks beginning in November in 1993, 1994 and 1995. Twenty-five fruits (five from each site) were randomly selected and tested weekly, during the sampling periods. This continued until all fruits had dehisced or begun to decompose (fresh fruits, n = 125; wrackline fruits, n = 125). Fruit numbers were limited because seedlings were needed for the mesocosm experiments (Chapter 2). Previous studies have not tested viability over time or reported the total number of seeds tested (Hubbard 1994). Conacher et al. (1994b) used three replicates of 30 seeds each (n = 90) to determine the viability of Z. capricorni seeds.

Seed longevity Seed longevity was estimated as the maximum length of time it took before all seeds either germinated or became inviable (i.e. stopped reacting with the tetrazolium solution or began decomposing). For example,

if after 14 days no more seeds reacted positively with the tetrazolium solution then longevity would be defined as 14 days. Alternatively, if all seeds germinated after 30 days then longevity would be defined as 30 days. Longevity in wrackline fruits was not determined because the exact date of fruit release from parent stock is unknown.

3.2.4 Seedling growth All collected fruits were separated into collection site and kept in mesocosms until they dehisced and released the seedlings. By keeping sites separate, comparisons between sites could be made. Few wrackline seeds developed into seedlings and hence, the wrackline data from Woodman Point and Cockburn Sound was pooled to give total wrackline seedling growth. Many fresh seeds germinated, allowing seedling growth to be examined for each site where fresh fruits were collected. Stages of seedling growth were measured and recorded as appearance of a cotyledon, appearance of roots and emergent leaves, lengthening of the root, root hair production and increasing leaf number. Growth was divided into three phases: 1) emergence of a cotyledon, 2) production of roots and emergent leaves and 3) lengthening of roots.

3.2.5 Rhizome elongation over depth and between species on Success Bank Ten Posidonia australis rhizomes and 15 rhizomes each of Halophila ovalis, Heterozostera tasmanica, Syringodium filiforme and P. coriacea were tagged (as described in Chapter 2) and measured at each site along Transect A (see Table 3.1). Transect B, sites 2 and 3 had 15 rhizomes of each species present tagged and measured. Rhizome elongation over depth (transect B) was measured monthly for 23 months at sites 1, 2 and 3 using the methods described in Chapter 2.

3.2.6 Vegetative propagule recruitment Natural vegetative propagule (rhizome) recruitment was observed and measured at sites 1 to 5 on Transect B, Success Bank (Figure 3.3) approximately three times a month (weather permitting) for 23 months. Rhizome recruitment was considered to have occurred when pieces of rhizome attached to the substrate and began to grow. Between each site parallel to Transect B two transect lines of approximately 20 m length and 2 m wide were monitored for new rhizome recruits. When a new recruit was found, the apical meristem was tagged and growth increments were measured monthly for rhizome elongation (as described in Chapter 2). Successful establishment was defined as the recruit remaining in place for at least two periods, producing roots and exhibiting rhizome extending.

3.2.7 Statistical analyses When analysing seagrass shoot density, seagrass biomass (aboveground and belowground), rhizome elongation, rhizome recruitment and comparisons of fruit viability (wrack versus fresh) one way ANOVA and repeated measure (RM) one way ANOVAs were used to test for differences between multiple sites and multiple treatments (F statistic). When one way ANOVA tests of normality or equality of variance failed, non-parametric analysis was done using a Kruskal-Wallis test (H statistic) (Zar 1996). Repeated

measure one way ANOVA non-parametric analysis was done using a Friedman’s RM ANOVA test (χ2 statistic) (Zar 1996).

Viability of fruit (wrack and fresh) was analysed using t-tests. Simple linear regressions were used to detect associations between seagrass shoot density (independent variable) and biomass; shoot density (independent variable) and fruit density; shoot density (independent variable) and seedling density; biomass (independent variable) and fruit density; and biomass (independent variable) and seedling density.

3.3 Results All meadows evaluated were considered healthy, based on the subjective monitoring of epiphyte loads, benthic herbivore density and shoot density. Although Posidonia coriacea was present at sites 1, 2, 6, 7, 8 and 9, it was not sampled in the quadrats.

3.3.1 Sexual Reproduction Meadow shoot density and biomass Density: During the three consecutive sampling periods (November 1993, 1994 and 1995) each species shoot density did not change significantly. The combined shoot density of all species at a site, significantly differed between sites (H[6] = 4.92, p < 0.05; Table 3.3). The combined shoot densities were similar at Carnac Island, Success Bank, Cosy Corner and Peak Island and were significantly greater than shoot densities at Wreck Rock (H[3] = 2.56, p < 0.05;), Woodman Point and Cockburn Sound which were

similar to each other (H[4] < 0.001 , p > 0.05; Table 3.3).

Shoot density also significantly differed between species (H[4] = 22.3, p < 0.05). Heterozostera tasmanica was the densest seagrass at all sites except Woodman Point, followed by Halophila ovalis, Syringodium filiforme, Posidonia australis and Amphibolis griffithii (Table 3.3). The understorey species H. tasmanica and H. ovalis had similar, high shoot densities that were higher than the dominant meadow species P. australis and A. griffithii (Table 3.3). Shoot densities of the climax species (P. australis, A. griffithii) and S. filiforme were low and were significantly different from the understorey species, H. tasmanica and H. ovalis (Table 3.4).

Table 3.3a: Mean shoot densities of a range of seagrasses at seven sites in Western Australia (see Figure 3.2). Sites are as in Table 3.1. Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera.

Shoot density (shoot m-2 ± standard error (SE)) Pos. Amph. Syrin. Hal. Het. Σ Total Site understorey density 1 488 ± 23 824 ± 41 792 ± 56 1431 ± 126 3212 ± 167 5435 ± 293 6747 ± 357 2 426 ± 16 0 ± 0 813 ± 44 1601 ± 135 3233 ± 144 5647 ± 279 6073 ± 339 3 188 ± 11 0 ± 0 0 ± 0 1382 ± 87 1250 ± 77 2632 ± 164 2820 ± 175 4 412.5 ± 38 0 ± 0 704 ± 47 560 ± 29 805 ± 52 2069 ± 81 3205 ± 119 5 512 ± 35 309 ± 14 0 ± 0 1563 ± 110 3163 ± 188 4725 ± 298 5546 ± 347 6 752 ± 43 282 ± 12 0 ± 0 1616 ± 122 4123 ± 201 5739 ± 323 6773 ± 378 7 365 ± 21 0 ± 0 0 ± 0 611 ± 17 1256 ± 94 1867 ± 111 2232 ± 132

Table 3.3b: Student-Newman-Keuls multiple comparison results illustrating significant differences (S) and no significant differences (NS) between shoot density of different species.

Species Significance (S) Posidonia NA Halophila S NA Heterozostera S NS NA Amphibolis NS S S NA Syringodium NS S S NS NA Species Posidonia Halophila Heterozostera Amphibolis Syringodium

Biomass: At all sites above and belowground biomass did not change significantly during the three consecutive sampling periods (H[6] = 7.1, p > 0.05). The understorey species (Heterozostera tasmanica, Halophila ovalis and Syringodium filiforme) had the smallest above and belowground biomass at all sites (Tables 3.4 and 3.5). Amphibolis griffithii had the greatest aboveground biomass, followed by P. australis, H. tasmanica, S. filiforme and H. ovalis (Table 3.4a). Amphibolis griffithii belowground biomass was also greater than P. australis, but not significantly so (H[2] < 0.001 , p > 0.05; Table 3.4b). Frequently the aboveground biomass of a species was greater than belowground biomass (Tables 3.4 and 3.5).

With the exception of P. australis, above and belowground biomass of each species differed significantly between sites (see Tables 3.4 and 3.5). Meadow biomass (all species combined) at each site had 2 2 significantly different total above and belowground biomass (χ [1, 90]= 14.1, p < 0.05; χ [1, 90]= 14.3, p < 0.05, respectively), with meadows at Carnac Island, Wreck Rock and Cosy Corner having large biomass and meadows at Success Bank, Woodman Point, Peak Island and Cockburn Sound having a low biomass (Tables 3.4 and 3.5).

Table 3.4a: Mean aboveground biomass of a range of seagrasses at seven sites in Western Australia (see Figure 3.2). Sites are as in Table 3.1. Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera.

Aboveground biomass (g dry weight m-2 ± SE) Pos. Amph. Syrin. Hal. Het. Σ Total Site understorey density 1 176 ± 12.8 398 ± 19.2 24.8 ± 1.1 18.9 ± 0.6 22.7 ± 0.8 66.4 ± 2.6 640 ± 34.6 2 242 ± 9.6 0 ± 0 38.4 ± 4.0 15.0 ± 1.6 28.8 ± 1.9 82.2 ± 7.5 324 ± 17.1 3 267 ± 11.7 0 ± 0 0 ± 0 1.12 ± 0.3 7.5 ± 2.1 8.62 ± 2.4 276 ± 14.1 4 291 ± 33.6 570 ± 274 80.0 ± 16.2 6.40 ± 1.6 42.1 ± 13.1 128.5 ± 30.9 989 ± 338 5 187 ± 54.4 144 ± 25.6 0 ± 0 4.64 ± 0.4 9.0 ± 0.8 13.6 ± 1.2 345 ± 81.1 6 210 ± 17.6 339 ± 68.8 0 ± 0 4.16 ± 0.3 0 ± 0 4.16 ± 0.3 553 ± 86.7 7 274 ± 16.0 0 ± 0 0 ± 0 4.32 ± 0.2 5.6 ± 0.3 9.92 ± 0.5 284 ± 16.5

Table 3.4b: Mean belowground biomass of a range of seagrasses at seven sites in Western Australia (see Figure 3.2, Table 3.1). Species names are abbreviated: Pos. = Posidonia; Amph. = Amphibolis; Syrin. = Syringodium; Hal. = Halophila; Het = Heterozostera. The understorey species are Syringodium, Halophila and Heterozostera.

Belowground biomass (g dry weight m-2 ± SE) Pos. Amph. Syrin. Hal. Het. Σ Total Site understorey density 1 140 ± 14.4 147 ± 9.6 21.8 ± 0.9 15.4 ± 1.3 17.44 ± 2.1 54.6 ± 4.2 343 ± 28.2 2 123 ± 46.4 0 ± 0 12.0 ± 1.8 19.4 ± 2.7 19.0 ± 1.7 50.4 ± 6.2 174 ± 52.6 3 130 ± 16.0 0 ± 0 0 ± 0 2.27 ± 0.3 25.3 ± 0.4 27.6 ± 0.7 157 ± 15.7 4 136 ± 67.2 134 ± 36.8 31.0 ± 0.2 5.76 ± 0.7 33.0 ± 33.6 69.8 ± 34.5 340 ± 138 5 107 ± 30.4 84.8 ± 14.4 0 ± 0 5.60 ± 0.5 6.72 ± 0.8 12.3 ± 1.3 204 ± 46.1 6 147 ± 9.6 224 ± 44.8 0 ± 0 6.72 ± 0.6 0 ± 0 6.72 ± 0.6 378 ± 55.0 7 101 ± 19.2 0 ± 0 0 ± 0 1.41 ± 0.2 1.79 ± 2.4 3.20 ± 2.6 104 ± 21.8

Table 3.5: Above and belowground biomass testing the hypothesis that above or belowground biomass was similar between sites, for seagrass species from sites in Western Australia (see Figure 3.2, Table 3.1).

Species Statistical value Significance Posidonia australis Aboveground biomass H[6] = 24.1 p < 0.05 Belowground biomass NA - Amphibolis griffithii Aboveground biomass H[6] = 61.9 p < 0.05 Belowground biomass H[6] = 61.7 p < 0.05 Syringodium filiforme Aboveground biomass H[2] = 7.06 p < 0.05 Belowground biomass H[2] = 48.2 p < 0.05 Halophila ovalis Aboveground biomass F[6,50] = 119 p < 0.05 Belowground biomass F[6,50] = 37.7 p < 0.05 Heterozostera tasmanica Aboveground biomass H[5] = 22.4 p < 0.05 Belowground biomass H[5] = 24.5 p < 0.05

Posidonia fruit, seedling and flower spike densities Posidonia australis fruit, seedling, floating fruit and flower spike densities are summarised in Table 3.6.

Fruit and seedling density did not significantly differ (fruit: F[2, 9] = 0.0001; p > 0.05; seedling F[2, 12] = 0.001; p > 0.05) during the three consecutive sampling periods (November in 1993, 1994 and 1995). No fruits were observed on plants or floating in the water column at Success Bank or Wreck Rock however, flower spikes were present. No seedlings were observed at Success Bank. Cosy Corner and Peak Island had the highest fruit and seedling densities, followed by Woodman Point, Wreck Rock and Carnac Island (Table 3.6). Floating fruit densities were highest at Woodman Point, followed by Cosy Corner and Peak Island.

Table 3.6: Average Posidonia australis densities (number m-2) of fruits, seedlings, floating fruits and flower spikes at sites in Western Australia (see Figure 3.2, Table 3.1); n = 30 for all sample sites.

Site Fruit Seedling Floating fruit Flower spikes 1 18.0 ± 1.5 23.7 ± 1.1 8.7 ± 1.2 8.7 ± 1.6 2 0 0 0 62.4 ± 17.6 3 26.1 ± 1.7 39.0 ± 4.3 53.2 ± 17.2 12.0 ± 1.5 4 0 32.0 ± 10.9 0 24.0 ± 8.5 5 38.1 ± 2.2 47.3 ± 3.4 35.6 ± 13.3 14.4 ± 1.9 6 39.1 ± 2.7 44.6 ± 1.3 37.2 ± 13.1 14.6 ± 1.7

The density of fruits from site to site were significantly different (H[6] = 64.7, p < 0.05), with Cosy Corner and Peak Island having higher densities than other sites (Figure 3.6). Seedling densities differed

significantly from site to site (H[6] = 50.7, p < 0.05) with Table 3.7 summarising the SNK results. Fruit density is positively correlated with seedling density (r2 = 0.69; Figure 3.6). Meadow shoot density or biomass did not influence the density of fruits and seedlings.

Table 3.7: Student-Newman-Keuls multiple comparison test results, indicating significant differences (S) and no significant (NS) differences between seagrass shoot densities at different sites in Western Australia (see Figure 3.2; Table 3.1).

Site Significance (S) 1 NA 2 S NA 3 S S NA 4 NS S S NA 5 S S NS S NA 6 S S Not tested S Not tested NA 7 S NS S S S S NA Site 1 2 3 4 5 6 7

) 50 -2

0 ling density (m density seagrass seed ling Mean

0 1020304050 Mean seagrass fruit density (m-2)

Figure 3.6: Correlation between seedling density and fruit density, in the seagrass Posidonia australis (r2 = 0.69, p > 0.05).

Posidonia seed viability and longevity

Seed viability was not significantly different (F[4,22] = 2.84; p > 0.05) during the sampling periods of November in 1993, 1994 and 1995.

Fresh fruits: Seed viability of fresh fruits did not differ significantly between the sites of Success Bank,

Woodman Point and Wreck Rock or during the sampling period (F[2,14] = 62.3; p > 0.05, respectively; Figure 3.7a). At Success Bank viability peaked at 81.3% after five days and declined to 0% by 31 days.

The longevity of Success Bank seeds was 31 days. Seed viability at Woodman Point, peaked at 72% on day five, remaining high (62.5%) until day 14. At day 31, 0% of the remaining seeds were viable (Figure 3.7a). Measurements stopped after 31 days because all seeds had either started to grow (indicating viability) or to decompose. Thus, Woodman Point seeds had at least 31 days to successfully settle in an area. Seed longevity was 31 days. Seed viability at Wreck Rock peaked at 76.5% after five days, decreasing to 62.5% by 14 days (Figure 3.7a). Viability testing was terminated at 21 days because seeds were no longer viable and were decomposing. Seed longevity at Wreck Rock was short (22 days).

Wrackline fruits: Seed viability of wrackline fruits differed significantly between Woodman Point and

Cockburn Sound (t[23] = 2.65, p < 0.05; Figure 3.7b). At Woodman Point, viability peaked at 11.2%, in nine days (Figure 3.7b). After 47 days, few seeds were viable, with many fruits decomposing. Cockburn Sound seed viability peaked at 8.8% in 12 days (Figure 3.7b).

The viability of fresh and wrackline fruits was significantly different (F[1,123] = 17.86, p < 0.05) and wrackline fruits are less viable. However, wrackline fruit longevity had the potential to be greater than fresh fruits because 46 days after collection wrackline fruits were still testing viable, whilst fresh fruits had stopped testing viable by the 30th day (Figure 3.7b).

Posidonia seedling growth With the exception of Carnac Island, seedlings at all sites had similar patterns of growth. Typically, seedlings grew to a length of 5-7 cm and had a persistent green cotyledon. In situ seedling rhizomes were adventitious when growing on hard substrates but grew a taproot in soft substrate on meadow edges. Seedling density was highest at meadow edges, with fewer seedlings located inside the meadows and no seedlings observed in the blowouts.

The growth of seedlings in the mesocosms, collected from all the sites, is summarised in Table 3.8. Seedlings were also used in experiments described in Chapter 2. Growth of seedlings in the mesocosms

did not differ significantly between sites (F[3,8] = 0.29; p > 0.05) or between seedlings derived from fresh

or wrackline fruits (t[4] = 0.92; p > 0.05). Seedlings growing in the mesocosms had adventitious root growth. A high number (87.6 ± 0.9%) of wrackline seedlings succumbed to bacterial and fungal infection despite using a disinfectant solution, with slightly fewer (73.4 ± 1.0%) fresh seedlings dying from infection.

Seedlings from fresh fruits: These seedlings grew more rapidly than wrackline seedlings (Table 3.8). Regardless of site, these seedlings had similar average growth patterns, with rhizomes increasing in length by 3.4 ± 0.25 mm per day. Wreck Rock seedlings developed (i.e. cotyledon appearance, root and emergent leaves appearance etc) the fastest (22 days), but grew slowest with average root elongation rates of 3.0 ± 0.18 mm per day. Success Bank and Woodman Point seeds took longer to develop but grew an average root elongation rates of 3.5 ± 0.2 mm per day and 3.7 ± 0.19 mm per day, respectively.

100

20 Fresh seagrass seed viability (%)

0 0123456 Time elapsed (days)

100

20 Wrackline seagrass seed viability (%) 0 0 10203040506070 Time elapsed (days)

Figure 3.7: Posidonia australis seed viability of a) fresh fruits at Woodman Point !, Wreck Rock !and Success Bank !, and b) wrackline fruits at Woodman Point ! and Cockburn Sound ! (see Figure 3.2). Time elapsed for wrackline fruits began from when the fruit was collected from wrack, not when it was released from the plant.

Seedlings derived from fresh meadow fruits, survived longer than wrackline seedlings with 11.9 ± 0.1% surviving after 20 weeks at the experiment completion. The majority of the surviving seedlings came from Woodman Point.

Wrackline seedlings: The number of seedlings was reduced when 50% of fruits did not dehisce. A small percentage (2.6%) of seedlings developed a cotyledon, emergent leaves and roots during the experimental period. Few (0.78%) seedlings produced roots and had roots that extended in length (Table 3.8).

Wrackline seedlings were less robust than fresh meadow seedlings. The majority of these seedlings stopped development within 5 days and all seedlings had died within 47 days (Table 3.8).

Table 3.8: Percentage growth of Posidonia seedlings collected from fresh meadow fruits (Wreck Rock, Success Bank and Woodman Point) and wrackline fruits (combined Woodman Point and Cockburn Sound). Growth is divided into three phases (1) emergence of a cotyledon, (2) production of roots and emergent leaves and (3) lengthening of roots. Time is measured in days.

Fresh seedlings Wrackline seedlings Combined Woodman Wreck Rock Success Bank Woodman Point Point and Cockburn Sound Time % Time % Time % Time % Site seedling seedling seedling seedlings Phase 1 5 d 86 5 d 90 5 d 96 5 d 21

Phase 2 9 d 12.8 9 d 8 9 d 6.3 9 d 1

Phase 3 22 d 2.4 31 d 2.4 31 d 3.6 47 d 0

3.3.2 Vegetative reproduction Rhizome elongation In situ rhizome elongation was significantly higher in spring and summer compared to the autumn and winter months (F[1,90] = 6.21, p <0.05). The species that had the fastest elongating rhizome, across all depths, was Heterozostera tasmanica, followed by Halophila ovalis (Table 3.9). Both species increased rates of elongation as depth increased. The increase was not significant in H. tasmanica, however it was significant for H. ovalis between sites 2 and 3 (F[4,90] = 4.09, q = 3.96, p < 0.05). Each species’ rhizomes

elongated at different rates (F[4,90] = 5.70; p < 0.05), with H. tasmanica rhizomes growing much faster than all other species (Table 3.9). Furthermore, at sites 1 and 2, all species grew at significantly different rates to each other (site 1: H[3] = 11.1, p < 0.05; site 2: H[4] = 11.9, p < 0.05), whilst at site 3 the species grew at similar rates.

The rhizomes of Amphibolis griffithii and Syringodium filiforme both elongated rapidly (Table 3.9), with elongation rates within site 2 being similar. A spatial comparison of rhizome growth in these two species was not possible because each species occurred at only one site (see Table 3.2).

Posidonia australis rhizome elongation decreased, although not significantly with increasing depth or site (Table 3.9). Posidonia australis rhizomes, at sites 1 and 2 elongate at approximately 1.04 ± 0.15 mm d-1 and 0.96 ± 0.12 mm d-1, respectively. The slowest of all species was P. coriacea (0.49 ± 0.06 mm d-1 and 0.46 ± 0.04 mm d-1, at sites 1 and 2 respectively) which, like P. australis, had a rate that decreased with increasing depth (Table 3.9).

Table 3.9: In situ mean (± SE) rhizome growth rate (mm d-1), from fastest to slowest, for six species of seagrass at three sites on Success Bank, with increasing depth. Site 1 is 4.9 m deep, site 2: 5.8 m, and site 3 is 8 .6 m depth (see Figure 3.4). NA – The species were not present at these sites.

Rhizome elongation rate (mm d-1) on Transect A Site 1 Site 2 Site 3 Species n SE n SE n SE H. tasmanica 15 1.4 1.2 15 4.2 1.2 15 4.8 0.9 H. ovalis 15 2.4 0.6 15 1.2 0.5 15 4.0 0.6 A. griffithii NA NA NA 15 1.7 0.3 NA NA NA S. filiforme 15 1.2 0.2 NA NA NA NA NA NA P. australis 51 1.0 0.2 15 1.0 0.1 NA NA NA P. coriacea 15 0.5 0.1 15 0.5 0.04 NA NA NA

Vegetative propagule Rhizomes (vegetative propagules) were recruited at all sites on Transect B at Success Bank. At site 1, recruited rhizomes (10) were short-lived, with three recruits establishing and elongating at an average rate of 0.78 ± 0.02 mm d-1. Sites 2 and 3 recruited many rhizomes (19 and 23, respectively), however none

elongated and all were short lived. Site 4 recruited (18) rhizomes that were all short lived with two recruits establishing and elongating (averaging 0.56 ± 0.01 mm d-1). Site 5 recruited the most rhizomes (36), with 28 recruits establishing and elongating at an average of 0.40 ± 0.02 mm d-1. Two species of Posidonia recruited vegetatively on Success Bank, P. australis and P. coriacea, however no P. coriacea recruits elongated. As depth increased elongation rates significantly decreased (F[5,31] = 92.4, p < 0.05).

The recruited rhizome growth rates were significantly lower than established rhizomes (H[40] = 47.6, p < 0.05).

3.4 Discussion This section of the study was designed to examine sexual and vegetative propagation in Posidonia australis and other seagrass species, and determine how meadows are maintained and extend into new territory. This in turn will provide an understanding of P. australis sexual and vegetative propagules efficacy in order to select an appropriate transplant unit. A comparison of data with other species is included.

3.4.1 Sexual reproduction Posidonia australis shoot density and plant biomass Meadow density and biomass was examined to see if it changed over time and between species or sites. During November 1993, 1994 and 1995, the density and biomass of the meadow species examined did not change significantly. Although meadow density and biomass presumably fluctuated over the year due to seasonal variations in conditions such as light, day length, water temperature and water motion. The constant density and biomass during November for these three years suggests that the plants were mature and perhaps, temporally stable (i.e. they may be following a cyclical pattern that recurs year after year). Study site meadows (see Figure 3.2) were composed of species that displayed classical colonising and climax species characteristics. The density and biomass between individual species within the meadows was significantly different. The colonisers, Heterozostera and Halophila, had high shoot densities and low biomass (Tables 3.3, 3.4 and 3.5). Syringodium (a coloniser) fell in-between the climax and pioneer species characteristics, having mid to high shoot density and low biomass (Tables 3.3, 3.4 and 3.5). Climax species, Posidonia and Amphibolis, had low shoot densities and high aboveground biomass (Tables 3.3, 3.4 and 3.5).

Total meadow density (all species combined) significantly differed between sites, with Carnac Island, Success Bank, Cosy Corner and Peak Island all supporting denser meadows than Wreck Rock, Woodman Point or Cockburn Sound (Figure 3.2). This may be due to differing hydrodynamic characteristics, predation or disturbance factors between sites. For example, both Cockburn Sound and Woodman Point are affected by high nutrient levels that are responsible for seagrass declines in these areas (Cambridge 1980; Cambridge and McComb 1984; Cambridge et al. 1986).

The meadow density and biomass results in this study have been within reported values for all the species examined. Posidonia australis shoot densities (Table 3.3) were within Cambridge’s (1980) reported

ranges (142-1,857 shoots m-2). Although, when comparing Cambridge’s sites with ones of similar depth in this study, Cambridge’s sites had much lower densities (2-3 times less). Posidonia australis biomass (294–427 g dry weight m-2; see Tables 3.4 and 3.5) was also within reported ranges (140–1,280 g dry weight m-2; Kirkman and Reid 1979; West and Larkum 1979; Cambridge 1980; Walker and McComb 1988; Paling 1994). Similarly, the ranges in this study are also comparable to the Mediterranean species, P. oceanica. Posidonia oceanica density can range from 164 to 1278 shoots m-2 and differs between localities at equivalent depths (Pergent et al. 1994; Pergent-Martini et al. 1994). The cause of the variation between localities is unclear but as mentioned above, it may relate to hydrodynamic characteristics, predation or disturbance factors between localities (Pergent et al. 1994; Pergent-Martini et al. 1994).

Thus, seagrass species density and biomass did not vary during the three measuring periods and was significantly different between species (r- and K-selected species) and sites

Posidonia fruit and seedling density: Fruit and seedling density did not change during the three sampling periods inferring that P. australis annual seed production did not vary between 1993, 1994 and 1995. The persistent number of fruits in the meadows, coupled with the similar shoot densities and biomass during the three years of sampling is indicative of meadows that are reproductively mature. As plants reach reproductive maturity the number of fruits in the canopy increases, peaks (or plateaus) and then gradually decreases as plants reach old age (Baskin and Baskin 1998).

Posidonia fruit and seedling density varied significantly between sites. Fruit density (39.1 ± 2.7 fruits per m2; Table 2.6) recorded in this study was much lower than those reported within the literature (391.4 ± 40.6 fruits per m2, Hocking et al. 1980). The reason for this discrepancy is unknown but may be related to plant age, disturbance events (storms and animals) and predispersal predation. These factors are known to influence both terrestrial and seagrass communities (Nienhuis and Groenendijk 1986; Clarke and Kirkman 1989; Sheperd et al. 1989; Merkel 1990a; 1990b; Ewanchuck 1995; Baskin and Baskin 1998).

Meadow density and biomass did not influence fruit or seedling densities. This result was unexpected. Terrestrial studies have demonstrated density-dependence where shoot density and biomass of a meadow will influence seed set (Ellner 1985a, 1985b; Venable and Brown 1988; Venable 1989; Baskin and Baskin 1998). However, in this study, this was not the case during the three sampling periods; Posidonia was density-independent. Furthermore, as fruit density increased, so did seedling density, implying that seedlings are recruited close to and within the parent meadow. Hence, seed set was density-independent (Figure 3.6). This reinforces Waycott’s (1995) research that Posidonia meadows are genetically diverse and are maintained by sexual reproduction (i.e. recruitment of seedlings). A disadvantage to density- independence is that seedlings may end up in a meadow where shoot density is high, creating a low-light environment that restricts growth and survival of the seedlings. Density-dependent species, on the other hand, recruit to areas of low shoot density, reflecting density-escaping characteristics. Density-escaping occurs when a species utilises a seed bank to escape the effects of sibling competition and hence reduce

the risk of extinction (Venable and Brown 1988). Zostera marina settlement is density-dependent using a seed bank and recruiting better to areas of low vegetative shoot density (Hootsmans et al. 1987; Orth et al. 1994; Ewanchuck 1995), however seedling mortality is not density-dependent (Harrison 1993).

Through competition, the environment organisms occupy often controls the r- and K-selected traits exhibited by many organisms. Typically, organisms fall between the two ranges, showing some characteristics of each trait. This seems to be the case for Posidonia australis at the examined sites because, Posidonia clearly had some K-selection traits: it invested heavily in growth and maintenance (i.e. large biomass; Tables 3.4 and 3.5), had a small reproductive commitment (i.e. producing few fruits in comparison to colonising species such as Halodule engelmannii; see McMillan 1987; Table 3.6) and takes a relatively long time to mature. However, K-selected species are also intensely competitive and limit recruitment by young within adult meadows (Krebs 1985). As mentioned above, Posidonia was density- independent and seedling density increased with fruit density suggesting that recruitment opportunities were not limited. Waycott and Les (1996) clearly show that some populations of P. australis exhibit either r- or K-selected traits.

Few other researchers consider the influence of shoot density and biomass on fruits and seedlings (an exception is Ewanchuck 1995), making comparisons, and calculations of density-dependence impractical. Thus, it is difficult to extract this information from the literature.

Seed viability and longevity Posidonia australis seed viability and hence inferred longevity was not significantly different during the sampling periods, suggesting that the sampled meadows were mature and stable. In this study, fresh fruit density was low (Table 3.6), viability was high (72 – 81.3%; Figure 3.7) and did not differ between sites. However, viability was low in wrackline fruits (Figure 3.7) and significantly differed between sites. Viability differences could be explained by locality differences, time spent by fruits drifting in the water column, disturbance (such as burial) and physical damage. Locality differences in physical factors were multi-variate and not considered in this study however, Pergent-Martini et al. (1994) demonstrated that even similar sites could exhibit significant differences.

Fruits of Posidonia spend, on average 1-2 days floating before the pericarp splits (Hocking et al. 1981), and thus are often within close to proximity to the parent meadow. Spending greater time drifting in the water column may result in fruits reaching unsuitable settlement sites, such as beaches or deep water. For example, P. oceanica fruits often settle in water that is too deep for survival or growth (Meinesz et al. 1993). Anecdotal evidence from Western Australia suggests that P. australis and P. sinuosa fruits often wash ashore during the reproductive season (R. Peters pers. comm.). Disturbance of seeds by either biotic or abiotic factors acts to decrease viability of seeds, especially those within seed banks (Venable and Brown 1988; Rees 1997; Baskin and Baskin 1998), including, pre- and postdispersal predation and once washed ashore, sunburn and dehydration. However, if seeds are orthodox, they can withstand desiccation while recalcitrant seeds will perish if moisture content drops below a relatively high, critical value (Chin

1988). It is unknown whether P. australis seeds are recalcitrant or orthodox, but it is likely that they are ‘recalcitrant’ because wrackline fruit had reduced viability (compared with fresh fruits) but increased seed longevity (Figure 3.7).

Viability of fresh fruits and wrackline fruits was significantly different. Wrackline seed longevity was greater than freshly collected seeds (Figure 3.7) which may be due to a physiological reaction bought about by the fruits being cast upon the shoreline. Reductions in moisture content (reductions can be in the range of 5–10%) are known to double the length of life of seeds in dry storage (Harrington 1973; Baskin and Baskin 1998). When the seeds imbibe water, they rapidly become functional and germinate (Baskin and Baskin 1998). The wrackline fruits collected in this study had been cast upon the shore for a short period of time (< 3 hours) and were still in relatively moist conditions (fruits on top of the wrack which were sunburnt or damaged were discarded, whilst fruits further down in the wrack looked healthy and were collected). If they had been cast onshore longer then a reduction in longevity may have been seen because extended dry storage of seeds is known to decrease viability (Baskin and Baskin 1998).

Few studies have investigated Posidonia seed viability in Australia. The difficulty in measuring viability in this genus relates to a lack of seed dormancy period making the exact time of germination obscure (Cambridge 1980; McComb et al. 1981; A. McComb, pers. comm 1997). Hocking et al. (1980) is one of the few studies to have investigated P. australis and P. sinuosa seed viability (P. australis had 6.50 ± 0.61 viable fruits per flowering shoot). However, Hocking et al.’s (1981) information is insufficient to calculate what percentage of fruits from a meadow is viable on a m2 basis. Larkum (1976) has observed that flowering in P. australis, in Botany Bay, was a rare event, but he did not provide information on seed viability.

More viability studies have been conducted with other Australian species including Zostera capricorni and Thalassia pachyrhizum. Conacher et al. (1994a and 1994b) demonstrated that the tropical seagrass Z. capricorni had low seed viability and germination rates (26 ± 2%). Although after storage for 50 days seed viability increased to 58 ± 10%, indicating the need for a dormancy period (Conacher et al. 1994b). Kuo and Kirkman (1987) detected low viability and germination rates in T. pachyrhizum, and only 10% of stems bore inflorescences, with less than 10% of flowers becoming seeds. Low viability and germination rates in Z. capricorni and T. pachyrhizum imply that vegetative reproduction is primarily responsible for maintaining extant populations. This is substantiated in Z. capricorni meadows, which use vegetative reproduction when flowering density is low, while meadows with high flowering density maintain the population through both sexual and vegetative reproduction and have been demonstrated to supply seeds to establish new beds (Conacher et al. 1994b).

Typically, information on viability within the literature focuses on fruit/seed numbers, germination and establishment. Studies that involve species with seed banks typically do not measure viability (for examples see Caye and Meinesz 1985b; Pirc et al. 1986; Hootsmans et al. 1987; McMillan 1987; Jewett- Smith and McMillan 1990; Loques et al. 1990; Churchill 1992; Moore et al. 1993; Terrados 1993;

Ewanchuck 1995). If viability is low and vegetative reproduction is poor then it is assumed that numerous seeds are necessary to ensure survival and maintenance of a meadow (Conacher et al. 1994b), but species that have persistent seed banks overcome this requirement for high seed set (Baskin and Baskin 1998). For example, some Zostera marina populations rely on a persistent seed bank for recruitment in order to survive and maintain a presence in the environment (Harrison 1993; Ewanchuck 1995). Baskin and Baskin (1998) described four types of seed banks (Table 3.10) and related the manner in which these ensure meadow maintenance and species survival in the environment. It appears that Z. marina is a type III species (Table 3.10), producing a persistent population that germinates in spring or autumn (locality dependent) and has a small seed reserve (Robertson and Mann 1984; Ewanchuk 1995). If both seed viability and recruitment is high then sexual reproduction may be the primary reproductive method if establishment potential is high. This is often not the case for Z. marina populations (Orth and Moore 1983; Harrison 1993; Ewanchuk 1995).

Table 3.10: Four types of seed banks described by Baskin and Baskin (1998).

Type Seed Bank description Germination season Seed reserve size I Transient Autumn Small II Transient Spring Small III Persistent Primarily autumn Small IV Persistent Autumn Large

Longevity significantly differed between sites, which was expected since viability and longevity are directly linked. Few studies have examined the seed longevity of Australian seagrass species (McMillan 1991) and none have examined Posidonia australis. The lack of a dormancy period in Posidonia seeds suggests that longevity may also be short. The difficulty in determining longevity for Posidonia may be responsible for this, however measurements in this study suggest that longevity of fresh fruits is less than 30-31 days and ~ 45 days for wrackline fruits (Figure 3.7). This is considered short in comparison with other seagrass species, which may remain viable for many years (McMillan 1991; Baskin and Baskin 1998). Colonising species, such as Syringodium, Halodule, Zostera and Cymodocea may have seeds that remain dormant but viable for years (Phillips 1972; McMillan 1983; 1991; Caye and Meinesz 1985b; Pirc et al. 1986). Prolonged longevity allows colonising species to be opportunistic, making the best of any opportunity to germinate and establish.

Posidonia australis seedling growth characteristics Posidonia australis seedling growth characteristics were examined over three years (1993, 1994 and 1995) to determine if this species establishes in high enough numbers to maintain a meadow. Seedling growth in the mesocosms was not significantly different between the three sampling periods. A high percentage (91%) of P. australis fresh fruits germinated (Table 3.8). However, few fresh or wrackline seedlings survived, with fewer (< 10%) producing emergent leaves and primary roots, and even less (< 3%) with roots that lengthened during the experiment (Table 3.8).

Despite seedlings settling on the substrate and attaching, the majority did not produce emergent leaves and consequently they were unable to photosynthesise and relied on stored nutrients. Hocking et al. (1980; 1981) demonstrated that Posidonia seedlings have stored carbohydrates that should last, on average, for nine months before the seedlings are required to photosynthesise their own carbohydrates. Seedling mortality in this experiment was well within this prescribed nine months (see Table 3.8) and therefore unexpected, yet was consistent over the three sampling periods. Failure of seedlings to establish is common in the environment for Posidonia (Hancock 1992; R. Peters pers. comm.) and other seagrass species, such as Z. marina and C. nodosa (Hootsmans et al. 1987; Duarte and Sand-Jensen 1990a; Harrison 1993; Orth et al. 1994; Ewanchuck 1995; van Lent and Verschuure 1995). However, the fragile nature of seedlings meant that the seedlings did not respond well to the handling that occurred during the experiment and this handling may have caused their demise (effects of handling seedlings has previously been discussed in Chapter 2).

Although Posidonia australis recruitment potential was high (fruit density matched seedling density; Figure 3.5) and seed viability was high (Figure 3.7), few seedlings established. Low establishment has also been observed in the species Zostera marina (Harrison 1993; Moore et al. 1993; Ewanchuck 1995) and Cymodocea nodosa where > 70% of seedlings fail to establish (Duarte and Sand-Jensen 1990). Seed dispersal is limited (Tutin 1938; Orth et al. 1990), with Zostera relying primarily on asexual propagation to maintain and establish meadows (Duarte and Sand-Jensen 1990; Ewanchuck 1995). Cymodocea is believed to rely on sexual reproduction to form patches and hence spread, although it fruits erratically, produces few seeds and seedling mortality is extremely high (Caye and Meinesz 1985b; Duarte and Sand- Jensen 1990a; Terrados 1993). Seagrass meadows in the Caribbean fruit prolifically, establish poorly and thus rely on vegetative fragments to recolonise the region after disturbance (Williams 1990).

Posidonia australis seedlings are commonly found in the areas studied. Dredge-sand storage sites at Woodman Point, are regularly seen with a dense covering of P. australis seedlings (R. Peters pers. comm.; unpubl. data). Similarly, P. coriacea and P. sinuosa seedlings settle in the Success and Parmelia Bank regions, however few establish (Hancock 1992; E. Paling pers. comm.). Failure to establish may be due to poor anchoring, predation, disturbance or, early death. The natural establishment process starts with seedlings settling and being pulled into the substrate via a primary root (Kirkman and Kuo 1990). This anchoring process is purported to be a slow process (Kirkman and Kuo 1990; Hancock 1992) and can be easily disrupted in conditions of high erosion, such as those experienced during storms.

The lack of establishment in this study may be due to unfavourable environmental characteristics of the sites. However, if seedlings settle in close proximity to parent meadows, as inferred for Posidonia australis in this study (Figure 3.6) and in Zostera marina elsewhere (Orth et al. 1994) then conditions should theoretically be appropriate. In terrestrial plants large numbers of seeds are produced but few are viable and fewer still settle and establish (Baskin and Baskin 1998). As mentioned earlier factors such as plant age, disturbance, pre- and postdispersal predation and infection may act to reduce the number of

available seeds and seedlings in seagrasses. In fact, it is common for few seedlings to settle and establish (Baskin and Baskin 1998). In this study, predation of seedlings may also be a major contributing factor to low establishment. The sea urchin, Temnopleuris michaelsenii, is commonly seen grazing upon Posidonia plants in the study areas (Cambridge et al. 1986; Hancock 1992; B. Wilkes pers. comm.).

Posidonia australis did not settle in high numbers and those that did died in this study, based on mesocosm work. This was verified by field observations. In conclusion, P. australis produced few fruits during 1993, 1994 and 1995. The production of fruit was independent of meadow density but settlement was correlated with fruit density. Seeds had high viability and yet, establishment was low. It appears therefore that sexual reproduction and recruitment may be limited in Posidonia at the study sites examined.

3.4.2 Vegetative reproduction Rhizome elongation Rhizome elongation was measured to determine elongation rates of different species. The rate of elongation was significantly greater in spring and summer, compared to autumn and winter. This was expected as a result of reduced light levels in autumn and winter, which directly results in a reduction of primary productivity and hence rhizome growth (Raven et al. 1986). Significant differences were detected between colonising and climax species. Colonising species, such as Heterozostera, Halophila and Syringodium elongated at significantly greater rates than climax species, such as Posidonia and Amphibolis (Table 3.9). At all depths Heterozostera, followed by Halophila (two r-selected species) had the fastest elongation rates (9.7 times faster than Posidonia coriacea, 4.5 times faster than P. australis, 4.0 times faster that S. filiforme and 2.8 times faster than Amphibolis griffithii), high shoot densities (Table 3.3), low biomass (Table 3.4 and 2.5) and a fluctuating population size (Krebs 1985; see Chapter 2). The fast rhizome elongation rates of colonisers (Heterozostera, Halophila and Syringodium) contribute to their opportunistic qualities. Slow rhizome elongation rates are a trait of K-selected species, which prevents them from being opportunistic colonisers. In contrast, Posidonia and Amphibolis are K- selected (climax species) and have restricted capabilities to colonise a region due to competition with colonising species.

The differences between colonising and climax species can be explained by differences in life history attributes and competition principles. For example, colonising species such as Heterozostera, Halophila and Syringodium, are capable of quickly invading an opening in the substrate, and rapidly colonising an area, yet because they do not invest heavily in growth, they are short lived, have no competitive ability and are easily displaced by larger species such as Posidonia. Such species have population sizes that vary with time, have a low long term biomass (Bulthuis and Woelkerling 1983; Birch and Birch 1984; Brouns 1987b; 1987c; Gallegos et al. 1994; Tables 3.4 and 3.5), produce many propagules and are short lived, all features typical of r-selection.

In contrast, Posidonia and Amphibolis have population sizes that are constant in time, meadows develop slowly, have a greater biomass (Kirkman and Reid 1979; West and Larkum 1979; Cambridge 1980; Walker and McComb 1988; Paling 1994; Tables 3.4 and 3.5) and live longer than one year. Such features typify a climax species, which are subjected to K selection (Krebs 1985; Raven et al. 1986). Other seagrass species that possess K-selected characteristics include Phyllospadix scouleri, Enhalus acoroides, Thalassia hemprichi and Thalassia testudinum; these are climax species that have a high biomass and low shoot density (Littler et al. 1983; Brouns 1987a; 1987b; 1987c; Erftemeijer et al. 1993; Gallegos et al. 1993).

The r- and K-selected characteristics displayed by species are frequently reported in the literature (Cambridge 1980; Bulthuis and Woelkerling 1983; Fry and Virnstein 1988; Aioi and Pollard 1993; Paling 1994; Paling et al. 1997). For example, the coloniser H. tasmanica typically grows in intertidal regions and has an aboveground biomass of 27-61 g dry wt. m-2 and a density of 600-2000 leaf clusters m-2 (Bulthuis and Woelkerling, 1983). Further examples of r-selected seagrass species (which have low biomass and high density) include Cymodocea rotundata, Halodule uninervis, Halodule wrightii, Syringodium isoetifolium and Zostera japonica (Brouns 1987b; 1987c; Erftemeijer et al. 1993; Baldwin and Lovvorn 1994; Gallegos et al. 1994).

In this study, Posidonia australis rhizomes elongated at a relatively slow rate (~36.5 cm y-1; Table 3.9) however this fell within reported rates in the literature (2.6 – 1800 cm y-1; Clarke and Kirkman 1989; West 1990; Paling 1992; Sinclair, Knight and Partners 1993; Kirkman 1995; MAFRL, unpub. obs.). It is a widely reported that Posidonia rhizomes elongate slowly (West 1980; Kirkman 1989; Kirkman and Kuo 1990; West et al. 1990; West 1990; Hancock 1992; Walker 1994; Kirkman 1995). West (1990) reported that actively growing Posidonia rhizomes elongated at rates close to this study (29 cm y-1), however it was estimated that there were few (1 in 600) actively growing shoots present in the meadows studies. Kirkman and Kuo (1990) suggest that both P. australis and P. coriacea complexes have no vegetative reproductive capabilities and slow rhizome growth. In contrast this study demonstrated that 100% of marked rhizomes (n = 60) were actively growing. This suggests that P. australis vegetative growth is active and elongation rates are sufficient for vegetative expansion, exceeding values for Zostera marina (16 cm y-1; Olesen and Sand-Jensen 1994) which spreads rapidly.

Rates of rhizome elongation in Posidonia oceanica are usually similar to P. australis (den Hartog 1970; West et al. 1990). However, in this study P. australis rates were much greater than previously reported P. oceanica rates (6-11.5 cm y-1; Caye 1980; Cooper 1981; Meinesz and Lefevre 1984; Meinesz et al. 1993). Other reported elongation rates for climax species are slower than observed in this study. Zostera marina is capable of 16 cm y-1 (Olesen and Sand-Jensen 1994) and Thalassia testudinum elongates at 27.2 – 36.5 cm y-1 (Erftemeijer et al. 1993; Gallegos et al. 1993). Measured Amphibolis elongation rates in this study (62 cm y-1; Table 3.9) exceeded rates reported in the literature (1.7-2.5 cm y-1; Hancock 1992; Walker 1994). Differences between reported rates and those in this study may be due to hydrodynamic effects (Fonseca et al. 1983; Brouns 1987; Fonseca and Kenworthy 1987; Duarte and Sand-Jensen 1990b) and

climatic differences (i.e. tropical versus temperate records) at the compared localities. For example, the constant movement of sand in high energy regions will necessitate the continual recolonisation of areas, stimulating vertical rhizome elongation and thus rates may differ if the area is continually disturbed (Littler et al. 1983; Baron et al. 1993; Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997).

The elongation rates of the colonising species Heterozostera tasmanica (126.5 cm y-1) and Halophila ovalis (93 cm y-1) were also comparable to those in the literature, although slightly slower (Brouns 1987; Kirkman and Kuo 1990). For example in tropical waters, Brouns (1987) has recorded H. ovalis rhizome elongation of 365 cm y-1. Cymodocea species are capable of much greater rates (1600 -7400 cm y-1; Caye and Meinesz 1985a; Brouns 1987; Duarte and Sand-Jensen 1990a). Syringodium elongation rates (43.8 cm y-1; see Table 3.9) were similar to reported values (50.0 – 200.7 cm y-1), yet slower (Williams 1990; Short et al. 1993; Gallegos et al. 1994). Once again slower rates than those reported in the literature may be due to the hydrodynamic (Fonseca et al. 1983; Littler et al. 1983; Brouns 1987; Fonseca and Kenworthy 1987; Duarte and Sand-Jensen 1990b; Baron et al. 1993; Marbà and Duarte 1994; Duarte et al. 1997) and climatic differences between localities.

The seagrasses in this study fit into the r- and K-selection strategies described by ecological theory. As the data described above suggests, Halophila ovalis, Heterozostera tasmanica and Syringodium filiforme are r-strategists, whilst Amphibolis griffithii and Posidonia australis and P. coriacea are K-strategists. The differences in elongation rates for the species studied support the conclusion that these meadows are mature, “climax” communities. The K-selected species are growing at comparable or greater rates to those in the literature but the r-selected species are all growing slower suggesting a less suitable environment.

The effect of changing depth on rhizome elongation The effect depth has upon rhizome elongation was examined for a number of species. Although elongation rates differed between species, there was no significant decrease with depth for any species (Table 3.9). Halophila ovalis however, elongated at a significantly greater rate at depth (Table 3.9). Halophila is a species well adapted to depth because of its a small rhizome, reduced leaf clustering and decreased leaf density; factors that enable it to maximise light capturing abilities. However, these adaptations may affect photoinhibiton potential in shallow water. The small rhizome is also unfavourable in shallow water because disturbance from water motion may easily damage the rhizomes. In contrast, Heterozostera was more of an understorey species possibly allowing it to escape the effects of water motion, while Halophila actively colonised bare areas at the study sites (pers. obs).

The larger species, Syringodium, Posidonia and Amphibolis, were unaffected by depth, perhaps because they have strong, large rhizomes making them more robust to water motion and disturbance. It was expected that shallower depth would be associated with higher growth rates due to increased light levels (Drew 1979; Dennison 1987; Williams 1988; Dawes and Tomasko 1988; Salisbury and Ross 1992) however, it was not statistically significant (Table 3.9). Differences in elongation between depths may be

due to P. australis growth plateauing and therefore, differences in rhizome elongation are negligible between these sites (Table 3.9). When light levels become saturating, growth plateaus or photoinhibition occurs in some species (Salisbury and Ross 1992). Light levels at site 1 (5 m) peaked at 321 µmol m-2 s-1 (during summer) and 144 ± 42 µmol m-2 s-1 (during winter) (unpub. data). Although light levels dropped marginally between sites 1 and 3 (5 to 8.6m) the levels still exceeded reported saturation irradiance (Isat) of P. australis from southern Western Australia (73 µmol m-2 s-1, LeProvost Environmental Consultants 1990; Masini et al. 1990). The plateau in growth could be true for P. coriacea, though it is difficult to verify because P. coriacea and P. australis saturation irradiance data in the study region does not exist.

Compensation and saturation irradiance levels for Halophila ovalis, Heterozostera tasmanica, Syringodium filiforme and Amphibolis griffithii are unknown. Since this data is lacking it is difficult to know why elongation decreased between sites 1 and 2 and then increased at site 3 (Table 3.9). Masini et al. (1990) detected that Posidonia sinuosa is affected by photoinhibition when light levels are high and water temperature is low (e.g. winter). It is possible that at the shallower depths in this study, elongation rates are affected by photoinhibition because the Isat exceeded those reported in the literature (LeProvost

Environmental Consultants 1990; Masini et al. 1990). Isat and compensation (Ic) levels reported in the literature for P. australis (LeProvost Environmental Consultants 1990; Masini et al. 1990) are based on

regional studies in southwestern Australia only. It is unlikely that Ic and Isat are consistent, particularly since water temperature significantly varies along the latitudinal gradient of the Western Australian coastline. Therefore, extrapolating Masini et al.’s (1990) values to other Western Australia sites has the

potential to grossly under- or over-estimate the true Ic and Isat of a region.

Rhizome recruitment The recruitment of vegetative fragments occurred at all depths with Posidonia australis and P. coriacea rhizomes having moderate to low recruitment success. After settling, the recruits often died quickly. Recruitment was more successful at depth, in terms of numbers and survival, however elongation rates were slower (0.40 ± 0.02 mm d-1) than shallower sites (0.78 ± 0.02 mm d-1). No P. coriacea recruits were seen to elongate, which may be due rhizome growth exhibited by this species (rhizomes extend down, not horizontally). Deeper sites were less disturbed by surge and swell causing material to be more readily deposited, greatly increasing the probability of rhizomes recruiting and establishing.

Elongation rates of vegetative recruits significantly decreased with increasing depth. This observation was opposite to in situ meadow rhizomes, which were unaffected by depth, however the recruit experiment depth gradient (4.9–12.0 m) was greater than the in situ rhizome elongation work (4.9–8.6 m) from above. As discussed above, deeper water has less available light and hence would be associated with slower elongation rates, as was the case for the rhizome recruits. Generally, higher light levels associated with shallow water induced faster rhizome elongation yet shallow water has greater water movement, which results in lower recruitment. For example, meadows in calm waters have few blowouts, are almost continuous in their extent and have an almost continuous annual seedling recruitment (Fonseca et al. 1983; Fonseca and Kenworthy 1987; Duarte and Sand-Jensen 1990b). This premise can be extended to

vegetative recruitment, with calmer waters enabling greater recruitment, although even in rough water the presence of a barrier can enhance recruitment of vegetative propagules.

The most significant observation from this work is that vegetative propagules of Posidonia australis established and their rhizomes elongated. Maintenance of seagrass meadows by vegetative growth occurs in Zostera marina populations (Tomlinson 1974), Cymodocea nodosa meadows (Duarte and Sand-Jensen 1990a) and mixed Thalassia testudinum, Syringodium filiforme and Halodule wrightii meadows (Williams 1990). This occurs when seed production, germination and seedling establishment is low (Williams 1990; Harrison 1993). Therefore, in this area, vegetative propagules of seagrasses, particularly P. australis, can contribute to meadow maintenance by recruiting to and establishing in disturbed areas and spreading.

3.4.3 Meadow maintenance and choice of transplant unit Posidonia seedling density matched fruit density in the sites examined (Figure 3.6), implying that seedlings settle within the meadow. Shoot density or plant biomass did not affect recruitment, suggesting density-independence in this species. Posidonia australis produces many viable fruits (Table 3.6; Figure 3.7) yet few seedlings establish and survive (Table 3.8). These observations are supported by the failure of seedling based transplant efforts in Western Australia. Seven major efforts have met with varying degrees of success. Five of these studies (Cambridge 1979; Kirkman 1989; Le Provost 1990; Hancock 1992; Kirkman 1995) placed seedlings in areas of low water quality and with poor results. Other studies met with higher success in less eutrophied water (Cambridge 1977, 1978; Kirkman 1995), however, the expansion of transplant units (seedlings) was very slow (Kirkman 1995).

Posidonia australis rhizome elongation was not significantly affected by depth. Rhizomes at 5 m elongated, with rates peaking at 37.96 ± 5.48 cm y-1. Vegetative recruitment of P. australis and P. coriacea was high; many P. australis recruits established and elongated at depth. Recruits in shallow water elongated fastest (almost equal to natural meadow rhizomes), but few were able to establish for longer than a short time. Posidonia australis recruits are therefore more likely to be more successful at depth because conditions are calmer, which allows recruitment to settle and attach.

In some seagrasses, such as Zostera capricorni, both sexual and vegetative reproduction sustains the community and may control meadow expansion into new areas (Conacher et al. 1994a; 1994b). The Mediterranean species Posidonia oceanica is transplanted using vegetative propagules because availability of its sexual propagules is low due to its limited flowering ability and low rate of successful sexual reproduction (Meinesz et al. 1991; Meinesz et al. 1993). Given the above observations for P. australis, seedlings would be considered a poor transplant choice for revegetating seagrass meadows. Vegetative propagules appear to be a more viable option given the measured elongation rates and the observed natural vegetative recruitment success.

Thus, by drawing from the present information and a review of pertinent literature a suitable transplant unit can be identified. The selection criteria for Stage I (site selection) and Stage II (transplant unit and technique) are: 1) Stage I (site selection) a) Irradiance levels above 5% SI; b) Water quality needs to be high and not eutrophied; c) Short burial periods with cyclical burial periods preferred, which infers that; d) Water movement should be low or reduced; and e) Substrate type can either be limestone rubble or coarse sand, with sand a better choice due to ease of transplanting and anchoring ability. 2) Stage II (transplant unit and technique) a) Vegetative propagules are better transplant units than sexual propagules; i) Plugs are preferred vegetative transplant units over sprigs; b) Plugs should be spaced in close proximity, however a trade-off between plug spacing and overall area should be evaluated (further investigations are recommended); c) Handling should be minimised; i) The donor meadow should be in close proximity to the transplant site to reduce handling time; d) Larger transplant units are preferred (> 15 cm diameter); and e) Donor meadows should be in deeper water than the transplant site (further investigation is recommended).

In conclusion, sexual propagules of Posidonia australis would appear to be a poor choice for transplant unit because establishment is low. In contrast, vegetative propagules such as rhizomes, have higher establishment rates and elongation is comparable to extant meadows. Vegetative propagules therefore appear to be the logical transplant unit for P. australis.

Chapter 4 Evaluating vegetative transplant success in Posidonia australis: a field trial with habitat enhancement

4.1 Introduction In an attempt to control the decline of seagrasses, much effort has been directed toward restoration research (Fonseca et al 1996; Butler and Jernakoff 1999). For restoration to succeed effective understanding of both community processes and ecosystem functions is required. Knowledge of both the fundamental and realised niche of a targeted restoration species will aid in the identification of appropriate mitigation sites and transplant techniques. In many cases this information may not be known or poorly investigated over the geographic range of this species. In the literature a number of hypotheses have been put forth such as those outlined in Chapter 1.

In addition, the recognition that habitat enhancement may be necessary to encourage species growth is required. Many innovative and often simple techniques have been used to enhance a transplant site. For example, anchors (e.g. Hancock 1992; Walker 1994) and artificial barriers are used in high energy regimes (Chapter 2), while mesh is used to increase accretion in high erosion regions (E. Paling pers. comm.) or the exclusion of burrowing animals (e.g. Merkel 1990a). The success of site enhancement is predicated on the selection of a transplant unit suitable to the enhancement method and selection of the transplant site (e.g. substrate). Anchoring easily damages seedlings and may increase drag, increasing the likelihood of dislodgment. While anchoring sprigs and plugs is more effective, it provides no protection from high water motion. Both mesh and artificial seagrass allow a choice of transplant unit (seedlings, sprigs or plugs) and simultaneously enhance the site by decreasing water motion (disturbance).

Successful transplant attempts, in general, select a site with high water quality and low disturbance, use vegetative propagules, and enhance the transplant habitat (see Appendix A). Transplant failure often occurs in regions where water quality (and therefore light levels) is low, disturbance still occurs, and where sexual propagules are used. For example, seeds are often consumed or “lost” in the environment. As a result, selection criteria can be identified which aid in both site selection, and transplant unit and technique selection as discussed in Chapter 3.

4.1.1 Restoration and mitigation Before attempting to restore or mitigate for seagrass losses, the interrelated factors of site selection (stable versus unstable conditions) transplant unit (sexual or vegetative propagules) and the requirement for habitat enhancement need to be examined.

Site selection As discussed in Chapter 2, selection of an appropriate seagrass transplant site is paramount to success. Transplantation from shallow to deeper water has been unsuccessful in Europe (Molenaar and Meinesz 1992), possibly due to the low light levels in the estuarine/bay environments where the seagrasses transplantations occurred (Thorhaug et al. 1985; Thorhaug 1985, 1986, 1987). However, this may not be the case in Western Australia. Water clarity is high in the oligotrophic coastal waters. Posidonia species in Western Australia can grow to a depth of 21 m (E. Paling pers. comm.) and have been shown to naturally recolonise dredged channels up to 2-14 m (Paling 1995). Similarly, deep water P. oceanica transplants have grown successfully in the oligotrophic waters of the Mediterranean where nitrogen and phosphate are limiting and hence, phytoplankton blooms are rare (Drew and Jupp 1976; Drew 1978; Pirc 1986; Pergent and Pergent-Martini 1991; Duarte 1991). If P. australis grows in deeper (>14 m) water then areas that have been dredged can be used as transplant sites. If growth in deeper water is poor or transplant success is low then dredged areas must either be restored to previous (pre-dredge) depths or disregarded as appropriate sites for P. australis transplants. To determine the depth limits of P. australis in Western Australia further work must be conducted.

When mitigating seagrass loss, the selection of transplant sites is often directed by political and/or economic rationale. For example, in the Florida Keys, urbanisation has destroyed many acres of seagrasses, which must be mitigated for under U.S. legislation. One such program saw 93.3 acres of seagrass destroyed and of the 62.8 acres deemed suitable for restoration only 47.5 acres were planted. After ten years this area was considered restored via natural recolonisation, not via the seagrass transplants (Ruckelshaus and Hays 1988). Similarly, in other restoration efforts the transplants themselves appear not have succeeded but natural recolonisation via recruitment mitigated the loss instead (Thorhaug 1983). Failure of transplant attempts in the U.S. is often directly related to the selection of sites where there is increased light attenuation or a high likelihood of disturbance (e.g. foraging animals, burial, physical displacement) (Roberts 1974; Thorhaug 1974, 1983, 1985; Williams 1988; Merkel

1990a; Zimmerman et al. 1995; Ruckelshaus and Hay 1998), resulting in many restoration efforts not succeeding because of poor site selection (Thorhaug 1974, 1983, 1985).

This is also true of many failed transplant attempts in Western Australian. Burial and erosion compromised the transplant sites of Hancock (1992) and Walker (1994), that were in close proximity to Carnac Island. LeProvost Environmental Consultants (1990) chose sites affected by light attenuation and strong currents, which killed all seedling transplants. Early work by Cambridge (1980) noted that sites in eutrophied waters were likely to fail. Transplant work by Walker (cited in Lord et al. 1999) and Kirkman (1995) in Warnbro Sound, Cockburn Sound and Rottnest Island, and Bastyan’s (1996; G. Bastyan pers. comm.) work in Oyster Harbour, southwestern Australia, confirm this. Although this work is under the influence of economic and political pressures (transplantation of seagrass needed to be in an area within close proximity to a sand mining operation, which may need to mitigate for seagrass losses in the future), there was a certain degree of freedom, which allowed the selection of a site that was most appropriate for this region.

The process of site selection examined in Chapter 2 concluded that light levels, burial and handling disturbance and substrate type have the potential to affect a seagrass transplantation attempt. Criteria for site selection determined that Posidonia australis transplants should be made into locations where burial periods are short, light levels are above 5% of the surface irradiance and substrate type can be either limestone rubble or coarse sand (see Table 2.7). Also handling disturbance should be minimised

A brief review of the environmental characteristics of the region directly south of Fremantle resolved that Success Bank (see Figure 4.1) was a suitable place for a Posidonia australis transplant attempt based on the criteria outlined in Chapter 2, Table 2.7. Sites that were disregarded included Carnac Island (water regime was too rough leading to prolonged burial periods) and Woodman Point (nutrient levels were high and close proximity to dredge spoil region increased likelihood of light attenuation).

Fremantle Success Bank Transplant Site Carnac Island Woodman Point

Cockburn Sound Garden Island

Figure 4.1: Posidonia australis transplant site ( situated on Success Bank, Western Australia.

Transplant unit Posidonia australis produces many viable fruits yet few seedlings establish and survive, while vegetative recruitment and establishment of P. australis is high with a potential for rhizome growth of 1.04 ± 0.15 mm d-1 at 5 m depth (see Chapter 3). This implies that vegetative propagules of P. australis are the best transplant unit. There are two choices of vegetative transplant unit to consider: sprigs or plugs (Phillips 1990b). Sprigs are rhizomes that have had the sediment removed (‘bare roots’). Sprigs tend to need more time to establish and spread and hence a time lag exists between transplant and meadow formation. In contrast, transplanting plugs creates an instant meadow. Plugs are capable of rapidly spreading and restoring an area because almost immediate post-transplant, plugs are capable of both sexual and vegetative reproduction. Large plugs are also very efficient at trapping debris, which helps to establish and bind the plug to the recipient site.

The choice of sprigs or plugs may largely be species dependent and determined by individual growth characteristics making one method more cost efficient. Both sprigs and plugs have been used with varying degrees of success. Within the last decade Posidonia oceanica sprigs have become the favoured transplant unit of French researchers (Molenaar and Meinesz 1991; Meinesz et al. 1991; Molenaar and Meinesz 1992; Molenaar 1992; Meinesz et al. 1992; Molenaar et al. 1993; Genot et al. 1994). A large majority of sprigs have survived transplantation in these efforts, leading to their continued use. Sixty percent of reported sprig transplants (those listed in Appendix A) in the United States are considered successful by the researcher. For example, Ware (1993) transplanted Zostera marina sprigs and recorded 80% survival. Similarly, Fonseca et al. (1979) reported that transplanted Z. marina sprigs, which were stabilised using mesh, resembled adjacent meadows one season after transplantation. In

Australia, transplants using plugs are most successful (50%), followed by sprigs (33%) and then seedlings (31%) (Lord et al. 1999; see Appendix A). In Australia, ironically, seedling transplant units are the preferred method.

Success with plugs has been varied, with early French work favouring this method, although results have been less satisfactory (Cooper 1976, 1977, 1978, 1979). Similarly, plug transplants in the United States have been less successful (45%) than sprigs (60%) (see Appendix A). Interestingly, few studies in the U.S. recorded expansion of transplants, despite a statement of success. Two exceptions are Hoffman (1990) and Harrison (1990) using Zostera marina transplants in which expansion of both sprig and plug transplant units was observed, but equivalency did not occur. More recently, Western Australian research has concentrated on plugs, rather than seedlings with good results. Walker (1994) had sound results with plugs (50-80% survival), while her sprig transplants had lower survival. For example, 12 months after transplantation, 40% of Posidonia coriacea plugs and 40% of Amphibolis griffithii plugs had survived. Furthermore, 85% of large (0.25 m2) plugs mechanically transplanted had also survived (Lord et al. 1999; E. Paling pers. comm.). Unfortunately expansion rates have yet to be measured for these transplants.

Western Australian researchers (see Cambridge 1980; LeProvost Environmental Consultants 1990; Hancock 1992; Paling 1992; Kirkman 1995; Lord et al. 1999) have had low transplantation success rates with Posidonia australis seeds, seedlings and sprigs and mid to high success with plugs. Based upon this outcome, the results from Chapter 3 (vegetative propagules as a preferable transplant unit) and the recommendations of Phillips (1980) and Thorhaug (1986) (see Chapter 1), it was determined that plugs would be used as the transplantation unit for further study.

Habitat enhancement Upon selection of a transplant unit (in this case plugs) a planting technique that ensures adequate anchorage and protection must be determined. If necessary, habitat enhancement may be needed to reduce sediment movement (erosion and accretion). In Australia and elsewhere, anchoring and mesh/matting are popular methods ensuring sprigs and seedlings remain in place (Phillips 1974, 1980; Thorhaug 1986; Kirkman 1989, 1990; Hancock 1992; Walker 1994).

Individual plant anchors, although used frequently by researchers (~27% of reviewed papers), have many disadvantages and a moderate success rate (58% success of efforts which used plant anchors). The most detrimental problem to using anchors is that seedling or sprigs must be used; both transplant unit types are vulnerable to damage that can easily occur through anchor usage (Hancock 1992; Kirkman 1995). The damage reduces the transplant units’ capacity to survive and establish and for rhizome elongation. An anchor may also hold the transplant unit in place but, it does not modify the habitat to increase the suitability of the region to transplants.

Artificial barriers and mesh (which has with no leaf structure) are used less frequently (1% and 10.5%, respectively) by researchers and have low success rates (0% and 10%, respectively). The prime benefit to artificial barriers is that they alter water flow and can decrease water velocity creating a calmer environment that is more suitable to transplantation. But, barriers are often costly because of size and due to a lack of extensive research, they remain an unknown entity. Barriers create ethical problems because they can be detrimental to one habitat to the benefit of another. However, artificial barriers can offer a solution for locations where the energy regime is high and subsequent water motion too rough for seagrass survival.

Mesh holds transplants in place better than anchors and is less time consuming to implement than anchors but, it too has a number of disadvantages. Like anchors, seedlings or sprigs are commonly used with mesh with similar problems. When mesh is used in synergy with plugs success is increased (Merkel 1990a, 1990b; Walker 1994; E. Paling pers. comm.). Various types of mesh break down at differing rates: wire mesh corrodes; plastic mesh can be weak; and biodegradable mesh can degrade rapidly (pers. obs.). Mesh encourages the accretion of sediments that can suffocate the transplant units (Merkel 1990a; Walker 1994). However, accretion can be used to combat erosion problems at a site (E. Paling pers. comm.). Neither barriers nor mesh mimic seagrasses and thus lack the three-dimensional structure of a seagrass plant. This lack of dimension results in a diminished floral and faunal assemblage and a lack of reduction in water motion seen in natural conditions.

Artificial seagrass (ASG) mats have been used at the same frequency (1%) and with a similar success rate (0%) as artificial barriers. ASG mats hold transplant units (sprigs and plugs) in place like mesh, but unlike mesh, they are an effective seagrass mimic. Thus, ASG mats

provide a three dimensional structure that: encourages similar floral and faunal assemblages to seagrass leaves, aiding in nutrient cycling; baffles water, slowing it’s progression when entering the ‘leaf’ canopy; stabilises the substrate; and appears to encourage the deposition of suspended material at a similar rate to seagrass plants. Disadvantages to ASG mats include cost and the effort required in manufacturing and placing the mats on site.

Seagrasses in Western Australia are found in both estuaries and coastal regions, however the majority inhabit coastal regions. Western Australian coastal regions are generally exposed to high-energy regimes, creating an environment too rough for unsecured transplants (see Cambridge 1980; LeProvost Environmental Consultants 1990; Hancock 1992; Nelson 1992; Paling 1992; Walker 1994; Kirkman 1995). As previously mentioned, the favoured method of stabilisation has been anchors (e.g. Growool™ spun rock cubes and metal pegs), which have been moderately successful (Hancock 1992; Paling 1992; Kirkman 1995). Mesh has been used less frequently and has also been unsuccessful (Kirkman 1995). However, when plugs and mesh are used together success increases (Walker 1994; E. Paling pers. comm.). ASG mats offer an improvement over mesh because of the physical mimicry of seagrass meadows, enhancing a site prior to transplanting and increasing transplant success (West et al. 1990; Nelson 1992).

No studies in Western Australia have used ASG mats to modify a transplant area so that it resembles a natural meadow. The site selected based on results from Chapter 2 (Success Bank) meets all of the criteria for a suitable habitat except it has unstable sediments (Nelson 1992; Walker 1994). Although Posidonia australis rhizomes respond well to burial, rhizomes cannot maintain growth during prolonged burial, such as what would occur on Success Bank. ASG mats will modify the Success Bank habitat by decreasing burial and stabilising the sediments, increasing the potential for transplant success. Because of limited transplantation success and the increasing pressures placed upon seagrass meadows it has become imperative to develop a cost-effective transplantation method that will accelerate seagrass recolonisation, revegetation or meadow expansion rate.

4.1.2 Aims This chapter aims to further evaluate the physical limitations influencing Posidonia australis transplant success and to investigate the development of an innovative method of seagrass transplantation using procedures where sites can be modified to improve survival of

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transplant units and accelerate meadow expansion. To do this my working hypotheses adapted from Chapter 1 are: HIa’ depth (i.e. light) has no affect on transplant success; HIe energy regime has no affect on transplant success (both from Stage I site selection); and HIIIe artificial seagrass mats enhance transplant survival (from Stage III habitat enhancement).

4.2 Materials and methods 4.2.1 Site selection The donor meadow on Success Bank was approximately 9 m deep and was selected because it was due to be dredged in the near future. The site was a patchy mixed meadow of Posidonia australis, P. sinuosa, P. coriacea, Amphibolis griffithii, Halophila ovalis and Heterozostera tasmanica. The mean density of this meadow (all species) at the time of removal of plugs was 1446 ± 0.012 shoots m-2 with a biomass of 1410.9 ± 0.31 g dry weight m-2. Density and biomass were measured using the techniques described in Chapter 3, section 3.2.2.

As discussed in Chapters 2 and 3, Transect B on Success Bank was used as the recipient site. This site ranged in depth from 5 to 12 m with meadow coverage in shallow water and bare substratum in deep water. The recipient site was selected by reviewing likely transplant sites using the Stage I (site selection) criteria discussed in Chapter 2 (see Table 2.7).

4.2.2 Habitat enhancement ASG mats were used to stabilise the sediments and were placed at the sites seven months prior to transplantation. One mat was placed and anchored at each of the five sites on the recipient meadow transect. Each ASG mat was 1.5 x 1.5 m with plastic mesh as the base of the mat (mesh size is 60 x 40 cm). Each cross bar of the mat held an individual artificial shoot, providing a total density of 1,944 shoots per mat. An artificial shoot consisted of three, 2 cm wide, clear polythene plastic leaf blades of three different lengths (6, 30 and 38.5 cm). The total surface area of one shoot was 298 cm2 and of one mat was 579,312 cm2. Shoots were attached to the mats using copper staples. A metal picket was fastened to the underside of each mat to stabilise it in the field and the mats were anchored using a metal picket at each corner of the mat.

Monitoring stabilisation effects

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Sediment grain size, erosion and accretion inside and outside of ASG mats were used as a guide to the effectiveness of the ASG mats at stabilising the environment (see Merkel 1990a, 1990b). Measurements were made on a fortnightly (erosion and accretion) or monthly (sediment grain size) basis, beginning three months before the ASG mats were put in place, continuing for seven months until seagrass plugs were transplanted and then for 14 months after the transplantation.

Sediment Grain Size Distribution: Sediment grain size fractions were used to indicate the water velocity of a site. In areas of high water velocity sediments are typically coarse, while low water velocity areas have silty sediment (Merkel 1990a, 1990b; Nielsen 1992). Four sediment cores (8 x 3 cm) were taken at each site: two within and two outside the ASG mat (approx. 1 m away). At each site, the core samples within the ASG mat were combined and sub-sampled to produce two 100 g samples. This treatment was repeated for the core samples outside the ASG mat. Sub-samples were dried at 70°C for 24 h, weighed and sieved through 2.0, 1.0, 0.5, 0.25, 0.125 and 0.0625 mm sieves. Each resulting size fraction was weighed and a mean was taken for each replicated pair to give the sediment grain size at each site, within and outside of the ASG mats. Sediment grain sizes were categorised by particle diameter into three classes; coarse sand (2.0 - 0.2 mm), fine sand (0.2 - 0.02) and silt (0.02 - < 0.002).

Erosion and Accretion: The trough height of a sand wave is considered to be closely related to hydraulic roughness (Nielsen 1992), making erosion and accretion (trough height) an effective measure of how ASG mats are stabilising a site. Trough height, at each site, was measured fortnightly using four marking pickets, two within the ASG mats and two in adjacent areas. Erosion was indicated by a decrease in the height of the sediment against the marking picket, whilst accretion was indicated by an increase. The two pickets inside and outside the ASG mats were averaged to give a mean erosion/accretion level.

4.2.3 Seagrass transplantation Seagrass transplants were collected from donor meadows and transplanted into the recipient meadow. Transplantation efforts were aided by the Australian Navy Clearance Divers (Clearance Diving Team Four, HMAS Stirling). Prior to beginning the transplantation work Navy divers were briefed in their task ensuring that plugs of the correct seagrass species were

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collected from the growing meadow edge. A brief description of each step in this process follows:

Transplant unit Each plug was 15 cm in diameter by 15 cm in depth. The plug size was selected because larger diameter plugs were difficult to remove from the substrate without the use of specialised lifting equipment. Also, as planting size increases survival follows a saturation response curve (Merkel 1990b). Walker (1994) demonstrated that 15 cm diameter cores were close to the largest plug size that survived before becoming saturated. Bird et al. (1994) and Walker (1994) also reported transplanting plugs with numerous apices was very effective. In this work, a 15 cm diameter plug of Posidonia australis had at least 3 ± 0.02 apices, whilst a smaller plug had fewer apices.

Transplant collection Plugs were removed from the growing edge of the donor meadow on Success Bank ensuring that plagiotropic rhizomes were collected. Donor plugs were removed by hammering 20 cm lengths of 15 cm diameter PVC pipe into the leading edge of the meadow. To aid in cutting through the meadow rhizomes, the PVC pipe edges were sharpened. Once the pipe was approximately 15 cm into the sediment it was levered from the substrate and the sharpened end was capped to prevent the loss of the plug. Plugs were then immersed in seawater, in low light conditions until transplanted (which occurred less than 1 h after removal). All plugs were collected and transplanted during one day.

Plug transplantation Each site on the recipient meadow was planted with 36 plugs of P. australis. The majority (28) of plugs were placed into an ASG mat area. The remaining plugs (8) were haphazardly placed outside the ASG mat to provide a control to determine the effect ASG mats have on plug survival. Prior to transplanting into the stabilised ASG mat the middle section of the mat was removed, to create an area for the transplantation. A plug was planted by positioning the PVC pipe containing the plug over the transplant plot, liquidising the sediment below the pipe, removing the cap and pushing the plug out of the pipe and into the sediment. Back- filling of the sediment then occurred. All transplants were placed into holes that were approximately 15 cm deep, allowing shoots to be at a similar height above the substrate as in situ plants.

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Monitoring transplant success Survival, leaf growth, primary leaf production and rhizome elongation were used to monitor transplant success. Primary leaf production was measured immediately before and after transplantation, giving a measure of a healthy meadow, before disturbance. Survival, leaf growth and rhizome elongation were measured either monthly or fortnightly (as described below) for 18 months following the methods described below:

Survival: Presence of leaves and later rhizomes with roots and green shoots was used to determine individual plug survival and success. During the monitoring period leaf presence fluctuated so that percent survival could drop to 0% and then rise.

Leaf growth: Twenty randomly selected shoots were marked by punching a small hole (18 gauge syringe needle) into the tip of the leaf sheath (see Dennison 1990b). The subsequent growth between the leaf scar and the sheath scar gave a measure of daily leaf growth.

Primary leaf production: The mean number of leaves per 20 random shoots, the mean length of leaves and leaf densities were monitored fortnightly (see Dennison 1990b). Equation 1 from Pergent-Martini et al. (1994) was used to compare temporal and spatial production of the primary leaf: PI = N x L x D Equation 1 where N = mean number of leaves, L = mean length of the leaves and D = leaf density.

Rhizome extension: Rhizome extension rates were measured monthly when rhizomes extended beyond the 15 cm plug diameter. Extending rhizomes were tagged and measured as described in Chapter 2.

4.2.4 Statistical analyses Three categories were established to compare changes that ASG mats and transplants might have on the habitat. These categories are 1) before ASG mats were placed on the transect (September to November 1993), 2) after ASG mats were placed on the transect but before transplants were made (December 1993 to April 1994), 3) after ASG mats and transplants were made (May 1994 to June 1995). Temporal data was analysed using RM one way ANOVA. Repeated measure one way ANOVA non-parametric analysis was done using a

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Friedman’s RM ANOVA test (Zar 1996). Spatial data was analysed using t-tests, with non- parametric analysis using a Mann-Whitney test (Zar 1996).

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4.3 Results 4.3.1 Habitat enhancement Sediment grain size The addition of ASG mats to the transect saw significant changes in the sediment composition. ASG mats appeared to stabilise the sediment composition over time, whilst in

2 the non-ASG areas the sediment composition fluctuated significantly over time (χ [29] = 112, p < 0.05). ASG mats, before seagrass transplants were in place, encouraged the deposition of silt and coarse materials, whilst non-ASG areas accumulated fine particles. With the addition of transplants the composition changed, with significantly more coarse material collecting in

2 the non-ASG area (χ [14] = 182, p < 0.05). Thus, the sediment grain size composition was significantly different between the ASG and non-ASG areas at each site both after placement of ASG mats and after seagrass transplantation (Figure 4.2).

Prior to ASG mat placement on the transect, there were significant differences in sediment

2 composition between sites (χ [29] = 67.5, p < 0.05), indicating differences with depth, however, no trend was detected. After the addition of ASG mats and seagrass transplants to

2 non-ASG areas, the sediment composition between sites still differed significantly (χ [29] =

2 88.5, p < 0.05; χ [29] = 84, p < 0.05) becoming sparse with depth.

Erosion and accretion Sites 1, 2 and 4 were accreting sediment before the placement of ASG mats on site, whereas sites 5 and 3 were eroding (Figure 4.2). The amount of fluctuation in sediment height observed at each site was similar. There was no apparent trend in sediment height movement before ASG mats were placed on site with depth. After placement, the ASG mats significantly enhanced accretion and erosion (F[1,6] = 1.03, p < 0.05) at all sites (Figure 4.2). Accretion and erosion occurred at the same time (followed a similar pattern) in the ASG mat and non-ASG areas (Figure 4.2) and did not appear to be depth dependent.

After seagrass transplants were added to the sites the amount of accretion or erosion was still

2 significantly enhanced in the ASG mats (χ [9] = 30.8, p < 0.05) and the pattern of accretion and erosion were similar (Figure 4.2). No trend was apparent with depth and accretion or erosion.

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a) 15

-5 Sediment height (cm) height Sediment

-10 Before ASG After ASG mat mat placement placement & before After ASG mat placement & transplantation transplantation & after transplantation -15 3 3 4 4 4 4 4 4 4 4 5 5 9 9 9 9 9 95 9 t 93 94 l 9 t 94 95 p 93 v 9 c n b pr 9 n u g 9 v 9 c b n e Ja e J u o e u S Oc No De F Mar A May 94 Ju A Sep 94 Oc N De Jan 95 F Mar Apr 9 May 95 J

Figure 4.2a: The amount of substrate erosion and accretion occurring before in ASG mats (() and non-ASG areas ()) at site 1 on the recipient transect, Success Bank in Western Australia.

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b) 15

-5 Sediment height (cm) height Sediment

-10 Before ASG After ASG mat mat placement placement & After ASG mat placement & transplantation before transplantation & after transplantation -15 3 3 4 4 4 4 4 4 5 5 5 9 9 9 9 94 9 9 9 9 93 r 94 l 94 v 93 c n r 9 p n g 94 p v 94 c r 9 n ay Ju u e o Jan 95 ay u Sep 9 Oct No De Ja Feb 94 Ma A M Ju A S Oct N De Feb 95 Ma Apr 95 M J

Figure 4.2b: The amount of substrate erosion and accretion occurring before in ASG mats (() and non-ASG areas ()) at site 2 on the recipient transect, Success Bank in Western Australia.

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c) 15

0 Sediment height (cm) height Sediment

-5 Before ASG After ASG mat mat placement placement & before After ASG mat placement & transplantation transplantation & after transplantation -10 3 3 4 4 4 4 4 4 4 4 4 5 5 5 9 9 9 9 9 9 95 9 9 t 93 l 9 t 94 v 9 c n pr 9 n u g 9 p 94 v 9 c n Ja J u e o u Sep 93 Oc No De Feb Mar A May 94 Ju A S Oc N De Jan 95 Feb Mar Apr 9 May 95 J

Figure 4.2c: The amount of substrate erosion and accretion occurring before in ASG mats (() and non-ASG areas ()) at site 3 on the recipient transect, Success Bank in Western Australia.

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d) 12

-2 Sediment height (cm) height Sediment -4

-6 Before ASG After ASG mat -8 mat placement placement & before After ASG mat placement & transplantation transplantation & after transplantation -10 3 3 4 4 4 4 4 4 5 5 9 93 9 9 9 9 9 94 9 9 t v 93 c n r 94 n l 9 g 94 p t v 94 c n p Ju u e o u Sep 9 Oc No De Ja Feb 94 Mar A May 94 Ju A S Oc N De Jan 95 Feb 95 Mar Apr 95 May 95 J

Figure 4.2d: The amount of substrate erosion and accretion occurring before in ASG mats (() and non-ASG areas ()) at site 4 on the recipient transect, Success Bank in Western Australia.

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e) 12

-2 Sediment height (cm) height Sediment -4

-6 Before ASG After ASG mat -8 mat placement placement & before After ASG mat placement & transplantation transplantation & after transplantation -10 3 3 4 4 4 4 4 4 4 4 4 5 5 5 9 9 9 9 9 9 95 9 9 t 93 r 9 l 9 t 94 v 9 c n p n u g 9 p 94 v 9 c n Ja J u e o u Sep 93 Oc No De Feb Mar A May 94 Ju A S Oc N De Jan 95 Feb Mar Apr 9 May 95 J

Figure 4.2e: The amount of substrate erosion and accretion occurring before in ASG mats (() and non-ASG areas ()) at site 5 on the recipient transect, Success Bank in Western Australia.

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4.3.2 Seagrass transplantation Plug survival Seagrass transplant success varied between depths and between ASG presence or absence. No trend was detected in transplant survival with depth although plug success varied. Deeper sites (5 and 4) had similar survival rates, peaking at comparable levels to the shallowest site (1). Survival levels appeared fairly stable at depth (Figure 4.3). Survival in both ASG and non-ASG plots fluctuated as depth increased.

2 ASG mats significantly increased plug survival (χ [14] = 53.6, p < 0.05). Plug survival was higher in

2 2 2 ASG mats at site 5 (χ [5] = 28.8, p < 0.05), site 4 (χ [5] = 18.4, p < 0.05) and site 1 (χ [5] = 17.7, p <

2 2 0.05) but lower at site 2 (χ [5] = 0.70, p <0.05) and site 3 (χ [5] = 18.1, p < 0.05). Plugs in non-ASG mat areas of sites 1, 4 and 5 did not recover from their initial decline, having 0% survival by the end of the monitoring period (Figure 4.3). In contrast, transplants at sites 2 and 3 had similar patterns of survivorship in ASG and non-ASG mats areas, until the final four months (Figure 4.3).

Leaf growth Leaf growth at all sites decreased and then increased at similar rates (Figure 4.4a). Differences in

2 leaf growth, were insignificant over depth (χ [29] = 0.013, p > 0.05; Figure 4.4). With the exception

2 of August 1995 (χ [4] = 19.1, p < 0.05; Figure 4.4b), leaf growth was similar over time, with no

2 differences between the seasons studied (χ [29] = 16.3, p > 0.05).

Primary leaf production Primary leaf production was not significantly different between sites, with the exception of site 2,

2 which was significantly lower than other sites (χ [29] = 15.8, p < 0.05; Figure 4.5a). Over time, there was a significant difference in primary leaf growth with the months of May 1994, October 1994 and

2 October 1995 producing more primary leaves than other months (χ [29] = 36.0, p < 0.05, Figure 4.5b). Yet, there was no clear trend in primary leaf growth over depth or season (Figure 4.5b).

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100 a) 80 60 40 20 0 100 b) 80 60 40 20 0 100 c) 80 60 40

Survival (%) Survival 20 0 100 d) 80 60 40 20 0 100 e) 80 60 40 20 0 MJ JASONDJ FM 1993 - 1994

Figure 4.3: Percentage survival of Posidonia australis seagrass transplants in ASG mats (() and non-ASG areas (") on the 5 sites on Success Bank, Western Australia: a) site 1, b) site 2, c) site 3, d) site 4, and e) site 5.

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1.2 a) 1.0 0.8 0.6 0.4 0.2 Leaf growth (cm) growth Leaf 0.0 SONFAJAO Month

1.2 b) 1.0 0.8 0.6 0.4 0.2 Leaf growth (cm) growth Leaf 0.0 0123456 Site

Figure 4.4: Posidonia australis transplant leaf growth on Success Bank in Western Australia at: a) site 1 ()), site 2 ("), site 3 (h), site 4 (’) and site 5 (v); and b) during October 1994 ()), December 1994 ("), February 1995 (h), April 1995 (’), June 1995 (J), August 1995 (() and October 1995 (!).

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45 40 a) 35 30 25 20

(x 1000) 15 10 5 0 Primary leaf production production leaf Primary MJAOJMAJAO Month

45 40 b) 35 30 25 20

(x 1000) 15 10 5 0 Primary leaf production production leaf Primary 12345 Site

Figure 4.5: Primary leaf production of Posidonia australis seagrass transplants on Success Bank in Western Australia at: a) site 1 ()), site 2 ("), site 3 (h), site 4 (’) and site 5 (v); and b) during May 1994 ()), June 1994 ("), August 1994 (h), October 1994 (’), January 1995 (J), March 1994 ((), April 1995 (!), June 1995 (f), August 1995 (&) and October 1995 (q).

Rhizome elongation Rhizome elongation in the seagrass transplants only occurred at site 1. At this site, 39% of transplants in ASG areas had rhizomes that elongated, at a mean rate of 3.2 ± 0.11 mm d-1 during April and May (1,168 mm y-1). Of the transplants made into non-ASG areas, no rhizomes elongated.

4.4 Discussion

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The feasibility of transplanting Posidonia australis into sites of differing depth (and therefore light) and energy regimes was examined. In addition the effect of habitat enhancement using ASG mats to stabilise the substratum was investigated. Success Bank was identified as the transplant site meeting the Stage I (site selection) criteria (Chapter 2, Table 2.7) and plugs with orthotropic rhizomes were selected as the transplant unit with the greatest likelihood of success (Chapter 3, Stage II [transplant unit and technique] criteria). The aims of this study were to establish the feasibility of transplant efforts and identify a cost-effective transplantation method with greatest success for P. australis (and other Posidonia species). It was anticipated that a method which accelerates recolonisation, revegetation and meadow expansion rates, as well as being feasible for both small and large-scale transplant attempts, would be necessary to meet these ends.

4.4.1 Habitat enhancement Artificial seagrass mats altered their immediate environment: they significantly increased both erosion and accretion levels and significantly changed the sediment composition. The sediment grain size composition reflects a stabilisation of the environment, with the composition fluctuating less with time. The increased erosion and accretion infers a less stable environment, however levels of erosion and accretion may counterbalance each other, resolving transplant suffocation problems that are seen in mesh. Thus, habitat enhancement did occur; if this enhancement is beneficial to seagrass transplants is discussed below.

Sediment grain size Nelson (1992) documented the capacity for ASG mats to reduce sediment fluctuation and increase fine material. In this study, ASG mats created an environment that was calmer (determined by sediment grain size) than non-ASG areas. This was reflected in the finer sediment grain size composition of ASG mat plots with transplants, and the reduced fluctuation in sediment composition. The link between sediment composition and energy regime is well established (Eckman 1983; Nielsen 1992; Gambi et al. 1990; Nielsen 1992). In rough water a substantial layer of sediment is in motion with the sediment composition of these regions being predominantly large particles (Nielsen 1992), as seen in the non-ASG areas. The creation of a calm environment in ASG mats result from a reduction of water velocity through the baffling effect of “leaves” acting as a barrier to entrained particles (Fonseca et al. 1982; Fonseca and Kenworthy 1987; Gambi et al. 1990). This results in a sheltered habitat with finer particles and a more stable sediment composition, which should increase survival of seagrass transplants (Gerard and Mann 1979;

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Fonseca et al. 1982; Fonseca et al. 1983; Fonseca and Fisher 1986; Gambi et al. 1990; Nelson 1992; Walker 1994).

The sediment composition was significantly different between sites, but did not follow a trend with depth. Generally, deeper areas are calmer resulting in a higher proportion of fine particles (Hughes 1991; Nielsen 1992). However, in this study no such trend occurred, inferring that deep sites were not calmer than shallow sites. A number of reasons may explain this. The wave harmonics, for example, on Success Bank may be such that the transect is influenced by both surface and long waves, creating high water motion at both deep and shallow sites (Sverdrup et al. 1942; Komar 1989; Neilsen 1992). Another possibility explaining the lack of fine particles relates to the increased dissolution rates of calcium carbonate in cold (< 20°C) seawater (Sverdrup et al. 1942; Borowitzka 1977). This trend is reflected in the limited distribution of many marine organisms such as corals (restricted to warm water; e.g. Fagerstrom 1987; Hughes 1991; Pearson 1995; Valiela 1995), calcareous algae (generally restricted to warm water; e.g. Borowitzka 1977; Pearson 1995) and siliceous diatoms (generally restricted to temperate water; e.g. Sverdrup et al. 1942; Hasle 1976; Pearson 1995; Valiela 1995). On Success Bank the scarcity of fine particles may be due to dissolution rates because the sediments have an extremely high calcium carbonate content (R. Wilson pers. comm.) and the water temperature ranges between 13-22°C. A further reason for the lack of fine particles at depth, which is purely speculative, is the possibility that an historic disturbance (such as a 100 year cyclone) may have removed a large proportion of the fines, thus depleting fines from the region.

Thus, ASG mats dampen water velocity, creating a calmer environment for seagrass transplants. No trend existed between sediment composition and depth, although the composition differed significantly between sites.

Erosion and accretion The pattern of accretion and erosion did not differ between ASG mats and non-ASG areas, although accretion and erosion both increased significantly in the ASG mats (Figure 4.2). An increase in accretion levels within an ASG mat was expected because they act as a barrier to particles and alter water flow (see 4.1.1 above; Merkel 1990a; Nelson 1992; Walker 1994). This has been analysed in previous, similar studies (Fonseca et al. 1982; Fonseca and Fischer 1986; Fonseca and Kenworthy 1987; Gambi et al. 1990). One other study has used ASG mats as a stabilising agent when transplanting seagrasses (West et al. 1990). That study was inconclusive due to the premature

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end of the experiment resulting from a large storm disturbance. Mesh has more frequently been used and has had mixed results (Merkel 1990a; Walker 1994). For example, Merkel (1990a) used mesh to stabilise an area and increase protection from foraging stingrays, resulting in 100% survival of transplants. Mesh is not always successful however as Walker (1994) found when large amounts of accretion led to transplant suffocation.

The increases in accretion and erosion within the ASG mats may be due to the subsequent ‘meadow form’ the ASG mats and transplants create. Typically, erosion increases on meadow edges and within meadow patches because of accelerated flow around the meadow (Gambi et al. 1990). This acceleration of flow and the baffling of currents entering meadows produce distinct meadow shapes, as discussed by Fonseca et al. (1983). Both plugs and ASG mats resemble a patchy meadow similar to those expected in high to mid current strength areas (Fonseca et al. 1983), with a mound shaped discontinuous vegetative cover (Fonseca et al. 1983; Fonseca and Kenworthy 1987). The mound shape tends to enhance turbulence (Fonseca et al. 1982; Fonseca et al. 1983; Fonseca and Kenworthy 1987), increasing both sediment accretion on the leading edge and erosion on the trailing edge and altering sediment grain size composition. In this study the ASG mats act as small artificial meadows that have increased both accretion and erosion, as predicted (Figure 4.2). By increasing both accretion and erosion, the ASG mats counterbalance the sediment fluxes, stopping burial or bare-rooting of transplants.

It is generally thought that deeper water is calmer, as discussed above (see the sediment grain size composition discussion). This did not occur on this transect (Figure 4.2). Considering both the sediment grain size composition and erosion and accretion levels, it seems likely that the Success Bank transect is influenced by a combination of surface and long waves, which act to create a similar levels of turbulence along the entire study transect.

Thus, the ASG mats created a different environment (increased both accretion and erosion levels) than the non-ASG areas. Again, there was a lack of trend between in depth and erosion and accretion levels.

4.4.2 Seagrass transplantation Transplant survival Survival of plugs, exceeded those reported for Posidonia species in other studies (see Kirkman 1989; LeProvost Environmental Consultants 1990; West et al. 1990; Hancock 1992; Paling 1992; Walker

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1994; Kirkman 1995; Lord et al. 1999; see also Appendix A). Survival in previous studies was hampered by a number of factors, such as poor site selection with low irradiance, high energy regimes and poor transplant unit choice. As reviewed in Chapter 1, transplant survival in other studies has been mixed (Appendix A; Lord et al. 1999). In this study, plug survival generally increased within ASG mats suggesting that the presence of seagrass structure (either artificial or natural) enhances plug survival (Figure 4.3). Site 1 for example, was affected by the May 1994 storms and plugs in both ASG mats and non-ASG plots recovered after three months (Figure 4.3). This rapid recovery was possibly aided by the close proximity of a naturally established seagrass meadow. This was the only site that had a healthy seagrass meadow present prior to beginning experimental work. The presence of the natural meadow surrounding the transplants may have acted to effectively increase the patch size, hence reducing currents and turbulence. At other sites, where meadow presence was reduced or absent, the majority of transplants in non-ASG areas died after the May storms, resulting in significant differences between ASG and non-ASG survival.

A number of factors affected plug survival. First, severe storms hit the study coastline in late May 1994 (< 1 month after the transplant attempt). These storms caused massive damage, with new blowouts forming and ASG mats in shallow sites (1, 2 and 3) sustaining damage. Storm disturbance is a natural phenomenon that has been documented to destroy ASG mats and dislodge transplants (see West et al. 1990). Second, anthropogenic disturbances also affected plug survival. At site 3, boat anchors (pers. obs.) destroyed half the ASG mat and dislodged 31 (86%) transplants. The damage to the ASG mat itself reduced any benefit that may have been received. Thus, the similarity between ASG and non-ASG plug survival at this site (Figure 4.3) is likely due to this disturbance. Similarly, site 2 became a blowout during the May 1994 storms. A blowout is an unvegetated depression within a meadow that has an eroding seaward edge and an accreting shoreward edge (Patriquin 1975; Clarke and Kirkman 1989). The storms weakened the ASG mat, which helped scour the immediate area prior to re-bracing which assisted in the creation of the blowout that increased in size over time. Within the blowout turbulence is high, creating a mobile substratum. Thus, any benefit the transplant may have received from the ASG mat was lost when the area became a blowout. It is of interest that approximately 25% transplants in the non-ASG areas survived (Figure 4.3). This supports the theory that damaged ASG mats actively contributed to blowout formation.

Depth also has the potential to affect plug survival (Molenaar and Meinesz 1992). It has been demonstrated that transplants derived from donor meadows shallower than the transplant site usually have lower survival rates, and growth and development of shoots is reduced (West 1990;

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Molenaar and Meinesz 1992). In this study, transplants into sites deeper than the donor meadow (sites 5 and 4) managed to survive relatively well (Figure 4.3). This result may be due to similar light levels recorded at the donor and the deep transplant sites (donor: ~94 µmol m-2 s-1 and site 5: 112 µmol m-2 s-1; measured using Dataflow Systems Pty Ltd. light loggers). Light attenuation was greater at the donor meadow due to sea sand dredging operations proximity and even though the donor meadow was shallower than sites 5 and 4, recorded irradiance was not lower. Although irradiance levels decreased with depth, irradiance was still above the compensation levels reported by Masini et al. (1990) for southwestern Australian P. australis populations and the surface irradiance (SI) at all sites was above 25% SI. Therefore, growth and survival was expected to continue at all the depths examined on the transect. Typically, light is considered to be the primary limiting factor of seagrass growth and distribution (Dennison and Alberte 1985; Orth and Moore 1988). Many studies have reported transplant failures due to light attenuation. For example, increased turbidity (either by animals or anthropogenic effects) leading to increased light attenuation is a major cause of transplant failure in the US (Thorhaug 1974, 1983, 1985; Merkel 1990a; Zimmerman et al. 1995; Ruckelshaus and Hay 1998; see Chapter 1). Light attenuation has also been linked with transplant failure in Australia (Clarke and Kirkman 1989; Kirkman 1989, 1995; Lord et al. 1999).

Forty two percent of plugs survived the experimental period (Figure 4.3), with ASG mats significantly increasing plug success. There were a number of other factors influencing survival such as natural and anthropogenic disturbances, however, light intensity was found to be insignificant.

Leaf growth The presence of ASG mats had no significant effect on leaf growth rates, inferring that primary production was unaffected. Leaf growth rates were similar at all sites (Figure 4.4), suggesting a correlation between depth and leaf growth. Responses in leaf growth are triggered by a combination of environmental factors such as irradiance, nutrients, temperature and photoperiod (Raven et al. 1986; Küppers et al. 1988). None of the factors were altered by the presence of ASG mats, nor were they changed across the transect with the exception of light levels (unpub. data.). Furthermore, Küppers et al. (1988) discussed the insignificance of photosynthetic characteristics for plant growth when environmental conditions are intermediate. This may explain why, in this study leaf growth was unaffected by depth (i.e. irradiance) or photoperiod.

Primary leaf production

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Neither depth nor season significantly affected primary leaf production and thus primary productivity in this study (Figure 4.4). An exception was site 2, which had significantly lower production rates than other sites. This was probably due to the disturbance at this site, which reduced shoot density and thus productivity. It has been hypothesised by other researchers that primary leaf production can vary widely with depth, because meadow density controls the primary leaf production (Vermaat et al. 1993; Pergent-Martini et al. 1994). Typically, meadow density decreases with depth. Meadow densities correlate with decreased primary leaf production, hypothesised to be due to a reduction in light levels that triggers maximisation of light harvesting ability. To maximise light harvesting at depth plants may have a small rhizome, reduced leaf density and reduced clustering, thus reducing the possibility of self-shading (Bulthuis 1983; Josselyn et al. 1986; West 1990). This is seen in species such as Heterozostera tasmanica (Bulthuis 1983; pers. obs.) and Halophila ovalis (pers. obs.). In an artificial situation (such as transplanted plugs), the densities are similar at each site and are initially controlled by the researcher and thus, no reduction in primary leaf production would be observed. A discussion of the lack of influence depth and photoperiod have had on primary production (leaf growth, primary leaf production and rhizome elongation) is provided below.

Rhizome elongation Rhizome elongation was found to be independent of depth, with rhizomes at most sites (sites 5, 4 and 3) failing to elongate. Irradiance on the Success Bank transect decreased with depth (on average 321 µmol m-2 s-1 to 112 µmol m-2 s-1), and growth occurred at site 1 but not at site 2, although both sites were at similar depth (4.9 and 5.8m, respectively). Also, contrary to expected responses, the natural meadow on Success Bank, reinforced the observations that rhizome growth was independent of depth (see Chapter 3). This may be because irradiance at all sites on this transect was above reported compensation (25 µmol m-2 s-1) and saturation (73 µmol m-2 s-1) levels.

However, the reported Isat and Ic values are based on data from Albany (Masini et al. 1990), not Success Bank where this study was undertaken. Therefore, surface irradiance ranges will be used to discuss the potential of light attenuation as a limiting factor in rhizome growth.

Chapter 2 determined that Posidonia australis is capable of growth when irradiance levels are between 5 and 25% SI. Irradiance levels at all of the sites on the Success Bank transect, remained above the 25% SI (102 µmol m-2 s-1) level, with the deepest sites (site 5) irradiance being equivalent to 27.5% SI during summer. Using surface and 12m irradiance on the transect, an extinction coefficient for the body of water was calculated (k = 0.065) from equation 2:

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–kd Id = I0 e Equation 2

where Id is the light intensity travelling a distance of (d), I0 the incoming light intensity and k is the extinction coefficient (from Parsons et al. 1977).

This extinction coefficient is comparable to that experienced in clearest oceanic water (k = 0.06, Parsons et al. 1977). Thus, based on the log/linear association between depth and clearest oceanic water light intensity, SI less than 5% (22 µmol m-2s-1) requires a depth of ~27.5m, which is not available in the surrounding regions such as Success Bank (5-6m), Owen Anchorage (16m) or Cockburn Sound (22m) from RAN chart data (Aus 117). Thus, there is no indication that light attenuation is a limiting factor at the depths examined, especially considering that primary production (see leaf growth and primary leaf production above) continued, although there was no observed rhizome growth at depth. This study does not support the theory that growth decreases with depth because of a reduction in irradiance (see Drew 1979; Dennison 1987; Williams 1988; Dawes and Tomasko 1988; Salisbury and Ross 1992).

A number of other factors are likely to influence growth: collection methods, growth type of the collected rhizomes, measurement techniques, and the hydrodynamic processes. The task of plug collection was allocated to the Navy clearance divers. Quality control measures involved directing divers to the removal area (meadow edge), instructing them on the growth mode (plagiotropic rhizomes) and the species (Posidonia australis) to be removed. None-the-less, quality may have been improved by using divers with a biological/botanical background.

Rhizomes grow in two forms: orthotropic or plagiotropic. If the divers collected the incorrect form (orthotropic instead of plagiotropic) there may have been a bias that affected growth rates. For example, more orthotropic rhizomes may have been collected if the plugs were removed from the middle of the meadow instead of the growing edges. If this were the case the rhizomes may assume a plagiotropic form to grow horizontally and colonise the surrounding bare area (Caye and Meinesz 1985a). This has been documented in meadows that have similar densities to the transplant plugs (Caye and Meinesz 1985a). Modification of growth type may involve a lag period (Meinesz et al. 1991) and hence a slow growth rate until orientation is completed. However, after re-orientation rhizome growth rates tend to increase, not decrease (Caye 1980; Meinesz et al. 1991). Similarly, Fonseca (1997) has surmised that transplant material should be selected from growing meadow edges because the plants here exhibit a guerilla growth strategy. Generally, guerilla strategists exhibit a higher rate of vegetative expansion, capable of colonising new areas (Lovett-Doust 1981).

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Alternatively, a lack of growth may occur if meristems are missing or damaged. The methods employed to collect transplants are destructive and thus some damage to the meristems is expected although the size of the plug is hoped to overcome this damage. Walker (1994) examined plug size, alluding that plugs of > 15cm diameter are more likely to survive transplantation because of a correlation between increased plug size and an increase in leaf base number and below ground biomass. Also, larger plugs experience fewer edge effects (Fonseca et al. 1982; Fonseca et al. 1983; Thayer et al. 1984; Duarte and Sand-Jensen 1990a; Olesen and Sand-Jensen 1994) and are likely to directly influence surrounding hydrodynamics.

Molenaar and Meinesz (1992) and Meinesz et al. (1992) discuss an apparent lack of growth in transplants due to growth of the terminal end being countered by necrosis at the basal region (see Figure 3.4, Chapter 3). This idea has been discussed in detail in Chapters 2 and 3, where it was decided to measure growth increments, not total length of rhizomes, thus overcoming this problem. Other researchers have opted to measure rhizome growth via patch movement (Duarte and Sand-Jensen 1990; Olesen and Sand-Jensen 1994) or using plastochrone interval (Brouns 1985, 1987; Pergent and Pergent-Martini 1990; Gallegos et al. 1993; Molenaar et al. 1993; Duarte et al. 1994; Gallegos et al. 1994; Marbà et al. 1994). Thus, because of the method used to measure rhizome elongation, it is unlikely that lack of growth was affected by the cancellation of growth through necrosis.

In a calmer environment, rhizomes are able to extend without the threat of burial or disturbance from strong currents and tend to elongate at a slower rate (Fonseca et al. 1983; Meinesz et al. 1991). Whereas in high-energy regimes, meadows are patchy, mounded and often unable to spread successfully (primarily because they have more vertical rhizomes; see Fonseca and Bell 1998). Although, the stress of burial in high-energy regime, often stimulates growth rates (Caye and Meinesz 1985; Littler et al. 1983; Baron et al. 1993; Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997; see also Chapter 3). For example, Thalassia testudinum responds to moderate burial by increasing vertical growth, which relocates meristems to the surface (Marbà et al. 1994). Yet in high-energy regimes, plants concentrate their efforts on recolonising areas and thus little extension of growth may be observed (Littler et al. 1983; Baron et al. 1993; Marbà and Duarte 1994; Marbà et al. 1994; Duarte et al. 1997; see also Chapter 3). Results already discussed indicate that the transect is in a mid to high current environment and thus rhizomes are continuously trying to maintain a presence, instead of increasing their area. This type of pattern occurs in natural

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environments, where fragmented meadows have a predominantly vertical relief and have a reduced ability to spread and colonise new areas (Fonseca and Bell 1998). This may explain why little growth was observed on the Success Bank transect, with the exception of site 1.

At site 1 rhizome elongation (1,168 ± 40.15 mm y-1) significantly exceeded that recorded for field rhizome growth (transplants 1,168 ± 40.15 mm y-1; field rhizome 379.6 ± 0.15 mm y-1; see Chapter 3). Irradiance at this site exceeds the 5% SI, and a healthy meadow was present, suggesting no limitations on growth. The shape of the meadow at this site also indicates potential for successful growth. The natural, established meadow shape at site 1 is characteristic of a low to mid current environment; it is almost continuous in extent, with some blowouts (Patriquin 1975; Fonseca et al. 1983). The ASG mats with transplants shape is characteristic of a mid to high current environment (e.g. a mounded patch). It is conceivable that the surrounding meadow at site 1 moderated the current, making the environment slightly more sheltered which enabled rhizome elongation. Whilst at other sites, the lack of an established meadow prevented any moderation in current strength and yet the size was still enough to create turbulence, which increased scour and retarded spread. However, as discussed in the stabilisation section above, the energy regime of the area was not significantly different over the transect. Therefore, it is improbable that the energy regime affected the transplant rhizome growth.

Rhizome growth only occurred in plugs transplanted into ASG mats, suggesting that ASG mats conferred an advantage on plug rhizome elongation. Once again, the indication is that the ASG mats reduce current flow, creating a calmer environment, more conducive the rhizome elongation. The presence of the ASG mat effectively increases the size of the plug, thus reducing edge effects on the actual plug. As discussed above, the presence of a meadow, whether artificial (such as an ASG mat) or natural, may directly influence currents and turbulence. Typically as the size of meadow increases, erosion is reduced and accretion increases, which improves growing conditions (Fonseca et al. 1982; Fonseca et al. 1983; Thayer et al. 1984; Duarte and Sand-Jensen 1990a, 1990b). This has been documented in coalescing patches of Cymodocea nodosa (Duarte and Sand- Jensen 1990a). This infers that the ASG mats created an environment that is more favourable to rhizome growth, reiterating that ASG mats enhance a transplant site. There are no published studies on Australian species patch formation and coalescence. More research in this area will help elucidate the role of hydrodynamics within meadows and patches in the survival, spread and maintenance of meadows and transplants.

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4.4.3 Conclusions: success of enhancement and transplantation ASG mats enhanced the sites and increased survival and rhizome elongation. However, for the effort they required to produce and place on site, a higher rhizome elongation rate would be needed to ensure wide-spread utility. Using the definition developed in Chapter 1 (> 50% survival and > 50% expansion), this transplant attempt failed as a whole (42% survival and 39% expanded). Even when reviewed individually no sites met the criteria of success. Site 1 had 50% of plugs surviving, yet only 39% expanded beyond the original 15cm diameter plug size. Other sites, ranged from 50% survival but no expansion, to 0% survival (Figure 4.3). Although the attempt failed, several important points were developed. First, transplants can survive in deep water but rhizome elongation may not occur. Survival can be hampered in shallow water by natural (increased surge and swell action; increased epiphyte loading due to increased light) and anthropogenic (boat anchors, fishing lines and dumping of refuse) effects. Shallow sites however, are more favourable for transplant sites because rhizome elongation was detected. Primary production was significantly different between sites, although no significant effect of depth was detected suggesting that light attenuation was not a limiting factor in the depths examined on this transect. Additional research into the patch dynamics of Australian seagrasses is necessary to understand the influence that meadow form has on mitigating environmental factors on rhizome elongation.

1) Stage I (site selection) a) Irradiance levels above 5% SI; b) Water quality needs to be high and not eutrophied; c) Short burial periods with cyclical burial periods preferred, which infers that; d) Water movement should be low or reduced; and e) Substrate type can either be limestone rubble or coarse sand, with sand a better choice due to ease of transplanting and anchoring ability. 2) Stage II (transplant unit and technique) a) Vegetative propagules are better transplant units than sexual propagules; i) Plugs are preferred vegetative transplant units over sprigs; b) Plugs should be spaced in close proximity, however a trade-off between plug spacing and overall area should be evaluated (further investigations are recommended);

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c) Handling should be minimised; i) The donor meadow should be in close proximity to the transplant site to reduce handling time; d) Larger transplant units are preferred (> 15 cm diameter); and e) Donor meadows should be in deeper water than the transplant site (further investigation is recommended). 3) Stage III (habitat enhancement) a) Reducing water flow and increasing accretion by using ASG mats is preferred; and b) Transplants should be made into stabilised sites (such as ASG mats).

Thus, with regards to the working hypotheses the evidence supports the tenet that ASG mats enhance habitats, which increased two of the measures of success (survival and rhizome elongation). There is evidence to support the hypothesis that increased water motion negatively affects transplant survival. This is partly due to increased burial disturbance. The evidence does not support the dogmatic hypothesis that depth and light limit transplant success. No significant difference was detected in rhizome growth rates or transplant unit survival in relation to depth or light. The depths observed in this experiment were not beyond the previously measured compensation depth of Posidonia australis, but had been assumed to limit seagrass distribution. This study demonstrated that P. australis transplants can survive and grow at depth and similarly, transplants from shallow donor sites can succeed at deep restoration sites. Additional research is needed to identify the limitations on transplant success.

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Chapter 5 General Discussion: the development of a decision-based framework for restoration planning

5.1 Introduction The destruction of seagrass meadows via anthropogenic and biological disturbance generally leads to the loss of habitat. This loss can create environmental conditions, such as light attenuation, sediment instability, and sediment consistency and distribution, which interact to diminish the success of seagrass transplantation. The decline of seagrass meadows and the failure of subsequent transplantation attempts thus, triggers a shift in community structure and dynamics within a seagrass ecosystem (see Chapter 1). Such changes have been documented worldwide and can influence the ecology of a region (e.g. change from a vegetated to a sandy substratum) and fisheries (e.g. prawn fisheries in Gulf of Carpentaria). In many instances the loss of seagrass and the community shift is dramatic, resulting in the examination and reporting of the event(s) (den Hartog and Polderman 1975; Thayer et al. 1975; Chiffings and McComb 1983; Cambridge and McComb 1984; Neverauskas 1985a, 1985b; Bastyan 1986; Cambridge et al. 1986; Silberstein et al. 1986; den Hartog 1987; Shepherd et al. 1989; Clarke and Kirkman 1989; Larkum and West 1990; Walker and McComb 1992; Paling 1995; Ruckelshaus and Hays 1998). As discussed in Chapter 1, the continual loss of seagrasses has directed research into restoration and mitigation.

In some countries, there has been a push for ‘no net loss’ in marine developments. In legislative terms, this becomes an equivalency policy. In such legislation it is recognised that a return to pre-existing conditions is unlikely, however there must be a return of the ecosystem to a close approximation of its conditions prior to disturbance (USNRC 1992; Gordon 1995; Ruckelshaus and Hays 1998).

Human populations are expanding at great rates and with this expansion comes habitat loss and fragmentation, increasing pollution, excessive consumption and decreases in biological diversity (see Chapter 1). The need for development is unopposed yet it needs to be balanced against environmental issues. With this balance, there is recognition that development and hence anthropogenic disturbance will continue to occur. Disturbance often brings degradation and if degradation continues, then the question moves into the biological arena and the difficulties are; is transplantation/mitigation theoretically possible and feasible.

Restoration and mitigation of seagrass ecosystems is often relegated to the “too hard basket”. For example, in the United States (U.S.), one of the funding bodies for mitigation projects (NMFS) will not accept seagrass mitigation in exchange for habitat destruction, instead, mitigation banking occurs (Ruckelshaus and Hays 1998). Although in some regions of the U.S. (e.g. Pacific northwest) mitigation banking is not allowed (S. Wyllie-Echeverria pers. comm.). Why is seagrass mitigation and restoration untenable? The truth lies in the success rates of transplantation efforts. In the U.S., where the vast majority of seagrass transplants have occurred, it is estimated that more than 50% of transplantation projects fail (Thorhaug 1986; Fonseca et al.

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1996; Fonseca 1997). Furthermore, of the attempts that succeed equivalency rarely occurs. In Australia, using the definition of transplant success defined in Chapter 1, even fewer (22%) of the published transplantation efforts have succeeded (see Appendix A). These results are reflected in European transplantation attempts.

When attempting to transplant we generally borrow from the terrestrial paradigm and look for sites that are similar to the ones the target species is already occupying (Guerrant and Pavlik 1998). If mitigation is worthwhile, the target species of seagrasses should be treated as a rare and endangered species (Schemske et al. 1994; Guerrant and Pavlik 1998), with a rigorous evaluation of the best practices to ensure mitigation success. This would involve a process similar to that outline in Chapter 1. Transplant sites should be critically examined against scientifically sound selection criteria (Figure 5.1). Limiting restoration projects to areas where seagrasses are currently found need not be the case (fundamental niche). In many instances, seagrasses are capable of utilising areas outside of their realised distribution, both in theory and in practice (see Chapter 4). Objections exist and should be empirically evaluated. The ‘habitat argument’ and the application of metapopulation dynamic theories to seagrass recruitment discussed by Fonseca (1997) should be incorporated into a broad framework.

The habitat argument is often discussed in ecological and conservation work and revolves around the concept of transplanting into areas where the species have historically been present (Guerrant and Pavlik 1998; Ruckelshaus and Hays 1998). Generally, it is perceived that success is more likely when transplanting into a habitat where the species has been known to occur. This concept ignores the larger available habitat represented by the fundamental niche of the species. A species limitation to its realised niche may occur through ecological constraints such as competition for space and resources, or physiological constraints such as depth, light or wave action. Yet, that does not mean that they cannot survive elsewhere (i.e. fundamental niche). An example is the distribution of the cirrepedia Chthamalus stellatus and Balanus balanoides (Connell 1961). Both species have overlapping fundamental niches that ranges from the high intertidal to the shallow subtidal. Chthamalus is located in the upper intertidal where its upper distribution is limited by desiccation and its lower distribution is limited by competition with Balanus (Connell 1961). Alternatively, Balanus is limited in its upper distribution by desiccation (it’s more susceptible to desiccation stress and therefore is located in the shallow subtidal). The predatory whelk Mucella limits Balanus’s lower distribution to low in the subtidal. Thus, both species have similar fundamental niches, although Chthamalus is capable of occupying deeper water, it is restricted to its realised niche through competition whereas Balanus is limited by predation.

Historically vicariance played a role in limiting the realised niche of the species, however the breakdown of biogeographic barriers (e.g. lessapsian migration and human mediated dispersal) has increased the chance of species fully attaining the fundamental niche. The development of canals through isthmuses often allow the natural dispersal of species. For example, before the opening of the Suez Canal in 1869, the realised niche of

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the mussel Brachidontes was the Red Sea. Once this body of water was opened, a natural dispersal event occurred, with Brachidontes migrating into the Mediterranean. Brachidontes realised niche expanded filling more of its fundamental niche (Ayal and Safriel 1983; Safriel and Sansson-Frostig 1988). The opening of the Suez Canal has contributed to the introduction of the seagrass Halophila stipulacea into the Mediterranean (den Hartog 1970; Clarke and den Hartog 1989). The Corinth Canal in the Mediterranean, has also enabled migration of species between the eastern Mediterranean and the Adriatic Sea (G. Rellini pers. comm. 1998).

Human mediated dispersal is an extreme example of the attainment of fundamental niche through transport between non-contiguous bioregions. For example, the oyster, Crassostrea gigas, has been transported intentionally around the world for mariculture practices (Elton 1958). Its fundamental niche was worldwide but its realised niche is its native bioregion (Northwest Pacific and East Asian Seas). Typically, C. gigas was collected for mariculture by shoveling the mudflat and thus, species associated with the oyster were also transported. This is reflected in the distribution of the introduced seagrass Zostera japonica, which maps Crassostrea’s distribution (Phillips and Shaw 1976; Carlton 1979, 1989; Harrison and Bigley 1982; Posey 1988). A lack of occupancy does not mean inability to survive in that habitat and therefore, the habitat argument is flawed because it ignores the concept of the fundamental niche.

Fonseca (1989, 1992, 1994, 1997) has argued against transplanting into realised niches, such as within meadow gaps, meadows or into currently unoccupied habitat. His argument is based upon observations of annual populations of Zostera marina in the US and is extrapolated to Posidonia meadows on Success Bank, Western Australia. Fonseca states that “a dynamic mosaic of seagrass cover means that currently unoccupied space is part of the future spatial requirement for the maintenance of the local population over time”. He considers Z. marina as an archetypal “blinking lights” species (Levins 1970); i.e. Z. marina meadows are density-dependent and regulated by patch occupancy, inferring that to remain self-sustaining Z. marina needs unoccupied habitat space to recruit into each year. This produces a classic single-species metapopulation dynamic model where, at any given time there is unoccupied habitat space (patches) and local patches blink into and out of existence as extinction and recolonisation occur (Levins 1970).

Fonseca’s extrapolation to Western Australia is based upon his observations that Posidonia meadows are a shifting landscape mosaic influenced by fast-spreading colonising species and the high frequency of seedling success. To this end, he argues that single-species metapopulation dynamics needs available space. However, as Hanski and Simberloff (1997) state “if a conservation strategy is based on metapopulation dynamics that do not exist, it can misfire”. This is the case for Posidonia meadows in Western Australia. For example, P. australis is density independent on Success Bank, Woodman Point, Cockburn Sound and Wreck Rock (seedlings are recruited close to and within the parent meadow; see Chapter 2). Therefore, it self recruits, which is demonstrated by the genetics of Posidonia populations in the Perth metropolitan region (Waycott 1995). Waycott (1995) established that Posidonia meadows are genetically diverse, suggesting recruitment

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via sexual propagules. Ruckelshaus (1994) and Ewanchuk (1995) have also provided evidence that Zostera marina is genetically differentiated, which does not support Fonseca’s argument that Z. marina fits the classic metapopulation model.

Metapopulation explosion or extinction are the only possible outcomes without density-dependence regulating the metapopulation model (Olivieri and Gouyon 1997). This does not mean that all metapopulation models are density-dependent; density-independent metapopulation models also exist (see Cipollini et al. 1994). The Posidonia meadows examined in this study are density independent and hence do not fit Fonseca’s proposed model. Thus, there is insufficient evidence to support Fonseca’s (1997) argument, with regard to transplanting Posidonia into available realised habitat space on Success Bank.

Furthermore, by using this argument, Fonseca assumes that transplanting into a habitat space saturates the metapopulation (i.e. the population is at maximum carrying capacity determined by the rate of colonisation and extinction of patches). If you transplant into a realised but empty habitat space in a saturated system the habitat space cannot be recolonised, resulting in system collapse. However, in the face of anthropogenic and natural disturbance the likelihood that a system is saturated is low. Thus, because it is a theoretically dynamic equilibrium, it is more likely that mitigation will enhance the system with decreasing payoffs as the system approaches saturation.

The process of seagrass transplantation is often recognised as complex, producing few successes. Globally, there have been many attempts to restore, revegetate and mitigate for seagrass losses yet there is still no single method that ensures success. Perhaps this is because of the dynamic nature of seagrass species, but more likely it is because there is a basic lack of knowledge about individual seagrass species autecology and therefore selection of an appropriate transplant site, transplant unit and transplant method is prone to flounder. To redress these failings, this dissertation aimed to develop a framework that identifies the key elements for transplant success. Dennison and Kirkman (1996) proposed a ‘Seagrass Survival Model’ which relied solely on light attenuation as a predictive element for site selection. In contrast, the framework outline in Chapter 1 extends beyond single element evaluations. In order to aid in identifying these elements, several lesser-known aspects of seagrass ecology were investigated, culminating in a field trial of the transplant framework (Chapter 4).

5.2 Decision-based framework for restoration planning As discussed in Chapter 1 there are three general steps in the mitigation process: planning; implementation and evaluation. This general plan can be expanded into a broader framework (Figure 5.1) that includes aspects of the restoration framework outlined in Chapter 1 (Figure 1.2). The planning process is broken down into five steps.

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Planning Objective Setting

Stage I Site Selection

Stage II Transplant Unit & Transplant Technique

Stage III Habitat Enhancement

Review of Objectives, Evaluation of Transplant Program and Preparation of Activity Plan

Proceed

Implementation Transplant / Restoration Activity

Evaluation

Monitoring Success

Figure 5.1: Broad decision-based framework for restoration planning.

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The first planning step, objective setting, begins with the identification of a target species or community (Figure 5.2). For example in this study Posidonia australis was selected as the target species. A review of the scientific literature is needed to elucidate the target species’ general biology and ecology (such as reproductive capacity, habitat preferences, light requirements, substrate preference, etc.). The key stakeholders need to be identified to aid in financing the mitigation effort as well as participate in developing the mitigation objectives (e.g. decide if habitat equivalency is needed) placing the project into context. The mitigation objectives need to consider and meet legal responsibilities or requirements (such as mitigation criteria and monitoring requirements). This step feeds back into the identification and quantification of financial constraints. Finally, broad success parameters must be stated, which consider both short and long term goals.

Objective Setting

Identify biological target (species or community)

Review scientific literature

Identify key stakeholders and legal responsibilities

Identify and quantify financial constraints

State broad success parameters in proximal (short term) and ultimate (long term)

Figure 5.2: Step 1, objective setting, in the planning process.

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Site selection is of prime importance for successful mitigation efforts (Chapter 2). A decision tree has been developed to guide the process of selecting an optimal restoration or mitigation site (Figure 5.3). The decision tree takes the potential mitigator through a number of steps designed to gather knowledge of the target species or community and highlight information gaps. When an information gap exists, an evaluation of this missing information must occur, feeding information back into the site selection process. The steps within this decision tree were demonstrated to be important limiting factors in site selection evaluation that occurred in Chapter 2. These were light requirements, water clarity and quality, water motion, depth of donor meadow and proximity of donor meadow relative to transplant site. Additional limiting factors may be identified for other species and can be readily incorporated into the decision tree.

Selecting the transplant unit and technique is the next step in the planning process (Figure 5.4). This step is highly target species dependent and involves the investigation of an appropriate transplant unit (e.g sexual or vegetative propagule; see Chapter 3). Two further decision trees (one for sexual and one for vegetative propagules) are developed in this step. To decide if sexual propagules are suitable seed production, seed viability and seedling establishment must be quantified (Figure 5.4). Similarly, to determine if a vegetative propagules are suitable, rhizome growth and density of active growing shoots must be quantified (Figure 5.4). Once a transplant unit is selected, minimisation of handling must occur (for seedlings, sprigs and plugs) and habitat enhancement (optional) should occur.

Habitat enhancement the fourth step in the planning process and is optional (Figure 5.5). Based on the outcomes of the Stage II – transplant unit and technique step, a number of transplant units are available: seeds; seedlings; sprigs; and plugs. Seeds are the least likely to succeed because they are often “lost” in the environment and/or consumed (see Chapter 1). The general technique for using seeds is broadcast dispersal. If seedlings are selected then the habitat should be modified by using either anchors, artificial barriers, mesh or ASG mats (Figure 5.5). Mesh and ASG mats are the preferred choice of habitat enhancement for seedlings because they don’t restrict the seedlings growth (like anchors) and they ensure the seedling stays on site. Similarly, if sprigs are selected as the transplant unit, then anchors, artificial barriers, mesh or ASG mats can be used. Like seedlings, the preferred habitat enhancement technique for sprigs is mesh and ASG mats (Figure 5.5). If plugs are selected, artificial barrier, mesh and ASG mats are the preferred habitat enhancement technique (Figure 5.5). Of these three methods ASG mats are preferred (see Chapters 1 and 4).

The final planning step is a review of objectives, an evaluation of the transplant program and the preparation (documentation) of an activity plan. Thus, this final process will define success and failure, and set performance indicators of success or failure. At this point a cost : benefit analysis of the various results from the site selection, transplant unit and technique and habitat enhancement decision trees can be undertaken. Finally, a monitoring program will need to be determined. The outcomes of this step feed back into objective setting to ensure that all information has been gathered and that the correct objectives have been set. Once the

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planning process is complete implementation proceeds and then evaluation occurs (Figure 5.1). Evaluation should be involve monitoring the transplant effort for at least five years, and will rely on the objective setting and review of objectives established in the planning process.

The framework was evaluated in this dissertation with Posidonia australis as the trial organism. The process of site selection was quantified by investigating the role of light, burial and handling disturbance substrate type on Posidonia australis sexual and vegetative propagules. Second, the role of sexual and vegetative propagules in the context of the implications of colonisation and restoration potential for P. australis was examined. These two objectives established a site (Success Bank) and a transplant unit (plugs) for a P. australis transplant. Finally, a transplant attempt was made, which examined the effects of depth (i.e. light limitations) and site enhancement on P. australis transplant survival, primary productivity and rhizome elongation.

5.3 Summary of findings Stage I of the framework with site selection, which was examined in Chapter 2. This chapter provides no evidence to support the popular tenet that light was a limiting factor for Posidonia australis growth at the depths observed but at light levels below 5% SI P. australis rhizome and seedling growth was reduced. Cyclical burial was best for rhizome growth and handling regimes did not appear to affect the survival of growth. Similarly, substrate type did not limit growth in either seedlings or rhizomes, but sand appeared to be a better attachment substrate. Further work on the concepts of species independence of light (particularly surface irradiance, SI, required for P. australis growth on Success Bank), disturbance and substrate type is needed. Insufficient data was collected to determine the tenet that recruitment is more successful and less patchy in stable habitats. Chapter 2 did allow a prediction of a suitable site for P. australis transplantation (Table 2.7), based on the site selection criteria in the framework.

Transplant unit selection is examined in Stage II of the framework and was investigated in Chapter 3. Few seedlings established despite Posidonia australis producing large numbers of fruits, being density independent (suggesting possible self-recruitment) and having high seedling viability. The conclusion that sexual propagules are an inadequate transplant unit was drawn from multiple lines of evidence: rates of rhizome elongation 38 cm y-1; the potential to P. australis recruit vegetative propagules existed on Success Bank; and low establishment potential of P. australis seedlings. Vegetative propagules were selected as the transplant unit for the seagrass transplant (see Chapter 4).

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Stage I - Site Selection

Light requirements No known (Ic/Isat)

Yes Undertake evaluation

In transplant No bioregion?

Yes

Can site water clarity Isat>= site >= Ic? No No be enhanced?

Yes Yes

Can site water quality Is epiphyte loading Yes (nutrient) be No Reject Site high at site? enhanced?

No Yes

Consider habitat Is water motion high? Yes enhancement

No Yes

Relative depth of Proximity of donor donor meadow meadow

Figure 5.3: Step 2, site selection (Stage I), in the planning process.

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Stage II - Transplant Unit Selection And Transplant Technique

Sexual Propagule Vegetative Propagule

Is seed production No Is rhizome growth high? No rate high?

Yes Reject sexual Is enhancement propagule transplant Yes Yes possible? unit No Is seed viability high? No Reject vegetative propagule transplant Yes unit

Is density of active growing shoots high? Is seedling No No establishment high? Yes Potential with larger Potential with site transplant size for enhancement plugs OR increase density of sprigs Yes

Consider seedlings for Consider sprigs for Consider plugs for transplant unit transplant unit transplant unit

Minimise handling

Figure 5.4: Step 3, transplant unit and technique selection (Stage II), in the planning process.

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In the final Stage (III) of the framework habitat enhancement is considered, in Chapter 4. ASG mats, through site stabilisation (enhancement), increased transplant survival and rhizome elongation, thus supporting the notions that excessive water impedes seagrass transplant survival and ASG mats enhance a site. Light was not detected as a limiting factor to seagrass growth. No support for the opinion that seagrass transplants should be taken from deeper water and transplanted into shallow water was found. Further research is required to elucidate the role of light on seagrass transplants in Western Australia, particularly for regions deemed future mitigation sites. As discussed earlier in this chapter, the habitat argument of only transplanting into realised niche may limit potential seagrass transplant sites. Theory suggests there is a lower likelihood of success when transplanting into sub-optimal habitats outside of the realised niche in the fundamental niche. But as the transplants in the deep sites at this study demonstrate, transplant survival and productivity was fair and the chance of success when moving beyond the realised niche is possible.

By examining ecological and physiological parameters affecting the seagrass Posidonia australis, knowledge was gained to strengthen the theoretical basis for selecting a suitable transplant site, unit and technique and habitat modification for a seagrass mitigation effort. Chapter 3 determined that vegetative propagules of P. australis are favoured as transplant units along the Perth metropolitan coastline. Chapter 3 determined that the environmental factors seen as affecting P. australis may not be limiting and needs further investigation to help predict more suitable transplant sites. Chapter 4 determined that ASG mats enhance sites, which increases P. australis transplant survival and rhizome elongation.

5.3 Future considerations Over the previous two decades the UN, through the World Bank, has made recommendations for environmental reporting, to be used as an estimator of a nations productivity, thus creating a green estimate of Gross Domestic Product (GDP) (State of the Environment Advisory Council 1996). The basis for such recommendations revolves around management of environmental challenges, which if handled effectively should lead to ecologically sustainable development. These UN recommendations and the recognition of the importance of “natural assets” (air, soils, water, forests etc) are the motivation behind the State of the Environment reporting in Australia.

Seagrass is one such habitat covered by the State of the Environment reporting and is directly reported on in two (habitat extent, habitat quality) of the eight class of issues the make up the 61 key indicators for the national State of the Environment reporting on Estuaries and the Sea. Seagrass is also indirectly related to three more class issues (non-renewable products through sea sand-mining, water/sediment quality and ecosystem level processes via their role as a nursery for juvenile fish and crustacea). These key and class indicators provide the basis of future data requirements in Australia (State of the Environment Advisory Council 1996; Ward et al. 1998). In terms of data requirements for habitat extent, Ward et al. (1998) recommend implementation of seagrass mapping for estuaries and IMCRA subregions, with calculations of

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percentage significance for the species in each subregion. CSIRO and relevant State agencies have undertaken this work. Within the habitat quality indicator, seagrass species/assemblage changes for reference sites (linked with the habitat extent studies) in each subregion will be examined. Estimates of change and the size of change will be calculated, although no ongoing monitoring is designated for this work (Ward et al. 1998). To date, funding for research is strongly influenced by the outlined data requirements in these documents. Thus, future seagrass studies in Australia are linked to biodiversity studies.

At the State level (Western Australia), a number of additional research areas need further attention with respect to seagrass mitigation and hence maintenance of biodiversity. These include: 1) Examination of the potential for light limitation of seagrasses along the Western Australian coastline is needed. Investigations need to concentrate on light limitations (SI requirements) of different seagrass species, thus providing a theoretical framework for transplanting depths for seagrasses. Furthermore,

accurate light compensation (Ic) and saturation (Isat) levels are needed for all species along the range of water bodies in Western Australia. This is particularly important because extrapolation of values from one water body to another is fraught with assumptions, causing a gross under- or overestimation of P-I curve values. For example, extrapolating P-I curve data from cool temperate waters such as Albany, is inadequate for subtropical waters such as Shark Bay. This is especially true because P-I curves are affected by water temperature and the extinction coefficients along the coastline change with proximity to environmental factors such as pollution and run off; 2) Transient and permanent species diversity, abundance and richness within seagrass meadows along the Western Australian coastline is required, to expand biodiversity data. Without this data, we are unable to utilise this resource efficiently or attain equivalency in mitigation attempts; 3) A lack of data on the impacts of introduced species (both macrobenthos and pathogens) on seagrass meadows exists. The link to biodiversity issues is great; and 4) Metapopulation dynamic modelling of seagrass communities in Western Australia, will be a useful tool in determining how local populations of seagrass spread, persist and colonise. This provides data to develop a framework outlining how different seagrass species, in different regions, can be mitigated.

During the period in which this Ph.D. was conducted (1993 – 1999) the following list of seagrass habitats have been destroyed or are being threatened. Few of these examples have been mitigated and it is unlikely that many will be successfully restored.

1. Werribee, Port Phillip Bay, Victoria. Eutrophication, due to a sewerage outfall, decreases available light resulting in seagrass die off. Seagrasses at this site must also compete for space and resources with a number of introduced species, such as Sabella spallanzanii and Undaria pinnatifida (pers. obs.); 2. Geelong Arm, Port Phillip Bay, Victoria. The Geelong Arm has been dredged to allow the entry of international and national shipping. It has been suggested that the dredge spoil be dumped at Werribee on

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algal and seagrass regions to control the spread of U. pinnatifida (T. Burridge 1996 ASPAB conference pers. comm.); 3. Botany Bay, New South Wales. Seagrass meadows were reclaimed and built upon for the extension of airport runways. Seagrass transplants by N.S.W. Fisheries have mitigated some of this loss (P. Gibbs and J. Upstone, pers. comm.); 4. Eden, New South Wales. Seagrass meadows adjacent to the Harris-Daishowa woodchip berth will be built upon during planned extensions of the woodchip berth (pers. obs.); 5. Moreton Bay, Queensland. A revegetation project was destroyed when luxury apartments were built on the mud flats where the transplantation work was being done (C. Conacher 1995, pers. comm.); 6. Hinchinbrook Island, Queensland. Entrepreneur Keith Williams built a marina and luxury resort in an area rich in seagrass, invertebrate and marine mammal life. To develop the region seagrass habitats were reclaimed and dredged (Booth 1996; Moorhouse 1996; A. Morris pers. comm.); 7. Tinderbox Marine Park, Tasmania. An outbreak of the exotic alga Undaria pinnatifida has the potential to impact upon seagrass meadows (pers. obs.); 8. Princess Royal and Oyster Harbour, Albany, Western Australia. Eutrophication is affecting seagrasses (pers. obs.), although the levels of rural and urban runoff have decreased since Bastyan’s (1986) work. Bastyan (1986) found that eutrophication had caused the loss of 8.1 km2 and 7.2 km2 of seagrass meadows in each of the harbours, respectively. Both harbours are also affected by the exotic fanworm Sabella spallanzanii, which forms dense meadows and is capable of out-competing Posidonia plants for space (pers. obs.); 9. Rottnest Island, Western Australia. Scouring from boat moorings has resulted in localised losses of seagrass patches (Lukatelich et al. 1987; Walker et al. 1989; pers. obs.); 10. Rous Head, Fremantle, Western Australia. Land reclamation of seagrass habitats has occurred to create land for industrial use (pers. obs.); 11. Cockburn Sound, Western Australia. Eutrophication from urban, industrial and rural runoff caused the loss of 3,300 hectares of seagrass meadows (Cambridge and McComb 1984; Cambridge et al. 1986; Silberstein et al. 1986; pers. obs.). The introduced fanworm, Sabella spallanzanii, is present in Cockburn Sound and when in large numbers forms dense meadows that can effectively compete for space with seagrass plants; 12. Success and Parmelia Banks, Western Australia. Both Success and Parmelia Banks have good seagrass cover but are mined for high quality sea sand; 13. Shipping channel from Gage Roads into Cockburn Sound, Western Australia: regrowth of seagrasses in the shipping channel was destroyed when the channel was re-dredged (I. LeProvost pers. comm.); 14. Geraldton, Western Australia. An area rich in seagrass and associated fauna will be blasted and dredged in a proposed plan to build a deep water port and associated industrial area, north of an already existing deep water port and industrial area. Consultants have claimed that no seagrasses are in the area, however seagrass and marine ecology experts have since proven this to be false (H. Astill and M. van Keulen, pers. comm.); and

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15. Exmouth Gulf, Western Australia: A proposal for sea sand mining in Exmouth Gulf has been tabled. If this occurs, tropical seagrass meadows and invertebrate life will be damaged (R. McCloughlin pers. comm.).

It is hoped that the restoration framework presented here may prove useful in the development of an efficient and scientifically rigorous tool for establishing successful restoration and mitigation attempts.

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References Ackerman, J. D. (1983). Current flow around Zostera marina plants and flowers: implications for submarine pollination. Biological Bulletin 165: 504. Addy, C. E. (1947a). Eelgrass planting guide. Maryland Conservation 24: 16. Addy, C. R. (1947b). Germination of eelgrass seed. Journal of Wildlife Management 11: 279. Aioi, K. and Pollard, P. C. (1993). Biomass, leaf growth and loss rate of the seagrass Syringodium isoetifolium on Dravuni Island, Fiji. Aquatic Botany 46: 283-292. Anderson, S.M. and Charters, A. C. (1982). A fluid dynamics study of seawater flow through Gelidium nudifrons. Limnology and Oceanography 27(3): 399-412. Ayal, Y. and Safriel, U. N. (1983). Does suitable habitat guarantee successful colonization? Journal of Biogeography 10: 37-46. Backman, T. W. and Barilotti, D. C. (1976). Irradiance reduction effects on standing crops of the eelgrass Zostera marina in a coastal lagoon. Marine Biology 34: 33-40. Baldwin, J. R. and Lovvorn, J. R. (1994). Expansion of seagrass habitat by the exotic Zostera japonica, and its use by dabbling ducks and brant in Boundary Bay, British Columbia. Marine Ecology Progress Series 103: 119-127. Barko, J. W. and Smart, R. M. (1983). Effects of organic matter additions to sediment on the growth of aquatic plants. Journal of Ecology 71: 161-175. Barko, J. W. and Smart, R. M. (1986). Sediment-related mechanisms of growth limitation in submersed macrophytes. Ecology 67(5): 1328-1340. Barko, J. W., Gunnison, D and Carpenter, S. R. (1991). Sediment interactions with submersed macrophyte growth and community dynamics. Aquatic Botany 41: 41-65. Barko, J. W., Smart, R. M., McFarland, D. G. and Chen, R. L. (1988). Interrelationships between the growth of Hydrilla verticillata (L.f.) Royle and sediment nutrient availability. Aquatic Botany 32: 205-216. Baron, J., Clavier, J. and Thomassin, B. A.. (1993). Structure and temporal fluctuations of two intertidal seagrass-bed communities in New Caledonia (SW Pacific Ocean). Marine Biology 117: 139-144. Baskin, C. C. and Baskin, J. M. (1998). Seeds. Ecology, Biogeography, and Evolution of Dormancy and Germination. Academic Press. 666pp. Bastyan, G. R. (1986). Distribution of seagrasses in Princess Royal Harbour and Oyster Harbour on the southern coast of western Australia. Department of Conservation and the Environment. Perth, Western Australia. Technical series 1. Batuik, R. A., Orth, R. J., Moore, K. A., Dennison, W. C., Stevenson, J. C., Staver, L., Carter, V., Rybicki, N., Hickman, R. E., Kollar, S., Bieber, S. and Heasly, P. (1992). Chesapeake Bay submerged aquatic vegetation habitat requirements and restoration targets: a technical synthesis. U. S. Environmental Protection Agency, Chesapeake Bay Program. Report CBP/TRS 83/92. Bell, S. S. and Hicks, G. R. F. (1991). Marine landscapes and faunal recruitment: a field test with seagrasses and copepods. Marine Ecology Progress Series 73: 61-68.

136

Birch, W. R. and Birch, M. (1984). Succession and patterns of tropical intertidal seagrasses in Cockle Bay, Queensland, Australia: A decade of observations. Aquatic Botany 19: 343-367. Bird, K. T., Jewett-Smith, J. and Fonseca, M. S. (1994). Use of in vitro propagated Ruppia maritima for seagrass meadow restoration. Journal of Coastal Research 10(3): 732-737. Björkman, O., Demmig, B. and Andrews, T. J. (1988). Mangrove photosynthesis: response to high- irradiance stress. pp. 43-61. In: Evans, J. R., von Caemmerer, S. and Adams III, W. W. (eds.). Ecology of Photosynthesis in Sun and Shade. CSIRO; Melbourne. Booth, C. (1996). Hinchinbrook dugongs: extinction for profit. Green Left Weekly. October 16th. pp. 15 & 32. Borowitzka, M. A. (1977). Algal calcification. Oceanography and Marine Biology Annual Review 15: 189-223. Boudoureque, C. F., de Grissac, J. and Meinesz, A. (1984). Relations entre le sedimentation et l’allongement des rhizomes orthotropes de Posidonia oceanica dans la Baie d’Elbu (Corse). pp. 185-191. In: Boudoureque, C. F., de Grissac, J., Olivier, O. (eds.) International Workshop on Posidonia oceanica Beds. G. I. S. Posidonie publ. France Marseilles 1. van Breedveld, J. F. (1975). Transplanting of seagrass with emphasis on the importance of substrate. Florida Marine Research Publication 17: 26. Brouns, J. J. W. M. (1985). A comparison of the annual production and biomass in three monospecific stands of the seagrass Thalassia hemprichii (Ehrenb.) Aschers. Aquatic Botany 23: 149-175. Brouns, J. J. W. M. (1987a). Growth patterns in some Indo-West-Pacific seagrasses. Aquatic Botany 28: 39-61. Brouns, J. J. W. M. (1987b). Quantitative aspects and dynamics aspects of a mixed seagrass meadow in Papua New Guinea. Aquatic Botany 29: 33-47. Brouns, J. J. W. M. (1987c). Aspects of production and biomass of four seagrass species (Cymodoceoideae) from Papua New Guinea. Aquatic Botany 27: 333-362. Brown, J.S. and Venable, D.L. (1986). Evolutionary ecology of seed-bank annuals in temporally varying environments. American Naturalist 127(1): 31-47. Bulthuis, D. A. (1983). Effects of in situ light reduction on density and growth of the seagrass Heterozostera tasmanica (Martens ex Aschers.) den Hartog in Western Port, Victoria, Australia. Journal of Experimental Marine Biology and Ecology 67: 91-103. Bulthuis, D. A. and Woelkerling, W. J. (1981). Biomass accumulation and shading effects of epiphytes on leaves of the seagrass, Heterozostera tasmanica, in Victoria, Australia. Aquatic Botany 16: 116- 136. Bürger, R. and Lynch, M. (1997). Adaptation and extinction in changing environments. pp. 209-239. In: Bijlsma, R. and Loeschcke, V. (eds.). Environmental Stress, Adaptation and Evolution. Birkhäuser Verlag; Basel. Burkholder, P. R. and Doheny, T. E. (1968). The Biology of Eelgrass. Department of Conservation and Waterways: Town of Hempstead, Long Island, New York. Butler, A. J. and Jernakoff, P. (eds.) (1999). Australian Seagrasses. CSIRO; Marmion, Australia.

137

Calumpong, H. P., Menez, E. G. and Phillips, R. C. (1996). Factors affecting survival and growth of seagrass transplants on Negros Island, Central Philippines (Abstract). pp. 378. In: Kuo, J., Phillips, R. C., Walker, D. I. and Kirkman, H. (eds.). Seagrass Biology. Proceedings of an International Workshop. Rottnest Island, Western Australia. 25-26 January 1996. Faculty of Science, University of Western Australia; Perth. Cambridge, M. L. (1975). Cited in Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Cambridge, M. L. (1977). Cited in Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Cambridge, M. L. (1978). Cited in Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Cambridge, M. L. (1979). Cockburn Sound environmental study: Technical report on seagrasses. Department of Conservation and Environment. Report No. 7. Cambridge, M. L. (1980). Ecological Studies on Seagrass of South Western Australia with Particular Reference to Cockburn Sound. Ph.D. Dissertation. Botany department. University of Western Australia. Perth, Western Australia, Australia. Cambridge, M. L. and McComb, A. J. (1984). The loss of seagrass from Cockburn Sound, Western Australia. I. The time course and magnitude of seagrass decline in relation to industrial development. Aquatic Botany 20: 229-266. Cambridge, M. L., Chiffings, A. W., Brittan, C., Moore, L. and McComb, A. J. (1986). The loss of seagrasses in Cockburn Sound, Western Australia. II. Possible causes of decline. Aquatic Botany 24: 269-285. Carangelo, P. D., Oppenheimer, C. H. and Picarazzi, P. E. (1979) In: Proceedings of the 6th Annual Conference on restoration of coastal vegetation in Florida. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Carigan, R. and Kalff, J. (1980). Phosphorous sources for aquatic weeds: water or sediment? Science 207: 987-989. Carlton, J. T. (1979). History, Biogeography, and Ecology of the Introduced Marine and Estuarine Invertebrates of the Pacific Coast of North America. Ph.D. Dissertation. University of California, Davis, USA. 904pp. Carlton, J. T. (1989). Man’s role in changing the face of the ocean: biological invasions and implications for conservation of nearshore environments. Conservation Biology 3: 265-273. Carpenter, S. R. (1996). Microcosm experiments have limited relevance for community and ecosystem ecology. Ecology 77(3): 677-680. Caye, G. (1980). Analyse du polymorphisme caulinaire chez Posidonia oceanica (L.) Delile. Bulletin Societe Botanique de France 127: 257-262.

138

Caye, G. and Meinesz, A. (1984). Oberservations sur la floraison et la fructification de Posidonia oceanica. In: Boudouresque, C. F., Meinesz, A., Fresi, E. and Gravez, V. (eds.) International Workshop on Posidonia oceanica Beds. GIS Posidonie publ., 1: 193-201. Caye, G. and Meinesz, A. (1985a). Observations on the vegetative development, flowering and seedling of Cymodocea nodosa (Ucria) Ascherson on the Mediterranean coasts of France. Aquatic Botany 22: 277-289. Caye, G. and Meinesz, A. (1985b). Experimental study of seed germination in the seagrass Cymodocea nodosa. Aquatic Botany 26: 79-87. Chiffings, A. W. and McComb, A. J. (1983). Boundaries in phytoplankton populations. Proceedings of the Ecology Society of Australia 11: 27-38. Chin,. (1988). Recalcitrant seeds; a status report. International Board for Plant Genetic Resources, Rome. 28pp. Churchill, A. C. (1983). Field studies on seed germination and seedling development in Zostera marina L. Aquatic Botany 16: 21-29. Churchill, A. C. (1992). Growth characteristics of Zostera marina seedlings under anaerobic conditions. Aquatic Botany 43: 379-392. Churchill, A. C., Cok, A. E. and Riner, M. I. (1978). Stabilisation of Subtidal Sediments by the Transplantation of the Seagrass Zostera marina L., NYSSGP-RS-78-15 (New York Sea Grant, Garden City, U.S.A). Cipollini, M. L., Wallascesenft, D. A. and Whigham, D. F. (1994). A model of patch dynamics, seed dispersal, and sex ratio in the dioecious shrub Lindera benzoin (Lauradeae). Journal of Ecology 82: 621-633. Clarke, S. M. and Kirkman, H. (1989). Seagrass dynamics. pp. 304-344. In: Larkum, A. W. D., McComb, A. J. and Shepherd, S. A. (eds.) Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Coles, R. G., Poiner, I. R. and Kirkman, H. (1989). Regional studies – seagrasses of north-eastern Australia. pp. 261-278. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Conacher, C. A., Poiner, I. R. And O’Donohue, M. (1994a). Morphology, flowering and seed production of Zostera capricorni Aschers. in subtropical Australia. Aquatic Botany 49: 33-46. Conacher, C. A., Poiner, I.R., Butler, J., Pun, S. And Tree, D. J. (1994b). Germination, storage and viability testing of seeds of Zostera capricorni Aschers. From a tropical bay in Australia. Aquatic Botany 49: 47-58. Connell, J.H. (1961). The influence of interspecific competition and other factors on the distribution of the barancle Chthamalus stellatus. Ecology 42: 710-723. Cooper, G. (1976). Jardinier de la Mer 1, 1. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117.

139

Cooper, G. (1977). Jardinier de la Mer 2, 1. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Cooper, G. (1978). Peuples Méditerranée 4, 43. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Cooper, G. (1979). Jardinier de la Mer 3, 1. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Cooper, L. W. and McRoy, C. P. (1988). Anatomical adaptations to rocky substrates and surf exposure by the seagrass genus Phyllospadix. Aquatic Botany 32: 365-381. Copper, G. (1981). Jardinier de la Mer. L’association-foundation. G, Cooper, Hyeres, France. Cahier No. 5. Critchley, C. (1988). The molecular mechanism of photoinhibition – Facts and fiction. pp. 27–41. In: Evans, J. R., von Caemmerer, S. and Adams III, W. W. (eds.). Ecology of Photosynthesis in Sun and Shade. CSIRO; Melbourne. Crosby, A. W. (1986). Ecological Imperialism. The Biological Expansion of Europe, 900-1900. Cambridge University Press; Cambridge. 368pp. Daehler, C. C. and Strong, D. R. (1996). Can you bottle Nature? The roles of microcosms in ecological research. Ecology 77(3): 663-664. Davis, F. W., Stine, P. A., Stms, D. M., Borchert, M. I. and Hollander, A. (1995). Gap analysis of the actual vegetation of California, 1. The southwestern region. Madroño 41: 40-78. Dawes, C. J. and Tomasko, D. A. (1988). Depth distribution in two Thalassia testudinum meadows on the west coast of Florida: a difference in effect of light availability. Publ. Staz. Zool. Napoli (I: Marine Ecology) 9: 123-130. den Hartog, C. (1970). The Sea-grasses of the World. North Holland, Amsterdam. den Hartog, C. (1987). “Wasting disease” and other dynamic phenomena in Zostera beds. Aquatic Botany 27: 3-14. den Hartog, C. (1994). Suffocation of a littoral Zostera bed by Enteromorpha radiata. Aquatic Botany 47: 21-28. den Hartog, C. and Polderman, P. J. G. (1975). Changes in seagrass populations of the Dutch Wadden Sea. Aquatic Botany 1: 141-147. Dennison, W. C. (1987). Effects of light on seagrass photosynthesis, growth and depth distribution. Aquatic Botany 27: 15-26. Dennison, W. C. (1990a). Shoot density. pp 61-64. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Dennison, W. C. (1990b). Leaf Production. pp 77-80. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Dennison, W. C. (1990c). Rhizome/root production. pp 81-82. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Dennison, W. C. and Alberte, R. S. (1985). Role of daily light period in the depth distribution of Zostera marina (eelgrass). Marine Ecology Progress Series 25: 51-61.

140

Dennison, W. C. and Kirkman, H. (1996). Seagrass survival model. pp. 341-344. In: Kuo, J., Phillips, R. C., Walker, D. I. and Kirkman, H. (eds.). Seagrass Biology. Proceedings of an International Workshop. Rottnest Island, Western Australia. 25-26 January 1996. Faculty of Science, University of Western Australia; Perth. Dennison, W. C., Orth, R. J., Moore, K. A., Stevenson, J. C., Carter, V., Kollar, P. W., Bergstrom, P. W. And Batuik, R. A. (1993). Assessing water quality with submersed aquatic vegetation. Habitat requirements as barometers of Chesapeake Bay health. BioScience 43(2): 86-94. Denny, P. (1972). Sites of nutrient absorption in aquatic macrophytes. Journal of Ecology 60: 819-829. Derrenbecker, J, Jr. and Lewis, R. R. (1983). Seagrass habitat restoration in Lake Surprise, Florida Keys. pp. 132-154. In: Stoval, R. H. (ed.) Proceedings 9th Annual Conference on Wetlands Restoration and Creation. Hillsborough community college. Tampa, Florida. Disraeli, D. J. (1984). The effect of sand deposits on the growth and morphology of Ammophila breviligulata. Journal of Ecology 72: 145-154. Dobson, A. P., Bradshaw, A. D. and Baker, A. J. M. (1997). Hopes for the future: restoration ecology and conservation biology. Science 277: 515-521. Drake, J. A., Huxel, G. R. and Hewitt, C. L. (1996). Microcosms as models for generating and testing community theory. Ecology 77(3): 670-677. Drew, E. A. (1978). Factors affecting photosynthesis and its seasonal variation in the seagrasses Cymodocea nodosa (Ucria) Aschers, and Posidonia oceanica (L.) Delile in the Mediterranean. Journal of Experimental Marine Biology and Ecology 31: 173-194. Drew, E. A. (1979). Physiological aspects of primary production in seagrasses. Aquatic Botany 7: 139- 150. Drew, E. A. and Jupp, B. P. (1976). Some aspects of the growth of Posidonia oceanica in Malta. pp. 357–367. In: Drew, E. A., Lythgoe, J. N. and Woods, J. O (eds.). Underwater Research. Academic Press; London. Duarte, C. M. (1991). Seagrass depth limits. Aquatic Botany 40: 363-377. Duarte, C. M. and Sand-Jensen, K. (1990a). Seagrass colonization: patch formation and patch growth in Cymodocea nodosa. Marine Ecology Progress Series 65: 193-200. Duarte, C. M. and Sand-Jensen, K. (1990b). Seagrass colonisation: biomass development and shoot demography in Cymodocea nodosa patches. Marine Ecology Progress Series 67: 97-103. Duarte, C. M., Marbá, N., Agawin, N. S. R., Cebrián, J., Enriquez, S., Fortes, M. D., Gallegos, M. E., Merino, M., Olesen, B., Sand-Jensen, K., Uri, J. S. and Vermaat, J. E. (1994). Reconstruction of seagrass dynamics: age determinations and associated tools for the seagrass ecologist. Marine Ecology Progress Series 107: 195-209. Duarte, C. M., Terrados, J., Agawin, N. S. R., Fortes, M. D., Bach, S. and Kenworthy, W. J. (1997). Response of a mixed Philippine seagrass meadow to experimental burial. Marine Ecology Progress Series 147 (1-3): 285-294. Ducker, S. C., Foord, N. J. And Know, R. B. (1977). Biology of Australian seagrasses: the genus Amphibolis C. Agardh (Cymodoceaceae). Australian Journal of Botany 25: 67-95.

141

Dunton, K. H. (1994). Seasonal growth and biomass of the subtropical seagrass Halodule wrightii in relation to continuous measurements of underwater irradiance. Marine Biology 120: 479-489. Durako, M. J. and Moffler. M. D. (1984). Qualitative assessment of five artificial growth media on growth and survival of Thalassia testudinum (Hydrocharitaceae) seedlings. pp. 73-92. In: Webb, F. J. (ed.). Proceedings of the 11th Annual Conference Wetlands Restoration and Creation. Hillsborough Community College. Tampa, Florida. Durako, M. J., Hall, M. O., Sargent, F. and Peck, S. (1992). Propeller scars in seagrass beds: an assessment and experimental study of recolonisation in Weedon Island State Preserve, Florida. pp. 42-53. In: Webb Jr, F. J. (ed.) Proceedings of the Nineteenth Annual Conference on Wetlands Restoration and Creation. Hillsborough community college. Florida. Eckman, J. E. (1983). Hydrodynamic processes affecting benthic recruitment. Limnology and Oceanography 28(2): 241-257. Eldered, R. A. and Maun, M. A. (1982). A multivariate approach to the problem of decline in vigour of Ammophila. Canadian Journal of Botany 39: 1371-1380. Eleuterius, L. N. (1975). Estuarine Research. Volume 2. Academic Press, New York. Ellner, S. (1985a). ESS germination strategies in randomly varying environments. I. Logististic-type models. Theoretical Population Biology 28: 50-79. Ellner, S. (1985b). ESS germination strategies in randomly varying environments. II. Reciprocal yiel-law models. Theoretical Population Biology 28: 80-116. Elton, C. S. (1958). The Ecology of Invasions by Plants and Animals. Methuen and Co.; London. 181pp. Erftemeijer, P. L. A., Osinga, R. and Mars, A. E. (1993). Primary production of seagrass beds in South Sulawesi (Indonesia): a comparison of habitats, methods and species. Aquatic Botany 46: 67-90. Evans, R. P. and Rizler, F. H. (1980). Measurement of whole lake sediment accumulation and phosphorous retention using lead-210 dating. Canadian Journal of Fisheries and Aquatic Research 37: 817-822. Ewanchuk, P. J. (1995). Population Growth of Eelgrass (Zostera marina L.): the Relative Importance of Sexual versus Asexual Reproduction. Masters Thesis. San Diego State University, San Diego, USA. Fagerstrom, J. A. (1986). The Evolution of Reef Communities. John Wiley and Sons; New York. Fonseca, M. S. (1989). Regional analysis of the creation and restoration of seagrass systems. pp. 175- 198. In: Kusler, J. A. and Kentula, M. E. (eds.). Wetland Creation and Restoration: the Status of the Science. Volume I. Regional reviews. EPA/600/3-89/0831. Environ. Res. Lab., Corvallis, OR. Fonseca, M. S. (1992). Restoring seagrass systems in the United States. pp. 79-110. In: Thayer, G. W. (ed.) Restoring the Nation’s Marine Environment. Maryland Sea Grant College Pub. UM-SG-TS- 92-06. 716pp. Fonseca, M. S. (1994). A Guide to Planting Seagrasses in the Gulf of Mexico. Texas A & M University Sea Grant College Program. 26pp. Fonseca, M. S. (1997). Review of seagrass rehabilitation project S2. pp. 73-98. In: Paling, E. I., Walker, C. J. and van Keulen, M. Project S2, phase 2: Seagrass Rehabilitation in Owen Anchorage, Western Australia. MAFRL Report No. 97/8. Murdoch University, Perth.

142

Fonseca, M. S. and Bell, S. S. (1998). Influence of physical setting on seagrass landscapes near Beaufort, North Carolina, USA. Marine Ecology Progress Series 171:109-121. Fonseca, M. S. and Fischer, J.S. (1986). A comparison of canopy friction and sediment movement between four species of seagrass with reference to their ecology and restoration. Marine Ecology Progress Series 29: 15-22. Fonseca, M. S. and Kenworthy, W. J. (1987). Effects of current on photosynthesis and distribution of seagrass. Aquatic Botany 27: 59-78. Fonseca, M. S., Kenworthy, W. J. and Thayer, G. W. (1987). Transplanting of the seagrass Halodule wrightii, Syringodium filiforme and Thalassia testudinum for sediment stabilisation and habitat development in the southeast region of the United States. Department US Army Corps of Engineers. Technical Report EL-87-8, North Carolina. Fonseca, M. S., Kenworthy, W. J. and Thayer, G. W. (1996). Draft Guidelines for Mitigation and Restoration of Seagrass in the United States and Adjacent Waters. National Marine Fisheries Service, NOAA Coastal Ocean Program Decision Analysis series. Fonseca, M. S., Kenworthy, W. J., Cheap, K. M., Currin, C. A. and Thayer, G. W. (1985). A low-cost transplanting technique for shoalgrass (Halodule wrightii) and manatee grass (Syringodium filiforme). Department of Army, US Army Corps Engineers, Report No. E1-84-1, Washington D. C. Fonseca, M. S., Kenworthy, W. J., Homziak, J. and Thayer, G. W. (1979). Transplanting of eelgrass and shoalgrass as a potential means of economically mitigating a recent loss of habitat. pp. 279-326. In: Cole, D. P. (ed.) Proceedings of the Sixth Annual Conference on the Restoration and Creation of Wetlands. Hillsborough community college, Tampa, Florida. Fonseca, M. S., Meyer, D. L., Murphey, P. L., Kenworthy, W. J., Thayer, G. W. And LaCroix, M. W. (1990). Comparison of fauna among natural and transplanted eelgrass Zostera marina meadow: criteria for mitigation. Marine Ecology Progress Series 65: 251-264. Fonseca, M. S., Zieman, J. C., Thayer, G. W. And Fischer, J. S. (1983). The role of current velocity in structuring eelgrass (Zostera marina L.) meadows. Estuarine, Coastal and Shelf Science 17: 367- 380. Fonseca, M.S., Kenworthy, W. J., Thayer, G. W. (1982). A Low-cost Planting Technique for Eelgrass (Zostera marina L.). U.S. Army Corps of Engineers. Coastal Engineering Technical Aid No. 82-6, Coastal Engineering Research Centre, Ft. Belvoir, Va. 15p. Fry, B. and Virnstein, R. W. (1988). Leaf production and export of the seagrass Syringodium filiforme Kütz. in the Indian River Lagoon, Florida. Aquatic Botany 30: 261-266. Fuss, C. M. Jr. And Kelly, J. A. Jr. (1969). Survival and growth of sea grasses transplanted under artificial conditions. Bulletin of Marine Science 19: 351-365. Gallegos, M. E., Merino, M., Marbá, N. and Duarte, C. M. (1993). Biomass and dynamics of Thalassia testudinum in the Mexican Caribbean: elucidating rhizome growth. Marine Ecology Progress Series 95: 185-192.

143

Gallegos, M. E., Merino, M., Rodriguez, A., Marbá, N. and Duarte, C. M. (1994). Growth patterns and demography of pioneer Caribbean seagrasses Halodule wrightii and Syringodium filiforme. Marine Ecology Progress Series 109: 99-104. Gambi, M. C., Nowell, A. R. M. and Jumars, P. A. (1990). Flume observations on flow dynamics in Zostera marina (Eelgrass) beds. Marine Ecology Progress Series 61: 159-169. Genot, I., Caye, C., Meinesz, A. and Orlandini, M. (1994). Role of chlorophyll and carbohydrate contents in survival of Posidonia oceanica cuttings transplanted to different depths. Marine Biology 119: 23- 29. Gerard, V. A. and Mann, K. H. (1979). Growth and production of Laminaria longicruris (Phaeophyta) populations expsoed to different intensities of water movement. Journal of Phycology 15: 33-41. Giesen, W. B. J. T., van Katwijk, M. M. and den Hartog, C. (1990). Eelgrass condition and turbidity in the Dutch Wadden Sea. Aquatic Botany 37: 71-85. Ginsberg, R. N. and Lowenstan, H. A. (1958). The influence of marine bottom communities on the depositional environment of sediment. Journal of Geology 66: 310-318. Giraud, G. (1977). Recensement des floraisons de Posidonia oceanica (L.) Delile en Méditerranée. Rapport. Commission Internationale pour la Mer Méditerranee [France] 24: 127-130. Glantz, S. A. (1992). Primer of Biostatistics. McGraw-Hill; Singapore. pp. 218-268, 320-365. Gordon, D. (1995). Status of seagrass restoration: review of International literature. Cockburn Cement Limited Shellsand Dredging Environmental Management Programme. Project S2, phase 1. pp. 3- 12. Gordon, D. M., Grey, K. A., Chase, S. C. and Simpson, C. J. (1994). Changes to the structure and productivity of a Posidonia sinuosa meadow during and after imposed shading. Aquatic Botany 47: 265-275. van Groenendael, J. M., Klimeš, L., Klimešová, J. and Hendriks, R. J. J. (1997). Comparative ecology of clonal plants. pp. 191-209. In: Silvertown, J., Franco, M. and Harper, J. L. (eds.). Plant Life Histories. Ecology, Phylogeny and Evolution. The Royal Society, Cambridge University; Cambridge. Guerrant, E. O. and Pavlik, B. M. (1998). Reintroduction of rare plants: Genetics, demography and the role ex situ conservation methods. pp. 80-108. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology: For the Coming Decade. 2nd edition. Chapman and Hall; New York. Hancock, C. N. (1992). Seagrass Transplantation Trials and Factors Affecting their Success. B.Sc. Honours Thesis. Environmental Science, Murdoch University. Perth, Western Australia, Australia. Hanski, I. and Simberloff, D. (1997). The metapopulation approach, its history, conceptual domain, and application to conservation. pp. 5-26. In: Hanski, I. and Gilpin, M. E. (eds.) Metapopulation Biology. Ecology, Genetics, and Evolution. Academic Press; San Diego. Harrington, J. F. (1973). Problems of seed storage. pp. 251-262. In: Heydecker, H. (ed.). Seed Ecology. Pennsylvania State University Press; University Park. Harris, M. McD., King, R. J. and Ellis, J. (1980). The eelgrass Zostera capricorni in Illawarra Lake, N.S.W. Proceedings of the Linnean Society of New South Wales 104: 23-33.

144

Harrison, P. G. (1979). Reproductive strategies in intertidal populations of two co-occurring seagrasses (Zostera spp.). Canadian Journal of Botany 57: 263-268. Harrison, P. G. (1987). Natural expansion and experimental manipulation of seagrass (Zostera spp.). Abundance and the response of infaunal invertebrates. Estuarine, Coastal and Shelf Science 24: 799-812. Harrison, P. G. (1990). Experimental eelgrass transplants in southwestern British Columbia, Canada. Proceedings of the California Eelgrass Symposium. Chula Vista, California May 1988. pp. 46-57. Harrison, P. G. (1993). Variations in demography of Zostera marina and Z. noltii on in intertidal gradient. Aquatic Botany 45: 63-77. Harrison, P. G. and Bigley, R. E. (1982). The recent introduction of the seagrass Zostera japonica Aschers. and Graebn. to the Pacific coast of North America. Canadian Journal of Fisheries and Aquatic Science 39: 1642-1648. Harrison, S. and Taylor, A. D. (1997). Empirical evidence for metapopulation dynamics. pp. 27-42. In: Hanski, I. and Gilpin, M. E. (eds.) Metapopulation Biology. Ecology, Genetics, and Evolution. Academic Press; San Diego. Hemminga, M. A., Harrison, P. G. and van Lent, F. (1991). The balance of nutrient losses and gains in seagrass meadows. Marine Ecology Progress Series 71: 85-96. Hemminga, M. A., Koutstaal, B. P., van Soelen, J. and Merks, A. G. A. (1994). The nitrogen supply to intertidal eelgrass (Zostera marina). Marine Biology 118: 223-227. Hillman, K., Lukatelich, R. J., Bastyan, G and McComb, A. J. (1991). Water quality and seagrass biomass, productivity and epiphyte load in Princess Royal Harbour, Oyster Harbour and King George Sound. Environmental Protection Authority. Technical Series 39; Perth, W.A. Hillman, K., Lukatelich, R. J., Bastyn, G. and McComb, A.J. (1990). Distribution and biomass of seagrasses and algae, and nutrient pools in water, sediments and plants in Princess Royal Harbour and Oyster Harbour. Environmental Protection Authority. Technical series 40; Perth, Western Australia. Hillman, K., Walker, D. I., Larkum, A. W. D. and McComb, A. J. (1989). Productivity and nutrient limitation. pp. 635-685. In: Larkum, A. W. D., McComb, A. J. and Shepherd, S. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Hocking, P.J., Cambridge, M. L. and McComb, A.J. (1980). Nutrient accumulation in the fruits of two species of seagrass, Posidonia australis and Posidonia sinuosa. Annals of Botany 45: 149-161. Hocking, P.J., Cambridge, M. L. and McComb, A.J. (1981). The nitrogen and phosphorous nutrition of developing plants of two seagrasses, Posidonia australis and Posidonia sinuosa. Aquatic Botany 11: 245-261. Hoffman, R. S. (1990). Recovery of eelgrass beds in Mission Bay, San Diego, California, following beach restoration work. Proceedings of the Californian Eelgrass Symposium. Chula Vista, California, May 1988. 21-27pp.

145

Hootsman, M.J.M., Vermaat, J.E. and Van Vierssen, W. (1987). Seed-bank development, germination and early seedling survival of two seagrass species from the Netherlands: Zostera marina L. and Zostera noltii Hornem. Aquatic Botany 28: 275-285. Hopper, S. D. (1998). An Australian perspective on plant conservation biology in practice. pp. 255-278. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology for the Coming Decade. 2nd edition. Chapman and Hall; New York. Hubbard, P. (1994). Seedling Recruitment of the Mangrove Avicennia marina in the Port Hedland Area. B.Sc. Honours Thesis. Murdoch University, Western Australia, Australia. Hughes, R. N. (1991). Reefs. pp. 213-229. In: Barnes, R. S. K. and Mann, K. H. (eds.) Fundamentals of Aquatic Ecology. Blackwell Scientific Pulbications; London. Hurlbert, S. H. (1984). Pseudoreplication and the design of ecological field experiments. Ecological Monographs 54(2): 187-211. Hutchinson, G. E. (1958). Concluding remarks. Cold Spring Harbour Symposia on Quantitative Biology. 22: 415-427. Iizumi, H., Hattori, A. and McRoy, C. P. (1980). Nitrate and nitrite in interstitial waters of eelgrass beds in relation to the rhizosphere. Journal of Experimental Marine Biology and Ecology 47: 191-201. Iizumi, H., Hattori, A. and McRoy, C. P. (1982). Ammonium regeneration and assimilation in eelgrass (Zostera marina) beds. Marine Biology 66: 59-65. Jacobs, R. P. W. M. (1982). Reproductive strategies of two seagrass species (Zostera marina and Z. noltii) along West European coasts. pp. 150-155. In: Symoens, J. J., Hooper, S. S. and Compère, P. (eds.) Studies on Aquatic Vascular Plants. Royal Botanical Society of Belgium; Brussels. Jewett-Smith, J. and McMillan, C. (1990). Germination and seedling development of Halophila engelmannii Aschers. (Hydrocharitaceae) under axenic conditions. Aquatic Botany 36: 167-177. Johnson, J. E. (1988). Natural seagrass revegetation in Gulf St Vincent, South Australia. Posidonia Newsletter. Jordon, W. R., Gilpin, M. E., Aber, J. D. (1987). Restoration ecology: ecological restoration as a technique for basic research. pp. 3-21. In: Jordan, W. R., Gilpin, M. E. and Aber, J. D. (eds.). Restoration Ecology. A Synthetic Approach to Ecological Research. Cambridge University Press; Cambridge. Josselyn, M., Fonseca, N., Niesen, T. and Larson, R. (1986). Biomass, production and decomposition of a deep water seagrass, Halophila decipiens Ostenf. Aquatic Botany 25: 47-61. Keddy, C. J. and Patriquin, D. G. (1978). An annual form of eelgrass in Nova Scotia. Aquatic Botany 5: 163-170. Kelleher, G., Bleakley, C. and Wells, S. (1995). A Global Representative System of Marine Protected Areas. Volumes 1 - 4. The Great Barrier Reef Marine Park Authority; The World Bank and; The World Conservation Union (IUCN). Washington, U. S. A. Kelly, J. A. Jr, Fuss, C. M. Jr and Hall, J. R. (1971). The transplanting and survival of turtlegrass Thalassia testudinum, in Boca Ciega Bay, Florida. Fishery Bulletin 69: 273-280.

146

Kikuchi, T. (1980). Faunal relationships in temperate seagrass beds. In: Phillips, R. C. And McRoy, C. P. (eds.). A Handbook of Seagrass Biology: an Ecosystem Perspective. Garland; New York. King, R. J. and Hodgson, B. R. (1986). Aquatic angiosperms in coastal saline lagoons of New South Wales. IV. Long-term changes. Proceedings of the Linnean Society of New South Wales 109: 51-60. Kirkman, H. (1989). Restoration and creation of seagrass meadows with special emphasis on Western Australia. Environmental Protection Authority. Technical series No. 30, Perth, Western Australia. Kirkman, H. (1990). Cited in Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Kirkman, H. (1992). Large-scale restoration of seagrass meadows. pp. 110-140. In: Thayer, G. W. (ed.). Restoring the Nation’s Marine Environment. Maryland Sea Grant College. Maryland, USA. Kirkman, H. (1995). Pilot Experiments on Potential Restoration of Seagrass Meadows. CSIRO report. CSIRO Division of Fisheries, Marmion, Western Australia. Kirkman, H. and Kuo, J. (1990). Pattern and process in southern Western Australian seagrasses. Aquatic Botany 37: 367-382. Kirkman, H. and Reid, D. D. (1979). A study of the role of the seagrass Posidonia australis in the carbon budget of an estuary. Aquatic Botany 7: 173-183. Kirkman, H. and Walker, D. I. (1989). Regional studies – Western Australian seagrasses. pp. 157-181. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Komar, P. D. (1989). Physical processes of waves and currents and the formation of marine placers. Rev. Aquatic Science 1: 393-423. Krebs, C. J. (1985). Ecology. The Experimental Analysis of Distribution and Abundance. 3rd Edition. Harper and Row; New York. 486-510pp. Kuo, J. and Kirkman, H. (1987). Floral and seedling morphology and anatomy of Thalassodendron pachyrhizum Den Hartog (Cymodoceaceae). Aquatic Botany 29: 1-17. Kuo, J. and McComb, A. J. (1989). Seagrass taxonomy, structure and development. pp. 6-67. In: Larkum, A. W. D., McComb, A. J. and Shepherd, S. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Kuo, J. and Kirkman, H. (1990). Anatomy of viviparous seagrasses seedlings of Amphibolis and Thalassodendron and their nutrient supply. Botanica Marina 33: 117-126. Kuo, J., Iizumi, H., Nilsen, B.E. and Aioi, K. (1990). Fruit anatomy, seed germination and seedling development in the Japanese seagrass Phyllospadix (Zosteraceae). Aquatic Botany 37: 229-245. Küppers, M., Koch, G. and Mooney, H. A. (1988). Compensating effects of growth in dry matter allocation in response to variation in photosynthetic characteristics induced by photoperiod, light and nitrogen. pp. 287-298. In: Evans, J. R., von Caemmerer, S. and Adams, W. W. (eds.). Ecology of Photosynthesis in Sun and Shade. CSIRO; Melbourne.

147

Lakon, G. (1948) The topographical tetrazolium method for determining the germination capacity of seeds. Plant Physiology 24: 389-393. Larkum, A. W. D. (1976). Ecology of Botany Bay. I. Growth of Posidonia australis (Brown) Hook. f. in Botany Bay and other bays of the Sydney basin. Australian Journal of Marine and Freshwater Research 27: 117-127. Larkum, A. W. D. and den Hartog, C. (1989). Evolution and biogeography of seagrasses. pp. 112-156. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Larkum, A. W. D. and West, R. J. (1990). Long-term changes of seagrass meadows in Botany Bay, Australia. Aquatic Botany 37: 55-70. Lawton, J. H. (1996). The ecotron facility at Silwood Park: the value of “Big Bottle” experiments. Ecology 77(3): 665-669. van Lent, F. and Vershuure, J. M. (1994a). Intraspecific variability of Zostera marina L. (eelgrass) in the estuaries and lagoons of the southwestern Netherlands. I. Population dynamics. Aquatic Botany 48: 31-58 van Lent, F. and Vershuure, J. M. (1994a). Intraspecific variability of Zostera marina L. (eelgrass) in the estuaries and lagoons of the southwestern Netherlands. II. Relation with environmental factors. Aquatic Botany 48: 59-75. van Lent, F. and Vershuure, J. M. (1995). Comparative study on populations of Zostera marina L. (eelgrass): experimental germination and growth. Journal of Experimental Marine Biology and Ecology 185: 77-91. LeProvost Environmental Consultants (1990). Seagrass seedling planting trials in Owen Anchorage, Success and Parmelia Banks, February 1990. LeProvost Environmental Consultants Report No. R331; Perth. Levins, R. (1970). Extinction. pp. 75-107. In: Gerstenhaber, M. (ed.). Some Mathematical Problems in Biology. American Mathematical Society; Providence, RI. Lewis, R. R. (1987). The restoration and creation of seagrass meadows in the southeast United States. pp. 53-174. In: Durako, M. J., Phillips, R. C. and Lewis, R. R. (eds.). Proceedings of the Symposium on Subtropical-tropical Seagrasses of the Southeastern United States. Fla. Dept. Nat. Resources. Littler, M.M., Martz, D. R. & Littler, D.S. (1983). Effects of recurrent sand deposition on rocky intertidal organisms: importance of substrate heterogeneity in a fluctuating environment. Marine Ecology Progress Series 11: 129-139. Loques, F., Caye, G. and Meinesz, A. (1990). Germination in the marine phanerogram Zostera noltii Hornemann at Golfe Juan, French Mediterranean. Aquatic Botany 38: 249-260. Lord, D., Paling, E. and Gordon, D. (1999). Review of seagrass rehabilitation and restoration programmes in Australia. pp. 67-118. In: Butler, A. J. and Jernakoff, P. (eds.). Australian Seagrasses. CSIRO; Hobart.

148

Lovett-Doust, L. (1981). Population dynamics and local specialization in a clonal perennial (Rununculus repens): I. The dynamics of ramets in contrasting habitats. Journal of Ecology 69: 743-755. Lubchenco, J., Olson, A. M., Brubaker, L. B., Carpenter, S. R., Holland, M. M., Hubbell, S. P., Levin, S. A., MacMahon, J. A., Matson, P. A., Melillo, J. M., Mooney, H. A., Peterson, C. H., Pulliam, H. R., Real, L. A., Regal, P. J., Risser, P. G. (1991). The sustainable biosphere initiative: an ecological research agenda. Ecology 72(2): 371-412. Lukatelich, R. J., Bastyan, G., Walker, D. I. and McComb, A. J. (1987). Effect of boat moorings on seagrass beds in the Perth Metropolitan region. Environmental Protection Authority. Technical series No. 21; Perth. MacArthur, R. (1968). The theory of the niche. pp. 159-176. In: Lewontin, R. C. (ed.). Population Biology and Evolution. Syracruse Univeristy. Madsen, T. V. and Warnke, E. (1983). Velocities of currents around and within submerged aquatic vegetation. Archiv fuer Hydrobiologie 97: 389-394. Marbà, N. & Duarte, C.M. (1994). Growth response of the seagrass Cymodocea nodosa to experimental burial and erosion. Marine Ecology Progress Series 107: 307-311. Marbà, N. Gallegos, M. E., Merino, M., Duarte, C. M. (1994). Vertical growth of Thalassia testudinum seasonal and interannual variability. Aquatic Botany 47: 1-11. Marshall, J. K. (1965). Corynephorus canescens (L.) P. Beauv. as a model for the Ammophila problem. Journal of Ecology 53: 447-463. Masini, R. J., Cary, J. L., Simpson, C. J. and McComb, A. J. (1990). Effects of light and temperature on the photosynthesis of seagrasses, epiphytes and macroalgae and implications for management of the Albany harbours. Environmental Protection Authority. Technical Series No. 32; Perth. Maun, M. A. and LaPierre, J. (1984). The effects of burial by sand on Ammophila breviligulata. Journal of Ecology 72: 827-839. Maunder, M., Culham, A. and Hankamer, C. (1998). Picking up the pieces: Botanical conservation on degraded oceanic islands. pp. 317-344. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology: For the Coming Decade. 2nd edition. Chapman and Hall; New York. Mazzella, L., Gambi, M. C., Russo, G. F. and Buia, M. C. (1984). Deep flowering and fruiting of Posidonia oceanica beds around the island of Ischia (Gulf of Naples, Italy). In: Boudouresque, C. F., Jeudy de Grissac, A. and Olivier, J. (eds.) International Workshop on Posidonia oceanica Beds. GIS Posidonie publ., 1: 203-209. Mazzella, L., Gambi, M. C., Russo, G. F. and Wittman, J. (1983). Flowering in Posidonia oceanica (l.) Delile prairies around the island of Ischia (Gulf of Naples). Rapport. Commission Internationale Mer Méditerranee [France] 28: 117-119. McComb, A. J., Cambridge, M. L., Kirkman, H. and Kuo, J. (1981). The biology of Australian seagrasses. pp. 259-291. In: Pate, J. S. and McComb, A. J. (eds.). The Biology of Australian Plants. University of Western Australia Press, Perth. McConchie, C. A. and Knox, R. B. (1989). Pollination and reproductive biology of seagrasses. pp. 74- 111. In: Larkum, A. W. D., McComb, A. J. And Shepherd, S. A. (eds.). Biology of Seagrasses. A

149

Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. McConchie, C. A., Ducker, S. C. and Knox, R. B. (1982). Biology of Australian seagrasses: floral development and morphology in Amphibolis (Cymodoceaceae). Australian Journal of Botany 30: 251-264. McLaughlin, P. A., Treat, S. F. and Thorhaug, A. (1983). A restored seagrass (Thalassia) bed and its animal community. Environmental Conservation 10: 247-254. McMichael, A. J. (1993). Planetary Overload. Global Environmental Change and the Health of the Human Species. Cambridge University; Cambridge. 352pp. McMillan, C. (1983). Seed germination for an annual form of Zostera marina from the sea of Cortez, Mexico. Aquatic Botany 16: 105-110. McMillan, C. (1986). Fruits and seeds of Halophila engelmannii (Hydrocharitaceae) in Texas. Contributions to Marine Science 29: 1-8. McMillan, C. (1987). Seed germination and seedling morphology of the seagrass, Halophila engelmannii (Hydrocharitaceae). Aquatic Botany 28: 179-188. McMillan, C. (1991). The longevity of seeds. Aquatic Botany 40: 195-198. Meinesz, A. and Lefèvre, J. R. (1984). Régénération d’un herbier de Posidonia oceanica quarante années après sa destruction dans la rade de Villefranche (Alpes-Maritimes, France). In: Boudouresque, C. F., Jeudy de Grissac, A. and Olivier, J. (eds.) International Workshop on Posidonia oceanica Beds. GIS Posidonie publication 1: 39-44. Meinesz, A., Caye, G., Loques, F. and Molenaar, H. (1991). Growth and development in culture of orthotropic rhizomes of Posidonia oceanica. Aquatic Botany 39: 367-377. Meinesz, A., Caye, G., Loques, F. and Molenaar, H. (1993). Polymorphism and development of Posidonia oceanica transplanted from different parts of the Mediterranean into the national park of Port-Cros. Botanica Marina 36: 209-216. Meinesz, A., Molenaar, H., Bellone, E. and Loques, F. (1992). Vegetative reproduction in Posidonia oceanica. I Effects of rhizome length and transplantation season in orthotropic shoots. Marine Ecology 13(2): 163-174. Merkel, K. W. (1990a). Eelgrass transplanting in south San Diego Bay, California. Proceedings of the Californian Eelgrass Symposium. Chula Vista, California, May 1988. 28-42pp. Merkel, K. W. (1990b). Growth and survival of transplanted eelgrass: the importance of planting size and spacing. Proceedings of the Californian Eelgrass Symposium. Chula Vista, California, May 1988. 70-78pp. Merkel, K. W. and Hoffman, R. S. (1990). The use of dredged materials in the restoration of eelgrass meadows: a southern Californian perspective. In: Landin, M. S. et al., (eds.) Proceedings of a Regional Workshop: Beneficial Uses of Dredged Material in the Western United States, 21-25 May. San Diego, California. US Army Corps of Engineers, Waterways experiment station, Vicksberg, MS.

150

Mills, D. A. (Editor) (1987). An overview of environmental problems in Princess Royal Harbour and Oyster Harbour, Albany, with a discussion on management options. Environmental Protection Authority Technical Series No. 16; Perth. Molenaar, H. (1992). Etude de la transplantation de boutres de Posidonia oceanica (L.) Delile, phanerogrames marines. Modelisation de l’architecture et du mode de croissance. Ph.D. Dissertation. University Nice-Sophia, Antipolos, France. 1-221. Molenaar, H. and Meinesz, A. (1991). Valorisation biologique des fonds marins littoraux des Alpes Maritimes. Rapport final 1991. Etude realise sous convention entre le Conseil General des Alpes Maritimes et l’ADER-PACA (Association pour la Development de l’Enseigment et de la Recherche en Provence Alpes Cote d’ Azur). Molenaar, H. and Meinesz, A. (1992). Vegetative reproduction in Posidonia oceanica. II Effects on depth changes on transplanted orthotropic shoots. Marine Ecology 13(2): 175-185. Molenaar, H., Meinesz, A. and Caye, G. (1993). Vegetative reproduction in Posidonia oceanica. Survival and development in different morphological types of transplanted cuttings. Botanica Marina 36: 481-488. Molinier, R. and Picard, J. (1952). Recherches sur les herbiers de phanerogrames marines du littoral Mediterranean Francais. Annales Instut Oceanographique 27: 157-234. Moore, K.A., Orth, R.J. and Nowak, J.F. (1993). Environmental regulation of seed germination in Zostera marina L. (eelgrass) in Chesapeake Bay: effects of light, oxygen and sediment burial. Aquatic Botany 45: 79-91. Moorhouse, M. (1996). Whose heritage Hinchinbrook? Wildness News Issue No. 146: 1 & 2. Moriarty, D. J. W. and Boon, P. I. (1989). Interactions of seagrasses with sediment and water. pp. 500- 535. In Larkum, A. W. D., McComb, A. J. and Shepherd, S. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. National Research Council (1995). Wetlands: Characteristics and Boundaries. National Research Council, Washington, USA. Nelson, M. L. (1992). Environmental Characteristics Affecting Revegetation of Seagrasses in Owen Anchorage, Western Australia. B.Sc. Honours Thesis. Environmental Science, Murdoch University. Perth, Western Australia, Australia. Neverauskas, V. (1988). Response of a Posidonia community to prolonged reduction in light. Aquatic Botany 31: 361-366. Neverauskas, V. P. (1985a). Port Adelaide Sewerage Treatment Works Sludge Outfalls. Effect of Discharge on the Adjacent Marine Environment. Progress report, July 1982-May 1984. EWS Rep. 85/6. 32pp. Neverauskas, V. P. (1985b). Effects of the Port Adelaide treatment works sludge discharge on the adjacent marine environment. Proceedings of the 1985 Australasian Conference on Coastal Ocean Engineering. 1: 193-202.

151

Nichols, O. G., Carbo, B. A., Colquhoun, I. J., Croton, J. T. and Murray, N. J. (1985). Rehabilitation after bauxite mining in south-western Australia. Landscape Planning 12: 75-92. Nielsen, P. (1992). Coastal Bottom Boundary Layers and Sediment Transport. Volume 4. World Scientific; Singapore. 161-200pp. Nienhuis, P.H. and Groenendijk, A. M. (1986). Consumption of eelgrass (Zostera marina) by birds and invertebrates: an annual Budget. Marine Ecology Progress Series 29: 29-35. Nitsos, R. (1990). Morro Bay eelgrass transplant. Proceedings of the Californian Eelgrass Symposium. Chula Vista, California, May 1988. 43-46pp. Nomme, K. M. and Harrison, P. G. (1991). Evidence for interaction between the seagrass Zostera marina and Zostera japonica on the Pacific coast of Canada. Canadian Journal of Botany 69(9): 2004- 2010. Norse, E. A. (ed.) (1993). Global Marine Biological Diversity. A Strategy for Building Conservation into Decision Making. Island Press; Washington. 384pp. Olesen, B. and Sand-Jensen, K. (1993). Seasonal acclimation of eelgrass Zostera marina growth to light. Marine Ecology Progress Series 94: 9-99. Olesen, B. and Sand-Jensen, K. (1994). Patch dynamics of eelgrass Zostera marina. Marine Ecology Progress Series 106: 147-156. Olivieri, I. and Gouyon, P. (1997). Evolution of migration rate and other traits: the metapopulation effect. pp. 293-323. In: Hanski, I. and Gilpin, M. E. (eds.) Metapopulation Biology. Ecology, Genetics, and Evolution. Academic Press; San Diego. Orth, R. J. (1985). Submerged aquatic vegetation in the Chesapeake Bay: value, trends and management. In: Proceedings of the Symposium on Wetlands of the Chesapeake. Contrib. 1234, VIMS. 84- 106pp. Orth, R. J. and Moore, K. A. (1983). Chesapeake Bay, an unprecedented decline in submerged aquatic vegetation. Science 222: 51-53. Orth, R. J. and Moore, K. A. (1988). Distribution of Zostera marina L. and Ruppia maritima L. sensu lato along depth gradients in the lower Chesapeake Bay, U.S.A.. Aquatic Botany 32: 291-305. Orth, R.J., Lukenbach, M. and Moore, K.A. (1994). Seed dispersal in a marine macrophyte: implications for colonisation and restoration. Ecology 75(7): 1927-1939. Ott, J. A. (1990). Biomass. pp 55-64. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Paling, E. I. (1986). Cited in Paling E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Paling, E. I. (1992). The Relationship between Productivity and Nutrient Cycling in Macroalgal Stands and Seagrass Meadows. Ph.D. Dissertation. Botany Department, University of Western Australia, Western Australia, Australia. 330pp. Paling, E. I. (1994). Nitrogen and phosphorus uptake in seedlings of the seagrass Amphibolis antarctica in Western Australia. Hydrobiologia 294:1-4.

152

Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Unpublished report. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Paling, E. I., Walker, D. I. and van Keulen, M. (1997). Seagrass Rehabilitation in Owen Anchorage, Western Australia. Annual report August 1996 – August 1997. Environmental Science Report No. MAFRA 97/8, Murdoch University, Western Australia. Parsons, T. R., Taakahashi, M. and Hargrave, B. (1977). Biological Oceanographic Processes. 2nd edition. Pergamon Press; Oxford. Patriquin, D. G. (1973). Estimation of growth rate, production and age of the marine angiosperm Thalassia testudinum Konig. Caribbean Journal of Science 13: 111-123. Patriquin, D. G. (1975). ‘Migration’ of blowouts in seagrass beds at Barbados and Carriacou, West Indies, and its ecological and geological implication. Aquatic Botany 1: 163-189. Pearce, D. and Moron, D. (1998). The economics of biological diversity conservation. pp. 384-409. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology: For the Coming Decade. 2nd edition. Chapman and Hall; New York. Pearson, L. C. (1995). The Diversity and Evolution of Plants. CRC Press; Boca Raton. 646pp. Pergent, G. and Pergent-Martini, C. (1990). Some applications of lepidochronological analysis in the seagrass Posidonia oceanica. Botanica Marina 33: 299-310. Pergent, G. and Pergent-Martini, C. (1991). Leaf renewal cycle and primary production of Posidonia oceanica in the bay of Lacco Ameno (Ischia, Italy) using lepidochronological analysis. Aquatic Botany 42: 49-66. Pergent, G., Benmaiz, N., Boudouresque, C. F. and Meinesz, A. (1989). The flowering of Posidonia oceanica over the past fifty years: a lepidochronological study. In: Boudouresque, C. F., Meinesz, A., Fresi, E. and Gravez, V. (eds.) International Workshop on Posidonia oceanica Beds. GIS Posidonie publ., 2: 69-76. Pergent, G., Romero, J., Pergent-Martini, C. A., Mateo, and M., Boudouresque, C. F. (1994). Primary production, stocks and fluxes in the Mediterranean seagrass Posidonia oceanica. Marine Ecology Progress Series 106: 139-146. Pergent-Martini, C., Rico-Raimondino, V. and Pergent, G. (1994). Primary production of Posidonia oceanica in the Mediterranean basin. Marine Biology 120: 9-15. Phillips, R. C. (1972). Ecological Life History of Zostera marina C. (Eelgrass) in Puget Sound, Washington. Ph.D. Dissertation. University of Washington; Seattle, USA. 154pp. Phillips, R. C. (1974). Transplantation of seagrasses, with special emphasis on eelgrass, Zostera marina L. Aquaculture 4: 161-176. Phillips, R. C. (1978). Role of seagrasses in estuarine systems. In: Lewis, R. R. (ed.). Creation and Restoration of Coastal Plant Communities. CRC press; Boca Raton. Phillips, R. C. (1980). A Handbook of Seagrass Biology: An Ecosystem Perspective. Garland; New York. Phillips, R. C. (1982). Seagrass meadows. In: Lewis, R.R. (ed). Creation and Restoration of Coastal Plant Communities. CRC; Boca Raton.

153

Phillips, R. C. (1990a). Responses to environmental conditions. pp. 43-46. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Phillips, R. C. (1990b). Transplant methods. pp 51-60. In: Phillips, R. C. and McRoy, C. P. (eds.). Seagrass Research Methods. UNESCO; France. Phillips, R. C. (1996). Ecology of eelgrass (Zostera marina L.) transplants in Izembek Lagoon, Alaska. pp. 333-338. In: Kuo, J., Phillips, R. C., Walker, D. I. and Kirkman, H. (eds.). Seagrass Biology. Proceedings of an International Workshop. Rottnest Island, Western Australia. 25-26 January 1996. Faculty of Science, University of Western Australia; Perth.. Phillips, R. C. and Shaw, R. F. (1976). Zostera noltii Hornem. in Washington, U. S. A. Syesis 9: 355-358. Phillips, R. C., McMillan, C. and Bridges, K. W. (1983). Phenology of eelgrass, Zostera marina L., along latitudinal gradients in North America. Aquatic Botany 15: 145-156. Pimm, S. L. (1991). The Balance of Nature? Ecological Issues in the Conservation of Species and Communities. University of Chicago Press; Chicago. 434pp. Pirc, H. (1984). Depth adaptation in Posidonia oceanica. pp. 227-234. In: Boudoureque, C. F., de Grissac, J., Olivier, O. (eds.) International Workshop on Posidonia oceanica Beds. G. I. S. Posidonie publ. France Marseilles 1. Pirc, H. (1986). Seasonal aspects of photosynthesis in Posidonia oceanica: influence of depth, temperature and light intensity. Aquatic Botany 26: 203-212. Pirc, H., Buia, M. C. and Mazzella, L. (1986). Germination and seedling development of Cymodocea nodosa (Ucria) Ascherson under laboratory conditions. Aquatic Botany 26: 181-188. Poiner, I. and Conacher, C. (1992). To develop techniques of moving intertidal marine sediments supporting Zostera capricorni seagrass meadows, using conventional earth-moving equipment, so as to guarantee the survival/growth of the seagrass meadows. Report to Leighton Contractors Pty Ltd. CSIRO Marine Laboratories Report, August 1992. CSIRO; Cleveland. Poiner, I. R., Walker, D. I. and Coles, R. G. (1989). Regional studies – seagrasses of tropical Australia. pp. 279-303. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Posey, M. H. (1988). Community changes associated with the spread of an introduced seagrass, Zostera japonica. Ecology 69(4): 974-983. Ramage, D. L. and Shiel, D. R. (1998). Reproduction in the seagrass Zostera novazelandica on intertidal platforms in southern New Zealand. Marine Biology 130:479-489. Ramus, J. (1978). Seaweed anatomy and photosynthetic performance: The ecological significance of light guides, heterogeneous absorption and multiple scatter. Journal of Phycology 14: 352 – 362. Ramus, J., Beale, S. I. and Mauzerall, D. (1976a). Correlation of changes in pigment content with photosynthetically capacity of seaweeds as function of water depth. Marine Biology 37: 231 – 238. Ramus, J., Beale, S. I., Mauzerall, D. and Howard, K. L. (1976b). Changes in photosynthetic pigment concentration in seaweeds as a function of water depth. Marine Biology 37: 223 – 229.

154

Ranwell, D. S., Wyer, D. W., Boorman, L. A., Pizzey, J. M. and Waters, R. J. (1974). Zostera transplants in Norfolk and Suffolk, Great Britain. Aquaculture 4: 185-198. Raupach, M. R. and Thom, A. S. (1981). Turbulence in and above plant canopies. Annual Review of Fluid Mechanics 13: 97-129. Raven, P. H., Evert, R. F. and Eichhorn, S. E. (1986). Biology of Plants. 4th edition. Worth; New York: 775pp. Rea, N. and Ganf, G. G. (1994). The role of sexual reproduction and water regime in shaping the distribution patterns of clonal emergent aquatic plants. pp. 101-111. In: Brock, M. A., Boon, P. I. and Grant, . (eds.). Plants and Processes in Wetlands. CSIRO; Victoria. Rees, M. (1997). Evolutionary ecology of seed dormancy. pp. 121-142. In: Silvertwon, J., Franco, M. and Harper, J. L. (eds.). Plant Life Histories. Ecology, Phylogeny and Evolution. The Royal Soceity, Cambridge University; Cambridge. Ridley, M. (1996). Evolution. 2nd edition. Blackwell Science; Massachusetts. 719pp. Robblee, M. B., Barber, T. R., Carlson, P. R. Jr, Durako, M. J., Fourqurean, J. W., Muehlstein, L. K., Porter, D., Yarbro, L. A., Ziemen, R. T. and Ziemen, J. C. (1991). Mass mortality of the tropical seagrass Thalassia testudinum in Florida Bay (USA). Marine Ecology Progress Series 71: 297-299. Roberts, M.H., Orth, R.J. and Moore, K.A. (1984). Growth of Zostera marina L. seedlings under laboratory conditions of nutrient enrichment. Aquatic Botany 20: 321-328. Robertson, A. I. and Mann, K. H. (1984). Disturbance by ice and life-history adaptations of the seagrass Zostera marina. Marine Biology 80: 131-142. Robilliard, G. A. and Porter, P. E. (1976). Transplantation of Eelgrass (Zostera marina) in San Diego Bay. Undersea Sciences Department, Naval Undersea Center, San Diego, U.S.A. Rogers, R. G. and Bisterfield, F. T. (1975). In Proceedings of the 2nd annual conference on restoration of coastal vegetation in Florida. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Ruckelshaus, M. H and Hays, C. G. (1998). Conservation and Management of species in the sea. Pp. 112-156. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology: For the Coming Decade. 2nd edition. Chapman and Hall; New York. Ruckelshaus, M. H. (1994). Ecological and Genetic Factors Affecting Population Structure in the Marine Angiosperm, Zostera marina L. Ph.D. Dissertation. University of Washington, USA. 208pp. Safriel, U. N. and Sansson-Frostig, Z. (1988). Can colonizing mussel outcompete indigenous mussel? Journal of Experimental Marine Biology and Ecology 117: 211-226. Salisbury, F. B. and Ross, C. W. (1992). Plant Physiology. 4th edition. Wadsworth; California. 207-224, 504-530pp. Sand-Jensen, K. (1977). Effects of epiphytes on eelgrass photosynthesis. Aquatic Botany 3: 55-63. Sand-Jensen, K. and Søndergaard, M. (1979). Distribution and quantitative development of aquatic macrophytes in relation to sediment characteristics in oligotrophic Lake Kalgaard, Denmark. Freshwater Biology 9: 1-11.

155

Schemske, D. W., Husband, B. C., Ruckelshaus, M. H., Goodwillie, C. Parker, I. M. and Bishop, J. G. (1994). Evaluating approaches to the conservation of rare and endangered plants. Ecology 75(3); 584-606. Schwarzschild, A. C., MacIntyre, W. G., Moore, K. A. and Libelo, E. L. (1994). Zostera marina L. growth response to atrazine in root-rhizome and whole plant exposure experiments. Journal of Experimental Marine Biology and Ecology 183: 77-89. Scott, J.K., Dybdahl, R. and Wood, W. F. (1986). The ecology of Posidonia seagrass fish communities in Cockburn Sound. Department of Conservation and Environment. Technical series No. 11. Sealey, R. V., Williams, M. L. and Cobb, A. H. (1990). Adaptations of Codium fragile (Suringer) Hariot fronds to photosynthesis at varying flux density. pp. 455–458. In: Baltscheffsky, M. (ed.). Current Research in Photosynthesis. Volume II. The Hague: Kluwer Academic Publishers Shepherd, S. A., McComb, A. J., Bulthuis, D. A., Neverauskas, V., Steffensen, D. A. and West, R. (1989). Decline of seagrasses. pp. 346-393. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Short, F. T. (1981). Nitrogen Resource Analysis and Modelling of an Eelgrass (Zostera marina L.) Meadow in Izembek Lagoon, Alaska. Ph.D. Dissertation. 173pp. Short, F. T. (1983a). The response of interstitial ammonium in eelgrass (Zostera marina L.) beds to environmental perturbations. Journal of Experimental Marine Biology and Ecology 68: 195-208. Short, F. T. (1983b). The seagrass, Zostera marina L.: plant morphology and bed structure in relation to sediment ammonium in Izembek Lagoon, Alaska. Aquatic Botany 16: 149-161. Short, F. T. (1987). Effects of sediment nutrients on seagrasses: Literature review and mesocosm experiment. Aquatic Botany 27: 41-57. Short, F. T. and McRoy, C. P. (1984). Nitrogen uptake by leaves and roots of the seagrass Zostera marina L. Botanica Marina 17: 547-555. Short, F. T., Davis, M. W., Gibson, R. A. and Zimmerman, C. F. (1985). Evidence for phosphorous limitation in carbonate sediments on the seagrass Syringodium filiforme. Estuarine and Coastal Shelf Science 20: 419-430. Short, F. T., Montgomery, J., Zimmerman, C. F. and Short, C. A. (1993). Production and nutrient dynamics of a Syringodium filiforme Kütz. seagrass bed in Indian River Lagoon, Florida. Estuaries 16(2): 323-334. Short, F. T. and Wyllie-Echeverria, S. (1996). Natural and human induced disturbance of seagrasses. Environmental Conservation 23(1): 17-27. Silberstein, K., Chiffings, A. W. and McComb, A. J. (1986). The loss of seagrass in Cockburn Sound, Western Australia. III. The effects of epiphytes on productivity of Posidonia australis Hook. f. Aquatic Botany 24: 355-371. Sinclair, Knight and Partners (1993). Emu Point Coastal Protection Study. Town of Albany, Western Australis. Sinclair Knight and Associates, Oceanroutes (Australia) Pty Ltd and F. L. Wilkinson and Associates.

156

Smith, F.E. (1951). Tetrazolium salt. Science 113: 751-754. State of the Environment Advisory Council (1996). Australia: State of the Environment 1996. CSIRO; Victoria. Steedman, R.K. and Craig, P.D. (1983). Wind-driven circulation of Cockburn Sound. Australian Journal of Marine and Freshwater Research 34: 187-212. Summerson, H. C. and Peterson, C. H. (1984). Role of predation in organising benthic communities of a temperate-zone seagrass bed. Marine Biology Progress Series 15: 63-77. Sverdrup, H. U., Johnson, M. W. and Fleming, R. H. (1942). The Oceans. Their Physics, Chemisty and General Biology. Prentice-Hall, New Jersey. 1087pp. Terrados, J. (1993). Sexual reproduction and seed banks of Cymodocea nodosa (Ucria) Ascherson meadows on the southeast Mediterranean coast of Spain. Aquatic Botany 46: 293-299. Thayer, G. W., Kenworthy, W. J., Fonseca, M. S. (1984). The Ecology of Eelgrass Meadows of the Atlantic Coast: a Community Profile. U.S. Fish and Wildlife Service, FWS/OBS-84/02. Thayer, G. W., Wolfe, D. A. and Williams, R. B. (1975). The impact of man on seagrass ecosystem. American Scientist 63: 288. Thelin, I. and Boudouresque, C. F. (1985). Posidonia oceanica flowering and fruiting. Recent data from an international inquiry. Posidonia Newsletter 1: 5-14. Thom, R. M. (1990). A review of eelgrass (Zostera marina L.) transplanting projects in the Pacific Northwest. The Northwest Environmental Journal 6: 121-137. Thorhaug, A. (1974). Transplantation of the seagrass Thalassia testudinum Konig. Aquaculture 4: 177- 183. Thorhaug, A. (1983). Habitat restoration after pipeline construction in a tropical estuary: seagrasses. Marine Pollution Bulletin 14: 422-425. Thorhaug, A. (1985). Large-scale seagrass restoration in a damaged estuary. Marine Pollution Bulletin 16: 55-62. Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110-117. Thorhaug, A. (1987). Large-scale seagrass restoration in a damaged estuary. Marine Pollution Bulletin 18(8): 442-446. Thorhaug, A. and Hixon, R. (1975). In Proceedings of the Second Annual Conference on Restoration and Creation of Coastal Cegetation in Florida. Hillsborough community college. Tampa, Florida. Thorhaug, A. and Miller, B.(1986). Environmental Conservation. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15: 110-117. Thorhaug, A., Miller, B., Jupp, B. and Booker, F. (1985). Effects of a variety of impacts on seagrass restoration in Jamaica. Marine Pollution Bulletin 16(9): 355-360. Thorogood, C. A., Poiner, I. R., Somers, I. F. and Staples, D. J. (1990). Seagrass and Cyclones in the Western Gulf of Carpentaria. Information Sheet No. 7. CSIRO; Cleveland. Tietenberg, T. (1992). Environmental and Natural Resource Economics. 3rd edition. Harper Collins Publisheres; New York. 678pp.

157

Tomasko, D. A. and Lapointe, B. E. (1991). Productivity and biomass of Thalassia testudinum as related to water column nutrient availability and epiphyte levels: field observations and experimental studies. Marine Ecology Progress Series 75: 9-17. Tomlinson, P. B. (1974). Vegetative morphology and meristem dependence - the foundation of productivity in seagrasses. Aquaculture 4: 107-130. Trowbridge, C. D. (1998). Assessment of the spread of the introduced green macroalga Codium fragile ssp tomentosoides on Australian shores. Interim Report December 1998. Centre for Research on Introduced Marine Pests, Technical Report. CSIRO; Hobart. Turner, T. (1985). Stability of rocky intertidal surfgrass beds: persistence, preemption and recovery. Ecology 66(1): 83-92. Tutin, T. G. (1938). The autecology of Zostera marina in relation to its wasting disease. New Phytologist 37: 50-71. USNRC (1992). Restoration of Aquatic Ecosystems. Science, Technology and Public Policy. Committee on restoration of aquatic ecosystems: Science, Technology and Public Policy. National Academy Press; Wahsington DC. 552pp. Valentine, J. F., Heck, K. L., Harper, P. and Beck, M. (1994). Effects of bioturbation in controlling turtlegrass (Thalassia testudinum Banks ex Konig) abundance: evidence from field enclosures and observations in the Northern Gulf of Mexico. Journal of Experimental Marine Biology and Ecology 178: 181-192. Valiela, I. (1991). Ecology of water columns. pp. 29-56. In: Barnes, R. S. K. and Mann, K. H. (eds.). Fundamentals of Aquatic Ecology. Blackwell; London. Valiela, I. (1995). Marine Ecological Processes. 2nd edition. Springer; New York. 686pp. Venable, D. L. (1988). Modeling the evolutionary ecology of seed banks. pp. 67-87. In: Leck, M. A., Parker, V. T. and Simpson, R. L. (eds.). Ecology of Soil Seed Banks. Academic Press; San Diego. Venable, D. L., Brown, J. S. (1988). The selective interactions of dispersal, dormancy, and seed size as adaptations for reducing risk in variable environments. American Naturalist 131: 360-384. Vermaat, J. E., Beijer, J. A. J., Gijlstra, R., Hootsman, M. J. M., Philippart, C. J. M., van der Brink, N. W. and van Vierssen, W. (1993). Leaf dynamics and standing stocks of intertidal Zostera noltii Hornem. And Cymodocea nodosa (Ucria) Ascherson on the Banc d’ Arguin (Mauritania). Hydrobiologia 258: 59-72. Vitousek, P. M., Mooney, H. A., Lubchenco, J. and Melillio, J. M. (1997). Human domination of Earth’s ecosystems. Science 227: 494-499. Walker, C. J. (1994). The Effect of Size and Sediment Stabilisation on the Survival of Seagrass Transplants in Owen Anchorage, Western Australia. B.Sc. Honours Thesis. Environmental Science, Murdoch University. Perth, Western Australia, Australia. Walker, D. I. (1989). Regional studies – seagrass in Shark Bay, the foundations of an ecosystem. pp. 182-210. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam.

158

Walker, D. I. (1992). Cited in Paling, E. I. (1995). Seagrass meadow regrowth, transplantation and recovery after disturbance in Western Australia: A review. Marine and Freshwater Research Laboratory, Murdoch University. Perth, Western Australia. Walker, D. I. and McComb, A. J. (1988). Seasonal variation in the production, biomass and nutrient status of Amphibolis antarctica and Posidonia australis Hook. F. in Shark Bay, Western Australia. Aquatic Botany 31: 259-275. Walker, D. I. and McComb, A. J. (1992). Seagrass degradation in Australian coastal waters. Marine Pollution Bulletin 25(5-8): 191-195. Walker, D. I., Hutchings, P. A. and Wells, F. E. (1991). Seagrass, sediment and infauna - a comparison of Posidonia australis, Posidonia sinuosa and Amphibolis antarctica, in Princess Royal Harbour, south-western Australia. I. Seagrass biomass, productivity and contribution to sediments. pp. 597- 610. In: Wells, F. E., Walker, D. I., Kirkman, H. and Lethbridge, R. (eds.). The Marine Flora and Fauna of Albany, Western Australia. Volume 2. Western Australian Museum; Perth. Walker, D. I., Lukatelich, R. J., Bastyn, G. and McComb, A. J. (1989). Effect of boat moorings on seagrass beds near Perth, Western Australia. Aquatic Botany 36: 69-77. Wallen, B. (1980). Changes in structure and function of Ammophila during primary succession. Oikos 34: 227-238. Ward, T., Butler, E. and Hill, B. (1998). Environmental Indicators for the National State of the Environment Reporting. Estuaries and the Sea. Australia State of the Environmental Indicator Reports. Canberra. 80pp. Ware, R. (1993). Eelgrass (Zostera marina) in southern California bays and wetlands with emphasis on Orange County, California. Shore and Beach 20-30pp. Waycott, M. (1995). Assessment of genetic variation and clonality in the seagrass Posidonia australis using RAPD and allozyme analysis. Marine Ecology Progress Series 116: 289-295. Waycott, M. and Les, D. H. (1996). An integrated approach to the evolutionary study of seagrasses. pp. 71-78. In: Kuo, J., Phillips, R. C., Walker, D. I. and Kirkman, H. (eds.). Seagrass Biology. Proceedings of an International Workshop. Rottnest Island, Western Australia. 25-26 January 1996. Faculty of Science, University of Western Australia; Perth. West, R. J. (1980). A Study of Growth and Primary Production of the Seagrass Posidonia australis Hook f. Masters Thesis. University of Sydney; Sydney, Australia. West, R. J. (1983). The seagrasses of New South Wales estuaries and embayments. Wetlands (Australia) 3: 34-44. West, R. J. (1990). Depth-related structural and morphological variations in an Australian Posidonia seagrass bed. Aquatic Botany 36: 153-166. West, R. J. Jacobs, N. E. and Roberts, D. E. (1990). Experimental transplanting of seagrasses in Botany Bay, Australia. Marine Pollution Bulletin 21(4): 197-203. West, R. J. and Larkum, A. W. D (1979). Leaf productivity of the seagrass Posidonia australis in eastern Australian waters. Aquatic Botany 7:57-65.

159

West, R. J., Larkum, A. W. D. and King, R. J. (1989). Regional studies – seagrasses of south eastern Australia. pp. 230-260. In: Larkum, A. W. D., McComb, A. J. and Shepherd, A. A. (eds.). Biology of Seagrasses. A Treatise on the Biology of Seagrasses with Special Reference to the Australian Region. Elsevier; Amsterdam. Westoby, M. Leishman, M. and Lord, J. (1997). Comparative ecology of seed size and dispersal. pp. 143-162. In: Silvertwon, J., Franco, M. and Harper, J. L. (eds.). Plant Life Histories. Ecology, Phylogeny and Evolution. The Royal Soceity, Cambridge University; Cambridge. Williams, S. L. (1988). Disturbance and recovery of a deep-water Caribbean seagrass bed. Marine Ecology Progress Series 42: 63-71. Williams, S. L. (1990). Experimental studies of Caribbean seagrass bed development. Ecological Monographs 60: 449-469. Williams, S. L. and Davis, C. A. (1996). Population genetic analyses of transplanted eelgrass (Zostera marina) beds reveal reduced genetic diversity in Southern California. Restoration Ecology 4(2): 163-180. Winer, B. J. (1971). Statistical Principles in Experimental Design. McGraw-Hill; Tokyo. 907pp. Winter, M. (1976). Culture studies of the seagrass Ruppia maritima L.. New York sea grant. Garden City, U.S.A. From Thorhaug, A. (1986). Review of seagrass restoration efforts. Ambio 15(2): 110- 117. WRI, IUCN and UNEP (World Resources Institute, World Conservation Union and United Nations Environment Programme) (1992). Global Biodiversity Strategy. World Resources Institute; Washington. Wyant, J. G., Meganck, R. A. and Ham, S. H. (1995). The need for an environmental restoration decision framework. Ecological Engineering 5: 417-420. Zar, J. H. (1996). Biostatistical Analysis. Third edition. Prentice-Hall International. Sydney. Zedler, J. (1998). Replacing endangered species habitat: The acid test for wetland ecology. pp. 364-380. In: Fiedler, P. L. and Kareiva, P. M. (eds.). Conservation Biology: For the Coming Decade. 2nd edition. Chapman and Hall; New York. Zhang, J. and Maun, M. A. (1990). Effects of sand burial on seed germination, seedling emergence, survival and growth of Agropyron psammophilum. Canadian Journal of Botany 68: 304-310. Ziemen, J. C. and Wetzel, R. G. (1980). Productivity in seagrasses: methods and rates. In Phillips, R. C. And McRoy, C. P. (eds.). A Handbook of Seagrass Biology: an Ecosystem Perspective. Garland; New York. Zimmerman, R. C., Cabello-Pasini, A. & Alberte, R. S. (1994). Modeling daily production of aquatic macrophytes from irradiance measurements: a comparative analysis. Marine Ecology Progress Series 114: 185-196. Zimmerman, R. C., Reguzzoni, J. L., Wyllie-Echeverria, S., Josselyn, M. and Alberte, R. S. (1991). Assessment of environmental suitability for growth of Zostera marina L. (eelgrass) in San Francisco Bay. Aquatic Botany 39: 353-366.

160

Zimmerman, R. C., Reguzzonia, J. L. and Alberte, R. S. (1995). Eelgrass (Zostera marina L.) transplants in San Francisco Bay: Role of light availability on metabolism, growth and survival. Aquatic Botany 51: 67-86.

161 Appendices

Appendix A

Selected seagrass transplantation studies from the past 30 years (1966-1996) reviewed in this study. a = cited in Thorhaug 1986; b = cited in Paling 1995; c = cited in Kirkman 1989 and; d = Butler and Jernakoff 1999. Asterisked (*) efforts are used in Table 1.1. Transplant units (TU): s = seeds; sd = seedlings; sp = sprigs; p = plugs; and r = rhizome. Habitat enhancement (HE): A = anchors; AB = artificial barriers; M = mesh; ASG = artificial seagrass mats; SH = shading, and; F = fertiliser. ? = unspecified.

Species Study TU HE Location Success Amphibolis spp. Kirkman 1989* sd M Cockburn Sound, 0-22% Warnbro Sound 0-47% & Rottnest 0% Island, Western Australia (WA) Kirkman 1990 sd, r M Warnbro Sound 13–75% & Rottnest Island, WA Hancock 1992* sp A, M Cockburn Sound, Low success WA Kirkman 1995* sd A, M Cockburn Sound, < 15% unpublished WA Walker 1985* sd, p M Cockburn Sound 0 - 100% Warnbro Sound, WA Walker 1992* sd, p M Cockburn Sound 0 - 100% Warnbro Sound, WA Walker 1994* sp, p None Cockburn Sound, 50-70% WA Paling 1995* p None Success Bank, 40% unpublished WA Paling et al. p None Success Bank, 97% 1997* WA unpublished Halodule Kelly et al. 1971a sp A Florida, US Not implicitly wrightii stated in citation Eleuterius 1975 a s A, M Mississippi, US Not implicitly stated in citation Derrenbecker Not Not Florida, US 44-98% and Lewis 1983* stated stated Thorhaug 1983* sp, sd A Florida, US Successful Durako et al. sd, sp None Florida, US Not stated 1992

162 Species Study TU HE Location Success Durako and sp None Florida, US Not stated Moffler 1984 Fonseca et al. s, p None Virginia & North Not stated 1985 Carolina, US Thorhaug 1985* sp A Florida, US Successful Thorhaug et al. s, sp, p None Jamaica 0-100% 1985* Fonseca et al. s, p None Florida & north- Not stated 1987 east Gulf of Mexico Thorhaug 1987* sp ? Florida, US Successful Tomasko and sp ? Florida, US Successful Lapointe 1991* Halophila Williams 1988* r ? US Virgin Successful decipiens Islands Posidonia Larkum 1976 p ? Botany Bay, Inconclusive australis and P. New South sinuosa Wales (NSW) Cambridge sd A Cockburn Sound, P. australis 70% 1977* Warnbro Sound, P. sinuosa 0- WA 50% Cambridge sd A Cockburn Sound, 98-100% 1978* Warnbro Sound, WA Cambridge sd A Cockburn Sound, 0-70% 1980* Warnbro Sound, WA Paling 1986b sp, r A Owen Not implicitly Anchorage, WA stated Kirkman 1989* sd M Cockburn Sound, 0% Rottnest Island, 67% WA LeProvost 1990* sd A Cockburn Sound, 0% WA West et al. 1990* sp, p A, Botany Bay, Unsuccessful ASG NSW Hancock 1992* r A Cockburn Sound, Unsuccessful WA Paling 1992* sp A, M Cockburn Sound, Unsuccessful WA Walker 1994* sp, p None Cockburn Sound, 80% WA Kirkman 1995* sd A, M Cockburn Sound Unsuccessful WA Paling 1995* sd, p None Success Bank, 95% unpublished WA Paling et al. sd, p None Success Bank, 85% 1997* WA

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Species Study TU HE Location Success Bastyan 1997*, r None Oyster Harbour, >50% unpublished WA Posidonia Cooper 1976a s, p None France Not implicitly oceanica stated in citation Cooper 1977a s, p None France Not implicitly stated in citation Cooper 1978ab s, p None France Not implicitly stated in citation Cooper 1979 s, p AB France Not implicitly stated Molenaar and sp None France Successful Meinesz 1991* Molenaar and sp None France 96-100% Meinesz 1992* Molenaar 1992* sp None France Successful Ph.D. thesis Meinesz et al. sd, sp None France Seedlings low 1992* Rhizomes high Meinesz et al. sp None France 92-97% 1992* Molenaar et al. sp None France 31-100% 1993* Genot et al. sp None Monaco 80-92% 1994* Ruppia maritima Winter 1976a p None Unspecified Not implicitly stated in citation Bird et al. 1994* sd, sp A North Carolina, Cultured plants US 0% Vitro plants 20- 80% Syringodium van Breedveld p None Florida, US 20-100% filiforme 1975* Derrenbecker sd, sp A Florida, US 44-98% and Lewis 1983* Orth and Moore s None Chesapeake Bay, Not stated 1983 US Churchill 1983a s None Unspecified Not implicitly stated in citation Durako and sd, p None Florida, US Not stated Moffler 1984 Thorhaug et al. s, p None Jamaica 0-100% 1985* Thorhaug 1985* s, sp, p None Florida, US Seedlings successful Fonseca et al. sp M Virginia, North Not stated 1985 Carolina, US

164 Species Study TU HE Location Success Thorhaug 1987* sd, sp, A Florida, US Successful p Lewis 1987* p None Florida, US 15% Fonseca et al. sp None Florida, US & Successful 1987* north-east Gulf of Mexico Tomasko and sp None Florida, US Successful Lapointe 1991* Durako et al. sp None Florida, US Not stated 1992 Thalassia Fuss & Kelly sp A Florida, US Not stated testudinum 1969 Thorhaug 1974* s A Florida, US Successful Thorhaug and s None Florida, US Unsuccessful Hixon 1975* Eleuterius 1975a s A, M Mississippi, US Not implicitly stated in citation van Breedveld p None Florida, US 20-100% 1975* Churchill 1983a s None Unspecified Not implicitly stated in citation Derrenbecker sd, sp A Florida, US 44-98% and Lewis 1983* Orth and Moore s None Chesapeake Bay, Not implicitly 1983 US stated Thorhaug 1983* sd, sp None Florida, US Successful McLaughlin et sd None Florida, US Successful al. 1983* Durako and sd, p None Florida, US Unsuccessful Moffler 1984* Thorhaug et al. s, p None Jamaica 0-90% 1985* Thorhaug 1985* s, sp, p None Florida, US Seedlings successful Thorhaug and p None Unspecified Not implicitly Miller 1986a stated in citation Fonseca et al. sp None Florida, US & Successful 1987* north-east Gulf of Mexico Lewis 1987* p None Florida, US 100% Thorhaug, 1987* sd, sp, A Florida, US Successful p Tomasko and sp None Florida, US Successful Lapointe 1991* Durako et al. sp None Florida, US Not stated 1992

165 Species Study TU HE Location Success Zostera Harris et al. 1980 ? None Illawarra Lake, Not stated capricorni NSW Lee Long 1990d* p None Cairns Harbour, Survival but no Queensland expansion (QLD) West et al. 1990* sp, p A, Botany Bay, Unsuccessful ASG NSW

Poiner and p None Raby Bay & Successful Conacher 1992* Victoria Point, QLD P. Gibbs pers. r ASG Botany Bay, Moderately comm.* NSW successful Zostera japonica Harrison 1987* p None British Unsuccessful Columbia, Canada Nomme and p SH British Inconclusive Harrison 1991 Columbia, results Canada Zostera marina Ranwell et al. p None Norfolk & Successful 1974* Suffolk, Great Britain Phillips 1974* s, p A Alaska & Puget Successful Sound, Washington, US Robilliard and p None Unspecified Not stated Porter 1976 Churchill et al. sp None Unspecified Not implicitly 1978a stated in citation Fonseca et al. sp M North Carolina, Successful 1979* US Fonseca et al. sp A Virginia, US Successful 1982* Churchill 1983a s None Unspecified Not implicitly stated in literature Orth and Moore s None Chesapeake Bay, Not stated 1983 US Orth 1985* p F Chesapeake Bay, Successful US Fonseca et al. sp None Dredge Island, 0% 1990* US Harrison 1990* sp, p None British Successful Columbia, Canada Hoffman 1990* p A California, US Unsuccessful Merkel 1990a* p A and California, US Successful M

166

Species Study TU HE Location Success Merkel 1990b* p None California, US Successful Merkel and p None Morro Bay, US 84-97% Hoffman 1990* Nitsos 1990* sp A California, US Successful

Thom 1990* sp, p A San Francisco Unsuccessful Bay, US Nomme and p SH British Inconclusive Harrison 1991 Columbia, Canada Batuik et al. sp, p None Chesapeake Bay, 11-96% 1992* US Fonseca and sp, p A North Carolina, 33-100 % Kenworthy US 1992* Ware 1993* s, sp None California, US Sprigs and anchors 80% 17% plugs Zimmerman et p None San Francisco, 10-60% al. 1995* US Phillips 1996* p None Izembek Lagoon, Intertidal 50% Alaska, US Subtidal successful Zostera noltii Ranwell et al. p None Norfolk & Successful 1974* Suffolk, Great Britain Unspecified Rogers and p None Unspecified Not implicitly Bisterfield, 1975a stated in the citation Phillips 1978a p None Unspecified Not implicitly stated in the citation Carangelo et al. p M Unspecified Not implicitly 1979a stated in the citation Poiner et al. 1989 ? ? Western Gulf of 20% Carpentaria, Northern Territory (NT) Thorogood et al. ? ? Western Gulf of 20% 1990 Carpentaria, NT Calumpong et al. p ? Negros Island, Not implicitly 1996 (abstract) Philippines stated Kenyond ? ? Western Gulf of 20% Carpentaria, NT

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Appendix B

Viable seeds

Fully stained Root tip Distal end of Cotyledon and Hypocotyl unstained cotyledon root tips unstained unstained unstained

Hypocotyl, cotyledon and root tip unstained

Dead seeds

Larger areas Almost Unstained area Cotyledon Completely unstained unstained on cotyledon unstained unstained

Tetrazolium staining of viable Posidonia australis seed tissue, following the method by Lakon (1948). Live tissue stained red with formazon, whilst dead tissue remained colourless. The amount of staining was used to determine seed viability.

168