MACROPHYTE PHENOLOGY IN A TEMPORARILY OPEN/CLOSED ESTUARY COMPARED WITH A PERMANENTLY OPEN ESTUARY

By

Deborah Claire Vromans

Submitted in fulfilment of the requirements for the degree Magister Scientiae

in the

Department of Botany

Nelson Mandela Metropolitan University

Port Elizabeth

2010

Supervisor: J.B. Adams Co-Supervisor: T. Riddin

1 DECLARATION

In accordance with Rule G4.6.3,

4.6.3 A treatise/dissertation/thesis must be accompanied by a written declaration on the part of the candidate to the effect that it is his/her own work and that it has not previously been submitted for assessment to another University or for another qualification. However, material from publications by the candidate may be embodied in a treatise/dissertation/thesis.

I, Deborah Vromans, student number 209090006, hereby declare that the dissertation for MSc (Botany) is my own work and that it has not previously been submitted for assessment to another University or for another qualification.

Signature: ......

Date: ...... 25 January 2011

i

ACKNOWLEDGEMENTS

I am eternally greatful to Professor Janine Adams for putting her faith in me and for supporting my endeavour to undertake a Master of Science degree, without her this would not have been possible. I would like to sincerely thank Ms Taryn Riddin for her quiet kindness and unabiding patience with my seemingly endless throng of queries and her always timely responses. To Professor Eileen Campbell, I extend my appreciation for all the statistical support.

To Professor Alan Whitfield, the National Research Foundation and Nelson Mandela Metropolitan University, I extend my thanks for financing this project.

For my field and laboratory assistants, Mike Verala, Lucienne Human, Sabine Hoppe-Speer, Jacques Dennis and Amanda Khumalo, thank you for all your hard work and assistance at what seemed, at times, an endless task of wading in or filtering and weighing of muddy sediment.

And last but not least to those at home that have supported me through this academic journey, Tom, your astute yet occasional troublesome truths, thank you. To my family, I would like to acknowledge your unwavering support, and to my new furry, four legged friends that always manage to put a smile on my face.

In loving memory of my Layla and Roze.

ii ABSTRACT

Temporarily open/closed estuaries (TOCEs) are unpredictable environments that change in response to mouth condition, which is influenced by freshwater flooding or sea storm surges. The aim of the study was to determine whether macrophyte phenology in a TOCE was event driven rather than cyclically predictable and if it differed from permanently open estuaries (POEs). Macrophyte growth and flowering phenology in response to environmental conditions was investigated in the East Kleinemonde Estuary (TOCE) and the Kowie Estuary (POE) along the eastern coastline of South . The lack of freshwater flooding due to low rainfall coupled with several overwash events resulted in a prolonged period of mouth closure from September 2008 to the end of this study period in 2010. This in turn caused the inundation of the supratidal and intertidal habitats, high water level (> 1.57 m amsl) and high salinity (30 - 42 ppt) in the TOCE. Principle Components Analysis showed that high water level and reduced sediments were the most significant environmental factors affecting macrophyte phenology. Macrophyte phenology in the POE was primarily driven by temperature, sediment redox potential and salinity. The saline high water level and reduced sediment significantly reduced macrophyte cover in all habitats in the TOCE. Macrophytes in the POE maintained high cover abundance due to seasonal re-growth compared to the TOCE where cover declined over the sampling period due to the high water level. Subsequent to water level dropping by as little as 11 - 20 cm in the TOCE, the intertidal species tegetaria and Salicornia meyeriana completed their life-cycles and produced viable seeds within four and three months of germinating respectively. In contrast, the Sarcocornia hybrid and S. meyeriana in the POE took longer to complete their life-cycles, namely seven and nine months respectively, while S. tegataria did not germinate in situ but reproduced vegetatively despite producing seed. In the TOCE, the submerged species cirrhosa and Chara vulgaris completed their life-cycles within five and three months and produced a maximum of 26 242 and 196 998 seeds m-2 respectively. Due to high water level and prolonged inundation, the reproductive periods were shorter for the intertidal and reed and sedge species in the TOCE compared to the POE. Seed output during the two reproduction periods varied between the two estuaries. Sarcocornia decumbens and S. tegetaria produced a substantially higher number of seeds in the TOCE compared to the POE, namely 0 - 102 847 versus 20 661 - 48 576 seeds m-²; and 7 001 - 45 542 versus 1 587 – 16 958 seeds m-² respectively. Seed output in the TOCE was significantly higher in S. tegetaria during the second reproduction period despite the significantly lower cover, which may be a function of the stressful environment in the TOCE. Seed production of S. meyeriana was significantly higher in the POE compared to the TOCE, with 264 224 - 640 292 compared with 24 050 - 27 643 seeds m-², due to higher plant cover in the POE. The research suggests that macrophyte phenology in the TOCE was significantly influenced by mouth condition. Further, macrophytes were able to demonstrate considerable phenotypic plasticity in response to changing and unfavourable environmental conditions. These data can be used in mouth management plans and freshwater requirement studies in TOCEs to ensure that macrophytes can complete their life-cycles and produce viable seeds for the safeguarding of habitat persistence and ecological processes. In impacted estuaries where artificial mouth opening is practised and the macrophytes have been severely degraded or extirpated, management should ensure that the intertidal and supratidal habitats are not inundated during peak flowering and seed production periods i.e. late spring to early autumn (November to March).

iii TABLE OF CONTENTS

DECLARATION……………………………………………………………………………………………………….i ACKNOWLEDGEMENTS……………………………………………………………………………………………ii ABSTRACT ...... iii LIST OF FIGURES……………………………………………………………………………………………………vi LIST OF TABLES……………………………………………………………………………………………………..x 1. CHAPTER 1: INTRODUCTION ...... 1 2. CHAPTER 2: LITERATURE REVIEW ...... 6 2.1 Estuaries……………………………………………………………………………………………………..6 2.1.1 Temporarily open/closed estuaries compared with permanently open estuaries…………...6 2.2 Macrophyte habitats………………………………………………………………………………………..10 2.2.1 ...... 10 2.2.2 Reeds and sedges ...... 13 2.2.3 Submerged macrophytes ...... 15 2.3 Effect of environmental conditions on the phenology of macrophytes…………………………… ….19 2.3.1 Climate: Seasonal changes in temperature, rainfall, irradiance and photoperiod ...... 19 2.3.2 Salinity ...... 22 2.3.3 Water regime ...... 25 2.3.4 Tidal exchange ...... 30 2.3.5 Redox potential ...... 33 2.3.6 Sediment organic matter ...... 37 2.3.7 Sediment moisture content...... 40 2.3.8 pH ...... 44 2.3.9 Light, turbidity and temperature ...... 47 2.4 Seed viability…………………………………………………………………………………………….…51 2.5 Synopsis…………………………………………………………………………………………………….57 3. CHAPTER 3: MATERIALS AND METHODS ...... ……59 3.1 Location of study sites……………………………………………………………………………………..59 3.2 Climate of study sites…………………………………………………………………………………...... 59 3.3 The East Kleinemonde Estuary……………………………………………………………………...... 60 3.3.1 Physical characteristics ...... …60 3.3.2 Macrophytes ...... 61 3.4 The Kowie Estuary…………………………………………………………………………………………62 3.4.1 Physical characteristics ...... 62 3.4.2 Macrophytes ...... 63 3.5 Macrophyte species selected……………………………………………………………………………..63 3.6 Macrophyte growth and reproductive output in the selected species of the East Kleinemonde and Kowie Estuaries…………………………………………………………………………………………….64 3.6.1 Emergent macrophytes ...... 65 3.6.2 Submerged macrophytes ...... 67 3.6.3 Terms used in this study ...... 68 3.7 The time from seed germination to seed formation in three emergent macrophytes and two submerged macrophytes……………………………………………………………………………………………..…68 3.8 The viability of the seeds attached to the macrophytes and if viability changed after maturation…………………...... 69 3.9 Environmental variables and macrophyte phenology……………………………………………….…70 3.9.1 Submerged macrophytes ...... …70 3.9.2 Emergent macrophytes ...... 70 3.10 Statistical analysis…………………………………………………………………………………………72

iv 4. CHAPTER 4: RESULTS ...... 74 4.1 Abiotic conditions in a temporarily open/closed estuary compared with a permanently open estuary ………………………………………………………………………………………………………………..74 4.1.1 Mouth condition of the East Kleinemonde Estuary ...... 74 4.1.2 Supratidal habitat ...... 74 4.1.3 Intertidal habitat ...... 75 4.1.4 Reed and sedge habitat ...... 76 4.1.5 Submerged habitat...... 78 4.2 Macrophyte phenology in a temporarily open/closed estuary compared with a permanently open estuary……………………………………………………………………………………………………...81 4.2.1 kraussii and Juncus acutus ...... 81 4.2.2 Sporobolus virginicus ...... 87 4.2.3 Sarcocornia decumbens ...... 91 4.2.4 Sarcocornia hybrid ...... 95 4.2.5 Salicornia meyeriana ...... 97 4.2.6 Sarcocornia tegetaria ...... 100 4.2.7 Phragmites australis ...... 103 4.2.8 Bolboschoenus maritimus ...... 107 4.2.9 Ruppia cirrhosa and Chara vulgaris ...... 111 4.3 Multivariate analysis: Environmental variables and macrophyte phenolog...………………………119 4.3.1 Juncus kraussii and Juncus acutus ...... 119 4.3.2 Sporobolus virginicus ...... 119 4.3.3 Sarcocornia decumbens ...... 123 4.3.4 Sarcocornia hybrid ...... 125 4.3.5 Salicornia meyeriana ...... 127 4.3.6 Sarcocornia tegetaria ...... 127 4.3.7 Phragmites australis ...... 132 4.3.8 Bolboschoenus maritimus ...... 132 4.3.9 Ruppia cirrhosa and Chara vulgaris ...... 132 5. CHAPTER 5: DISCUSSION ...... 139 5.1 Abiotic conditions in a temporarily open/closed estuary compared with a permanently open estuary ………………………………………………………………………………………………………………139 5.2 Macrophyte phenology in a temporarily open/closed estuary compared with a permanently open estuary………………………………………………………………………………………………….....142 5.2.1 Juncus kraussii and Juncus acutus ...... 142 5.2.2 Sporobolus virginicus ...... 143 5.2.3 Sarcocornia and Salicornia species ...... 144 5.2.4 Phragmites australis ...... 148 5.2.5 Bolboschoenus maritimus ...... 149 5.2.6 Ruppia cirrhosa and Chara vulgaris ...... 150 6. CHAPTER 6: CONCLUSIONS AND RECOMMENDATIONS ...... 154 7. REFERENCES ...... 159 8. APPENDIX ...... 203 8.1 Abiotic conditions in the East Kleinemonde and Kowie estuaries………………………………... ..203 8.1.1 Supratidal habitat ...... 203 8.1.2 Intertidal habitat ...... 203 8.1.3 Reed and sedge habitat ...... 204 8.2 Macrophyte phenology of the selected species in the East Kleinemonde and Kowie estuaries ..204 8.2.1 Juncus kraussii and Juncus acutus ...... 204 8.2.2 Sporobolus virginicus ...... 206 8.2.3 Sarcocornia decumbens ...... 208 8.2.4 Sarcocornia hybrid ...... 212 8.2.5 Salicornia meyeriana ...... 213 8.2.6 Sarcocornia tegetaria ...... 216 8.2.7 Phragmites australis ...... 218 8.2.8 Bolboschoenus maritimus ...... 220 8.2.9 Ruppia cirrhosa and Chara vulgaris ...... 223

v LIST OF FIGURES

Figure 2.2: Macrphyte species selected for the study...... 18

Figure 3.1: Map of southern Africa showing the three major biogeographic regions (Whitfield, 2000) and the East Kleinemonde and Kowie estuaries along the south-eastern Cape coastline of (Cowley et al., 2001)...... 59

Figure 3.2: Monthly rainfall (mm) over the study period from February 2009 to June 2010 in the East Kleinemonde and Kowie estuaries...... 60

Figure 3.3: Sampling points (white circles) and transects (numbered white lines) in the East Kleinemonde Estuary...... 64

Figure 3.4: Sampling site in the Kowie Estuary...... 65

Figure 3.5: (a) S. decumbens that were 100 % seeding indicated by the corky ; and (b) S. meyeriana inflorescences that were 100 % flowering...... 66

Figure 4.1.1: Average water level and salinity in the East Kleinemonde Estuary (EK), including water salinity in the Kowie Estuary (KW). Arrows indicate the overwash events in the East Kleinemonde Estuary...... 74

Figures 4.1.2: (a) Transect one (west bank) in the East Kleinemonde Estuary with J. kraussii in the background, S. meyeriana, S. tegetaria and S. virginicus in the foreground. (b) Transect two (east bank) with P. australis and B. maritimus habitat in the East Kleinemonde Estuary. Both photographs were taken at the start of the sampling period (February 2009)...... 77

Figures 4.1.3: (a) Transect one in the Kowie Estuary with S. tegetaria and patches of S. meyeriana and S. decumbens. (b) Transect two with P. australis and J. acutus habitat in the Kowie Estuary. Both photographs were taken at the start of the sampling period (February 2009)...... 78

Figure 4.2.1: Structural components of the J. kraussii paniculate . The holds the entire inflorescence and the rachis is the main stem holding the flowers or branches of the inflorescence...... 82

Figure 4.2.2: Mean monthly percentage cover per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) of J. acutus (Ja) in the Kowie Estuary February 2009 to June 2010 (± SE)...... 84

Figure 4.2.3: Mean monthly height (cm) per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE)...... 84

Figure 4.2.4: Mean monthly number of inflorescences per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE)...... 85

Figure 4.2.5: Mean monthly percentage for the phase of the inflorescences per m² during (a) the first flowering period; and (b) the second flowering period for J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE)...... 85

Figure 4.2.6: Mean monthly (a) number of seed per m²; and percentage germination and number of days to germination for (b) J. acutus (Ja) seeds harvested from the Kowie Estuary from February 2009 to June 2010 and (c) J. kraussii seeds harvested from the East Kleinemonde Estuary (± SE)...... 86

vi Figure 4.2.7: A paniculate inflorescence of (a) S. virginicus changing into an infructescence from which a mean of 52 caryopsis (fruit) will develop and produce one seed each (b)...... 88

Figure 4.2.8: Mean monthly percentage cover per m² of S. virginicus in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 89

Figure 4.2.9: Mean monthly height (cm) per m² of S. virginicus in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 89

Figure 4.2.10: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seeds per m² of S. virginicus in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 90

Figure 4.2.11: Mean monthly percentage germination and number of days to germination of S. virginicus seeds harvested from the East Kleinemonde Estuary, for the period April 2009 to June 2009, and the Kowie Estuary, for the period January 2010 to March 2010 (± SE)...... 90

Figure 4.2.12: S. decumbens had an average of nine seeds (a) per fertile segment (b) and an average of 18 fertile segments per inflorescence...... 92

Figure 4.2.13: Mean monthly percentage cover per m² of S. decumbens in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 93

Figure 4.2.14: Mean monthly percentage cover per m² of S. decumbens in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 93

Figure 4.2.15: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of S. decumbens in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February/May 2009 to June 2010 (± SE)...... 94

Figure 4.2.16: Mean monthly percentage germination and number of days to germination of S. decumbens seeds harvested from the East Kleinemonde (EK) Estuary (for the period April 2009 to October 2009) and the Kowie (KW) Estuary (for the period April 2010 to June 2010) (± SE)...... 94

Figure 4.2.17: Mean monthly (a) percentage cover; and (b) height (cm) per m² of the Sarcocornia hybrid in the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 96

Figure 4.2.18: Average monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of the Sarcocornia hybrid in the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 96

Figure 4.2.19: Mean monthly percentage cover per m² of S. meyeriana in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 98

Figure 4.2.20: Mean monthly height (cm) per m² of S. meyeriana in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 98

Figure 4.2.21: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of S. meyeriana in the East Kleinemonde Estuary Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 99

vii Figure 4.2.22: Mean monthly percentage germination and number of days to germination of S. meyeriana seeds harvested from the East Kleinemonde (EK) Estuary (± SE)...... 99

Figure 4.2.23: Mean monthly percentage cover per m² of S. tegetaria in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 101

Figure 4.2.24: Mean monthly height (cm) per m² of S. tegetaria in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 101

Figure 4.2.25: Mean monthly (a) number of inflorescences, (b) phase of inflorescences and (c) number of seed per m² of S. tegetaria in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 102

Figure 4.2.26: Mean monthly percentage germination and number of days to germination of S. tegetaria seeds harvested from the East Kleinemonde (EK) Estuary (± SE)...... 102

Figure 4.2.27: P. australis panicle (inflorescence) with (a) an average of 27 rachis per peduncle, 59 spikelets per rachis and 2 flowers/fruits per spikelet; and (b) a caryopsis (fruit)...... 104

Figure 4.2.28: Mean monthly percentage cover per m² of P. australis in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 105

Figure 4.2.29: Mean monthly plant height (cm) per m² of P. australis in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 105

Figure 4.2.30: Mean monthly (a) number of inflorescences, (b) phase of inflorescences, and (c) number of seeds per m² of P. australis in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 106

Figure 4.2.31 (a) The flowering spikelets of B. maritimus and (b) seedlings germinating from an achene (fruit), which contains one seed...... 108

Figure 4.2.32: Mean monthly percentage cover per m² of B. maritimus in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 109

Figure 4.2.33: Mean monthly height (cm) per m² of B. maritimus in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 109

Figure 4.2.34: Mean monthly (a) number of inflorescences, (b) phase of inflorescences and (c) number of seed per m² of B. maritimus in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE)...... 110

Figure 4.2.35: Mean monthly percentage germination and number of days to germination of B. maritimus seeds harvested from the (a) East Kleinemonde Estuary Estuary (EK); and (b) the Kowie Estuary (KW) during 2010 (± SE)...... 110

Figure 4.2.36: Mean monthly biomass (g DW) per m² of R. cirrhosa; and (b) C. vulgaris in the East Kleinemonde Estuary (EK) from February 2009 to June 2010 (± SE)...... 112

Figure 4.2.37: Mean monthly height (cm) per m² of R. cirrhosa; and (b) C. vulgaris in the East Kleinemonde Estuary (EK) from February 2009 to June 2010 (± SE)...... 112

viii Figure 4.2.39: Growth periods and periods of decline for the selected species in the East Kleinemonde Estuary based on the mean monthly percentage cover m-² to determine maximum growing periods/season for open mouth recommendations...... 114

Figure 4.2.40: Growth periods and periods of decline for the selected species in the Kowie Estuary based on the mean monthly percentage cover m-² to determine maximum growing periods/season for open mouth recommendations...... 114

Figure 4.2.41: Phenogram of the reproductive cycles (including seed release periods) of the selected species in the East Kleinemonde Estuary (EK) and Kowie Estuary (KW)...... 116

Figure 4.3.1: Ordination diagram based on a PCA of species and environmental data for (a) J. kraussii and (b) J. acutus...... 120

Figure 4.3.2: Ordination diagram based on a PCA of species- and environmental data for S. virginicus in in the (a) East Kleinemonde Estuary; and (b) the Kowie Estuary...... 122

Figure 4.3.3: Ordination diagram based on a PCA of species - and environmental data for S. decumbens in the (a) East Kleinemonde Estuary; and (b) the Kowie Estuary...... 124

Figure 4.3.4: Ordination diagram based on a PCA of species and environmental data for the Sarcocornia hybrid in the the Kowie Estuary...... 126

Figure 4.3.5: Ordination diagram based on a PCA of species- and environmental data for (a) S. meyeriana in the East Kleinemonde Estuary; and (b) S. meyeriana in the Kowie Estuary...... 128

Figure 4.3.6: Ordination diagram based on a PCA of species and environmental data for S. tegetaria (a) first generation plants; and (b) second generation plants in the East Kleinemonde Estuary; and (c) in the Kowie Estuary...... 130

Figure 4.3.7: Ordination diagram based on a PCA of species and environmental data for P. australis in the (a) East Kleinemonde Estuary and (b) Kowie Estuary...... 133

Figure 4.3.8: Ordination diagram based on a PCA of species and environmental data for B. maritimus in the (a) East Kleinemonde Estuary and (b) Kowie Estuary...... 135

Figure 4.3.9: Ordination diagram based on a PCA of species and environmental data for (a) R. cirrhosa and (b) C. vulgaris in the East Kleinemonde Estuary...... 137

Figure 5.1: Plant cover of J. kraussii in the East Kleinemonde Estuary in April 2009 (left) compared to January 2010 (right), when significant die-back had occurred due to prolonged inundation. Once the water receded in January 2010, re-growth was rapid probably due to the significant die-back that occurred during inundation and the warmer summer temperatures...... 143

Figure 5.2: In the Kowie Estuary dead bushes of first generation plants (arrow) remained in situ for 11 months and held seeds for six months in the inflorescences, while second generation plants expanded around and under them...... 147

ix LIST OF TABLES

Table 1: Characteristics of the temporarily open/closed estuaries compared with permanently open estuaries in the warm temperate region of South Africa ...... 9

Table 4.1.1: Characteristics of the habitats of the emergent species in the East Kleinemonde Estuary compared with the Kowie Estuary...... 79

Table 4.1.2: Characteristics of the habitats of the emergent species in the East Kleinemonde Estuary ...... 80

Table 4.1.3: Characteristics of the habitat of the submerged species habitat in the East Kleinemonde Estuary...... 81

Table 4.2.1: Extrapolation results for the number of flowers, fruit and seed per inflorescence for a J. kraussii plant in the East Kleinemonde Estuary...... 82

Table 4.2.2: Mean monthly change in percentage cover and height of the selected macrophytes within 1 m2 quadrats in the East Kleinemonde Estuary and the Kowie Estuary …………………………………………………………………………………………………………………...... 115

Table 4.2.3: Mean monthly rate of new reproductive output (percentage), including the mean number of inflorescences and seed produced m-2 (±SE) per flowering period in the selected species of the East Kleinemonde and Kowie estuaries…...... 117

Table 4.2.4: Flowering characteristics of the selected species in the East Kleinemonde and Kowie Estuaries. N/a = Not applicable……...... 118

Table 4.3.1: Summary of PCA of species- and environmental data for J. kraussii in the East Kleinemonde Estuary...... 121

Table 4.3.2. Summary of PCA of species and environmental data for J. acutus in the Kowie Estuary...... 121

Table 4.3.3: Summary of PCA of species and environmental data for S. virginicus in the East Kleinemonde Estuary...... 123

Table 4.3.4: Summary of PCA of species and environmental data for S. viriginicus in the Kowie Estuary...... 123

Table 4.3.5: Summary of PCA of species and environmental data for S. decumbens in the East Kleinemonde Estuary...... 125

Table 4.3.6: Summary of PCA of species and environmental data for S. decumbens plants in the Kowie Estuary...... 125

Table 4.3.7: Summary of PCA of species- and environmental data for Sarcocornia hybrid plants in the Kowie Estuary...... 126

Table 4.3.8: Summary of PCA of species and environmental data for S. meyeriana in the East Kleinemonde Estuary...... 129

Table 4.3.9: Summary of PCA of species and environmental data for S. meyeriana in the Kowie Estuary...... 129

Table 4.3.10: Summary of PCA of species and environmental data for first generation S. tegetaria plants in the East Kleinemonde Estuary...... 131

x Table 4.3.11: Summary of PCA of species and environmental data for S. tegetaria in the Kowie Estuary...... 131

Table 4.3.12: Summary of PCA of species and environmental data for P. australis in the East Kleinemonde Estuary...... 134

Table 4.3.13: Summary of PCA of species and environmental data for P. australis in the Kowie Estuary...... 134

Table 4.3.14: Summary of PCA of species and environmental data for B. maritimus in the East Kleinemonde Estuary...... 136

Table 4.3.15: Summary of PCA of species and environmental data for B. maritimus in the Kowie Estuary. ....136

Table 4.3.16: Summary of PCA of species and environmental data for R. cirrhosa in the East Kleinemonde Estuary...... 138

Table 4.3.17: Summary of PCA of species and environmental data for C. vulgaris in the East Kleinemonde Estuary...... 138

xi

1. CHAPTER 1: INTRODUCTION

Macrophyte community structure and composition in temporarily open/closed estuaries (TOCEs) is influenced by freshwater inflow, tidal exchange, salinity, water level fluctuations and sediment dynamics (van Niekerk et al., 2008). The opening of the East Kleinemonde Estuary is unpredictable due to erratic rainfall in the region, unlike numerous small TOCEs in KwaZulu-Natal whose mouths open after seasonally high rainfall periods (Bennett, 1989; Harrison and Whitfield, 1995). Consequently, macrophyte responses in these estuaries are expected to be event driven. Due to the stressful nature of the TOCE environment, salt marsh and submerged macrophytes have low species richness in the East Kleinemonde Estuary. This means that any loss in macrophytes attributable to varying physico-chemical conditions represents a loss of biodiversity (van Niekerk et al., 2008). In comparison, permanently open estuaries (POEs) have regular seawater and freshwater inflows and are permanently connected to the sea (Adams and Riddin, 2007; Whitfield and Bate, 2007). Due to the large and regular tidal exchange in these systems, the salt marshes are better developed compared to TOCEs (Adams and Riddin, 2007). The cover of emergent macrophytes is frequently higher in POEs due to the availability of suitable habitat (Adams and Riddin, 2007) guaranteed by the more stable environment, as reflected in the regular seawater and freshwater inflows.

Estuaries, particularly temporarily open/closed estuaries (TOCEs), are highly unpredictable environments that experience rapid changes in environmental conditions in response to mouth condition. These changes have a significant affect on species distributions, although expansions and contractions are considered by some authors to be readily reversible, depending on the environmental conditions (Zedler et al., 1986). Macrophytes are known to develop plastic responses to variations in the abiotic environment, which is the expression of phenotypic plasticity (Brock, 1991; Davy, 2001; Silva et al., 2006). Flexibility of life-cycle pattern and plasticity of growth form enhanced survival in widely fluctuating environments (Brock, 1991).

Macrophytes in temperate TOCEs in South Africa, such as the East Kleinemonde Estuary, are exposed to a wide range of environmental fluctuations, particularly with respect to water level. Water level fluctuations result in wetting and drying of the supratidal and intertidal habitat, which is dependent on rainfall and/or overwash events. Both sexual and asexual reproductive strategies are important for the persistence of macrophytes in such dynamic environments. Flooding events may result in partial or complete biomass loss or die-back of submerged species in response to a drop in water level (Riddin and Adams, 2008a). Despite the fact that vegetative production is the predominant reproductive strategy in salt marsh plants and submerged perennial macrophytes (Verhoeven, 1979), seed banks play a critical role in their re-establishment when water level fluctuations result in local extirpation of species (Keddy and Reznicek, 1982; Welling et al., 1988; Adams and Riddin, 2007). It has been suggested that as disturbance increases, seed production and growth from seed reserves are more important (Verhoeven, 1979; Kautsky, 1990; Casanova and Brock, 1996; Combroux and Bornette, 2004; Davis and Stevenson, 2007). Seed banks of both emergent and submerged macrophytes offer resilience to these varying conditions as they can re-colonize their respective habitats when the waters retreat after protracted flooding or advance due to drought, limited rainfall and/or overwashing (Brock, 1991; Adams

1 and Riddin, 2007). However, macrophytes need to complete their life-cycles in order to replenish these seed stocks otherwise their persistence may be compromised.

Life-cycle and reproductive strategies are closely coupled with water level fluctuations and the associated fluctuation in environmental factors, such as salinity and temperature. Flowering in Salicornieae species, for example, generally occurs over a six month period, however during this time conditions may not be conducive causing the duration of the flowering period to be significantly curtailed. This may be the case where water level is high and the plants are inundated and unable to flower. Only once the water recedes can flowering take place, while favourable flowering conditions may occur out of the normal flowering season or not present themselves for several years (O‘Callaghan, 1992). When water level in the estuary drops, submerged species such as Ruppia cirrhosa are exposed and die but as soon as water level rises to suitable levels the plants are able to germinate rapidly to complete their life-cycle within three months (Adams and Bate, 1994c; Riddin and Adams, 2010). Similarly, in temporarily flooded areas, Gesti (2005) found that R. cirrhosa developed an annual cycle and completed its life-cycle in less than two months while producing a high abundance of flowers and seeds. In contrast, in permanently inundated waters and under favourable conditions R. cirrhosa completes its life-cycle over a longer period. Low temperature during winter can result in a quiescence period when the plant hibernates in a vegetative state. Such flexibility in life-cycle and reproductive strategies is essential to ensure the persistence of these macrophyte communities, particularly because mouth breaching events in TOCE are not always seasonal (Brock, 1991; Casanova, 1994).

Research on the East Kleinemonde Estuary by Riddin and Adams (2008; 2010) showed that water level fluctuations are a key driving force affecting the spatial and temporal distribution of macrophytes. Supratidal salt marsh dies if inundated for one to two months while intertidal salt marsh dies if inundated for three months. This means that seed reserves need to be replenished in order for the habitat to persist after a disturbance event, such as flooding which removes seeds. Submerged species have persistent seed banks (Riddin and Adams, 2009) and Ungar (1987a/b) notes the persistence of seed banks in coastal salt marshes while Salicornia europaea seed banks have been reported as persistent under more severe and less predictable conditions (Ungar, 1987b; Davy, 2001). In the East Kleinemonde Estuary, Chara vulgaris and Ruppia cirrhosa had low germination percentages (11 and 15 %), but due to their large seed banks that remain viable, yet dormant, for many years they were able to persist for an indefinite period despite the variable and dynamic nature of the TOCE environment (Riddin and Adams, 2009). Establishment of seedlings from a seed bank and the resultant pattern of germination are affected by dormancy, until favourable conditions return (Silvertown, 1988). Dormancy guarantees that seasonal germination is established if envioronmental conditions are favourable. The pattern of germination is often an adaptation to the predictability of variation within a habitat. In a less predictable habitat, such as a TOCE, the risk of germinating early could negate any competitive advantage (Silvertown, 1988; Casanova and Brock, 1991).

In contrast, this study hypothesises that macrophytes in permanently open estuaries (POE) have cyclical life- cycles and reproduction periods coupled with seasonal patterns (depending on their growth characteristics) due to regular tidal exchange and freshwater input. Salicornia meyeriana, an annual intertidal salt marsh plant for example, has the greatest growth from February to May (O‘Callaghan, 1992), while the vegetative growth of S.

2 tegetaria is greatest in spring (Adams and Bate, 1994b). It is therefore the suggestion that, due to the comparatively regular nature of the POE environment, these cyclical patterns of growth and reproduction tend to continue unabated. Once the salt marsh is established, succession is largely driven by biotic factors (Ungar, 1978; Gray, 1985; Bertness and Ellison, 1987; Shumway and Bertness, 1992) and changes in the system are not as dramatic as opposed to those in TOCEs. POEs usually have large catchments and relatively high runoff throughout the year and maintain a permanent connection to the sea through strong tidal forces (Whitfield and Bate, 2007). This study therefore tests whether the POE environment is more predictable and stable allowing macrophytes to complete their life-cycles and replenish their seed banks within a seasonal or cyclical annual pattern.

An understanding of macrophyte responses to changes in environmental factors is needed in order to predict the response of the macrophytes and other biota to changes in mouth condition. The duration of the open or closed mouth condition and salinity levels, influence the abundance of macrophytes and related fauna in these estuaries (Whitefield and Bate, 2007; Chuwen, 2009). Little information is available on the phenology of macrophytes in estuaries. Consequently, this research will provide input to mouth management plans and environmental flow requirement studies. An adequate freshwater inflow regime will ensure that the system continues to function ecologically, as the macrophytes will be able to complete their life-cycles and replenish their seed banks. Furthermore, water requirement data are necessary in terms of the South African National Water Act (36 of 1998). The Act requires that all estuaries are supplied with adequate water to maintain a recommended ecological state. The ecological state is the condition that reflects the state in which the estuary must be managed to ensure the survival of the system‘s flora and fauna. The East Kleinemonde Estuary is a relatively pristine estuary and results from this study can therefore be applied in the management of other TOCEs. The study can provide input on the best time for artificially opening the mouth of impacted estuaries.

The main objectives of the research were as follows:

1. Determine the rate of species growth and reproductive output (i.e. number of seeds produced m-² or gram of seeds m-²) in the following macrophyte species of the East Kleinemonde and Kowie estuaries:

 Supratidal salt marsh: Juncus kraussii Hochst, Juncus acutus L., Sporobolus virginicus (L.) Kunth.

 Intertidal salt marsh: Upper intertidal - Sarcocornia decumbens (Tölken) A.J. Scott.

Lower intertidal - Sarcocornia tegetaria S. Steffen, Mucina & G. Kadereit, Salicornia meyeriana Moss.

 Reeds and sedges: Phragmites australis (Cav.) Trin ex Steud, Bolboschoenus maritimus (L.) Palla.

 Submerged macrophytes: Ruppia cirrhosa (Petagna) Grande and Chara vulgaris L.

The above selected species are grouped per habitat, as determined in POEs by Adams et al. (1999), namely: salt marsh, reeds and sedges and submerged macrophytes. Salt marsh plants show distinct zonation patterns along tidal inundation and salinity gradients (Adams et al., 1999) and are therefore further subdivided into supratidal and intertidal salt marsh. However, TOCEs that are predominantly closed and

3 therefore not tidal, such as the East Kleinemonde Estuary, do not generally have these zones. According to Day (1981) and O‘ Callaghan (1994), S. decumbens is located in the area between the extreme high water spring tide level to the mean high water spring level, known as the upper intertidal zone. It should be noted that J. kraussii can either occur at the mean high water spring (MHWS) or in shallow parts of estuaries in the intertidal zone, in association with S. virginicus. In the East Kleinemonde Estuary, J. kraussii tends to be further up the elevation gradient (>1.6 m amsl) in the ‗supratidal zone‘, with S. virginicus growing across the ‗upper intertidal‘ to ‗supratidal zones‘. Although C. vulgaris is a submerged macroalgae, it functions like and occurs in the same habitat as R. cirrhosa; and has therefore been categorized accordingly (Riddin and Adams, 2008a).

It should be noted that Sarcocornia tegetaria has been relatively recently named and distinguished from Sarcocornia perennis (Steffen et al., 2007). Its distribution is restricted to southern Africa. Consequently, all references to Sarcocornia perennis Miller A.J. Scott refers to this species, which is found elsewhere in the world.

2. Compare the rate of reproductive output (number of seeds produced m-² or grams of seeds m-²) in order to establish if reproduction in the TOCE is more rapid compared to the POE. This will influence the recommended duration of an open estuary mouth in the TOCE to ensure that the macrophytes produce viable seeds.

3. Determine the time from seed germination to seed formation of three emergent species, namely S. decumbens, S. tegetaria and S. meyeriana in both estuaries; and of the two submerged species in the TOCE, namely R. cirrhosa and C. vulgaris. The three emergent species were selected because they readily germinate from seed compared to the other emergent species which reproduce vegetatively via . This makes it difficult to determine new seedling growth from seeds without destroying the plants. The submerged species were not found in the POE during the sampling period.

4. Determine the viability of mature seeds harvested from the plants and if viability changes after maturation in the TOCE. The importance of testing for seed viability was to determine if the seeds required an extended period of time to enable germination in a TOCE, i.e. an after ripening or over-wintering period, after maturation. This will influence the recommended period or timing of an open estuary mouth in the TOCE to ensure that the seeds are able to germinate.

5. Determine whether life-cycles are cyclical or related to the environmental conditions in the TOCE and POE by measuring the in situ physico-chemical conditions at each stage of the species life-cycles to identify the most significant factors influencing macrophyte growth and reproductive output. The following environmental variables were sampled:

 Salinity and electrical conductivity (water and sediment)

4  pH (water and sediment)

 Redox potential (water and sediment)

 Sediment moisture content

 Sediment organic content

 Water level

 Temperature (water)

 Turbidity (water)

Some of the specific hypotheses that the research addressed were:

1. The life-cycles of macrophytes in TOCEs are event driven i.e. they are driven by the estuary mouth being open or closed due to the associated water level fluctuations and environmental conditions.

2. Environmental conditions in a TOCE are more variable and/or stressful than environmental conditions in a POE, which are more stable.

3. Macrophytes respond to relatively small water level fluctuations (10 - 20 cm) without the mouth breaching in a TOCE.

4. Macrophytes complete their life-cycles more rapidly in a TOCE compared to a POE.

5. Macrophytes are highly plastic i.e. adaptable and flexible in response to fluctuating and stressful environmental conditions in a TOCE.

6. Intertidal salt marsh requires at least two months for plants to produce viable seeds after flowering and at least four months after germination in a TOCE.

7. Submerged macrophytes require stable water levels for at least three to four months for viable seed to develop after germination in a TOCE.

5 2. CHAPTER 2: LITERATURE REVIEW

2.1 ESTUARIES

Estuaries in South Africa have been classified into five key types, namely permanently open estuaries, intermittently open estuaries (or temporarily/open closed), estuarine bays, estuarine lakes and river mouths (Whitfield, 1998). There are approximately 250 functional estuaries along the coastline of South Africa that cover roughly 70 000 ha (Whitfield, 2000). They are divided into different bioregions based on rainfall, river flow and water temperatures of the region (Day, 1981; Harrison, 2004 Sheppard, 2009), and include cool temperate, warm temperate and subtropical estuaries (Whitfield, 1998). The majority of estuaries occur in the Eastern Cape, according to the bio-geographical classification of the South African coast. The Eastern Cape Province supports 123 warm temperate estuaries compared to 117 subtropical estuaries to the northeast and 10 cool temperate estuaries in the west (Whitfield, 1998; Strydom, 2002). The province is predominantly comprised of temporarily open/closed and permanently open estuaries, although river mouths are also present.

2.1.1 Temporarily open/closed estuaries compared with permanently open estuaries Climate, topography and catchment geology affects the type of estuaries along the South African coastline, but all estuaries are either permanently open or temporarily open/closed to the sea depending on their hydrodynamic characteristics (Whitfield and Bate, 2007). The South African coastline is comprised of approximately 175 temporarily open/closed estuaries (TOCEs) out of 250 estuaries (Whitfield, 1992). These estuaries are usually small and shallow (<2 m); and are either predominantly open or closed (more or less than 50% of the time) to the sea. When closed they are isolated from the sea by a sand bar, which is usually the result of low river inflow coupled with high wave action from the sea. During periods of mouth closure, water level can increase to the extent that the adjacent floodplain becomes flooded (Begg, 1984). Due to their irregular connection to the sea the physico-chemical condition of these systems can be unpredictable and highly variable, particularly water level and salinity (Riddin and Adams, 2008a). The mouth condition is primarily regulated by river flow regime, marine sediment supply, tidal prism and the local wave action (Reddering, 1988; Cooper, 2001). The volume of river inflow is the major regulator of these systems, the magnitude of which is governed principally by rainfall and catchment size (Riddin and Adams, 2008a). Reduced river inflow can lead to prolonged mouth closure and a shorter open phase (Whitfield, 2005). In the warm temperate region, where rainfall is low, breaching generally only occurs during periods of high fluvial discharge (Perissinotto et al., 2000; Cowley and Whitfield, 2001). Rainfall along the south-eastern coastal area of the Eastern Cape is highly variable and not as strongly seasonal (Kopke, 1988 and Cowling et al., 2001). Mouth opening events therefore occur infrequently throughout the year (Cowley, 1998), or not at all. Water temperatures of TOCEs are distinctly seasonal with summer highs of 20 - 27°C and winter lows of 13 - 20°C, which is a function of seasonal trends rather than due to linkages to the sea. Salinity does not show this seasonal trend but can change dramatically due to mouth condition (Snow and Taljaard, 2007). Most of the TOCEs in the warm-temperate region have a salinity range between 15 - 30 ppt, are well oxygenated and have low turbidity levels (<5 NTU) due to predominantly clear waters (Vorwerk et al., 2001; Harrison, 2004). The physical environment of small TOCEs is known to experience large fluctuations and estuarine habitats are primarily influenced by physical rather than biological factors (Riddin and Adams, 2008a).

6

Approximately 25 % of the 250 South African estuaries are permanently open estuaries (POEs) meaning that they are permanently connected to the sea. This is due to climate, rainfall patterns and coastal morphology (Whitfield, 1992; Whitfield, 1998). POEs are large systems (>150 ha) that are able to maintain an open mouth due to a strong tidal flow even when run-off decreases in the low flow season (Whitfield, 1992). They are therefore less sensitive to flow reduction because run off and/or tidal flow are sufficient to maintain an open mouth despite the occurrence of droughts. Warm-temperate POEs, such as the Kowie Estuary, maintain their connection to the sea primarily through tidal currents (Harrison, 2004). These estuaries are permanently tidal and have a well defined pycnocline, which is created by the dense tidally regulated ocean water and the fresh and lighter river inflow. Both these forces represent the mixing processes in the estuary, accelerated by high tidal ranges and turbulence from tidal currents (Taylor et al., 2006). Freshwater input into these estuaries is often very high in comparison to the TOCEs. Salinity is influenced primarily by the mixing of ocean water and freshwater inflow (Harrison, 2004). The reduction in seasonal base flow affects the salinity distribution in POEs because it modifies the extent of salinity penetration through the mouth. Salinity usually remains fairly stable however because these tide dominated estuaries are generally well flushed due to a large tidal prism, which causes the salinity to be low at the head of the estuary compared to the rest of the estuary, which is similar to that of the open ocean (Taylor et al., 2006). Most of the warm-temperate open estuaries have a salinity range of 25 - 35 ppt, are well oxygenated and have turbidities below 10 NTU (Harrison, 2004). Water temperatures of POEs are a function of their linkage to the sea and have maximum summer temperatures of between 18 and 22 °C with minimum winter temperatures of between 12 and 14°C (Harrison, 2004). POEs have a continuous supply of nutrients as opposed to TOCEs where this supply is decreased or depleted during low flows or drought years respectively (Whitfield and Bate, 2007). Refer to Table 1 for a comparison of TOCEs compared with POEs in the warm temperate region of South Africa.

POEs do experience physico-chemical fluctuations but these are not anticipated to be as spatio-temporally dynamic when compared to TOCEs. Physico-chemical fluctuations usually occur as a result of floods or droughts, which are generally short in duration e.g. oligohaline conditions (<5 %) occur in the upper reaches and frequently extend into the middle and lower reaches for brief periods during major river floods (Whitfield, 2005). Freshwater floods directly influence the estuary‘s water temperature, salinity, pH, turbidity, nutrient status, organic inputs and dissolved oxygen concentrations; and indirectly affect mouth status, tidal prism, habitat diversity, primary and secondary productivity, fish recruitment, food availability and competition (Whitfield, 2005). For example, in the Swartkops Estuary, a POE in South Africa approximately 150 km south of the Kowie Estuary, flooding introduced freshwater throughout the estuary (Baird et al., 1988). Salinity levels before the flood event were 33 – 35 ppt, during the flood salinity dropped to 0 ppt and just after the flood increased to 10 – 15 ppt. Pre-flood salinity levels returned after two months at the mouth (35 ppt), while in the middle reaches salinity remained lower (20 ppt vs 33 ppt). Other physico-chemical changes occurred subsequent to the flood, for example light attenuation, nutrients and particulate suspended material increased (Baird et al., 1988).

A study by Chuwen et al. (2009) of two seasonally open estuaries, one POE and four normally closed estuaries in southern investigated the physico-chemical parameters of these estuaries. The salinity variations in

7 the POE compared to the seasonally open estuaries tended to be smaller. For example, during short opening phases the salinity of Broke Inlet did not exceed 10 ppt but during extended periods of mouth opening reached that of seawater, showing discernible cyclical changes during the year. In comparison, the POE at Oyster Harbour always had salinity values close to that of seawater. The normally closed estuaries also showed marked salinity fluctuations due to changes in seasonal conditions, namely: droughts caused the salinity to increase from 53 to 112 ppt (July 1996 to March 1997), which was then followed by substantial riverine discharge causing the salinity to decline to 14 ppt (September 1997). The river inflow resulted in a breach event which in turn caused the salinity to rise again (Young and Potter, 2002).

The relative stability of physico-chemical factors in POEs was also demonstrated in the Kariega and Great Fish estuaries, which are located along the eastern coast of South Africa. Jennings (2006) recorded salinity during June, September, December and March (winter, spring, summer and autumn) over a one year period. In the Kariega Estuary, salinity ranged from 34.2 - 36.8 ppt, because it is marine-dominated and therefore salinity values usually remain close to that of the sea (Teske and Wooldridge, 2003) along the entire length of the estuary. The salinity of the Great Fish Estuary ranged from 14.4 - 20 ppt with the March sampling trip recording 4.7 ppt due to high rainfall and because the estuary is a freshwater or river dominated system. The Kromme Estuary, another marine dominated system, also reflects high salinity values along the entire length of the estuary and a lack of salinity variation (Teske and Wooldridge, 2003) due to its permanent connection to the sea. Other river-dominated POEs, such as the Keiskamma and Sundays estuaries, experience shifts in the salinity regime downstream during the rainy season (Teske and Wooldridge, 2003), but these are predictable due to their seasonal nature. Studies in the Kromme Estuary showed relatively constant average salinity levels, namely 28.7 ppt from 1993 to 1994, 23.7 ppt to 34.2 ppt from November 1998 to December 1998, 35.4 ppt in January 1999, 35 ppt in November 2003 and 36.3 ppt in July 2004. Average salinity therefore ranged from 23.7 - 36.3 ppt (Snow and Adams, 2005). In POEs, the water column mixing process is tidally and riverine driven, with mean salinity usually fluctuating between 15 and 35 ppt. In contrast, TOCEs usually experience fluctuations in salinity from 1 - 35 ppt and greater, which demonstrate the substantial fluctutations that may occur in this environment relative to the POEs (Whitfield, 1992; Whitfield, 2005).

As noted above, the physico-chemical conditions of TOCEs can vary dramatically, particularly water level and salinity (Riddin and Adams, 2008a) over short time periods due to their dynamic and variable connection with the sea (Pollard, 1994; Hastie and Smith, 2006). When TOCEs are closed to the sea they either become brackish due to reduced freshwater input, or hypersaline due to water evaporation and macrophytes rapidly decline (PERL 1990 and Zedler et al., unknown, Chuwen et al., 2009), while prolonged submergence can cause salt marsh plants to die-back (Adams and Bate, 1994a). Most biological impacts of mouth closure are thought to be a result of changes in the physico-chemical environment, but the timing of mouth closure also interrupts recruitment and reproductive processes (Griffiths, 2001; Young and Potter, 2002). When water levels eventually subside, the recruitment of new plants can also be prevented due to the absence of reproduction. It might be expected that, because TOCEs usually have small catchments and highly variable physical and chemical conditions, their ecology would differ compared to POEs (Hastie and Smith, 2006). In order for small TOCEs to maintain their structure and function they need a proportionally greater percentage of their mean annual run-off in comparison to POEs (van Niekerk et al., 2008b).

8

Physico-chemical changes in a TOCE during the closed compared to open mouth condition is demonstrated in the following example (van Niekerk et al., 2008b). During the intermittently open/closed (high flow) condition, the East Kleinemonde Estuary experiences salinity levels ranging from 37 ppt, in the lower to middle reaches, to 1 ppt in the upper reaches. In comparison, during the closed mouth condition the estuary is usually saline (>25 ppt) (van Niekerk et al., 2008b). These fluctuations in physico-chemical conditions are expected to have an impact on macrophyte phenology or life-cycle events. Limited data on the fluctuations in physico-chemical data in the Kowie Estuary over the long term was available in order to compare with the varying physico- chemical conditions during the closed versus open mouth conditions in TOCEs. Data predominantly related to average conditions rather than long term data demonstrating physico-chemical fluctuations. However, the fact that the Kowie River‘s annual flow is erratic due to frequent droughts and floods in the catchment does suggest fluctuations in salinity (Heydorn and Grindley, 1982) and other physico-chemical factors (Baird et al., 1988). These variations, however, are comparatively short in duration, such as freshwater flooding which which can last up to two to four weeks in the Kowie Estuary (Heydorn and Grindley, 1982) and are relatively predictable because such events are linked to seasonal rainfall. In contrast, these fluctuations occur unpredictably in TOCEs causing sudden changes that may last for several months or even years. Consequently, it is anticipated that due to the permanent open mouth condition in POEs, the fluctuations in physico-chemical conditions are not expected to be as spatio-temporally dynamic as they are considered to be in TOCEs. Further, that this elevated spatio-temporal dynamism can have significant ramifications for macrophyte phenology in TOCEs.

Table 1: Characteristics of the temporarily open/closed estuaries compared with permanently open estuaries in the warm temperate region of South Africa

Characteristics TOCE POE Source

Number in SA (of 250) 175 75 Whitfield, 1992; Whitfield, 1998

Size Small and shallow (<2 m) Large (>150 ha) Begg, 1984; Whitfield, 1992

Average period open to 50 % 100 % Begg, 1984 the sea

Mouth condition River flow regime, marine Strong tidal flow/currents Reddering, 1988; regulated by sediment supply, tidal prism and Whitfield, 1992; Cooper, the local wave action 2001

Temperature Summer: 20 - 27°C Summer: 18 - 22 °C Harrison, 2004

Winter: 13 - 20°C Winter: 12 - 14°C

Salinity 15 - 30 ppt 25 - 35 ppt Harrison, 2004

Dissolved Oxygen Well oxygenated Well oxygenated Vorwerk et al., 2001; Content Harrison, 2004

Turbidity <5 NTU <10 NTU Harrison, 2004

Physico-chemical Highly variable and unpredictable More stable, generally occuring Riddin and Adams,

9 Characteristics TOCE POE Source fluctuations over the short term due to 2008a; Whitfield, 2005 floods and droughts

2.2 MACROPHYTE HABITATS 2.2.1 Salt marsh Salt marshes are coastal ecosystems that form a transitional zone between the marine and terrestrial environment (Gribsholt and Kristensen, 2003). Salt marshes are saline (typically at or above seawater) ecosystems with characteristic geomorphology (sedimentary environments, fine soil texture, and relatively flat topography), herbaceous vegetation, and diverse invertebrates and birds (Zedler et al., 2008). They are colonized by halophytic macrophytes that can complete their life-cycles in salty environments (Silvestri et al., 2005). Salt marshes are characteristically comprised of herbaceous forbes, graminoides, succulents and dwarf subshrubs that can tolerate tidal inundation (Adams, 1990; Zedler et al., 2008). Salt marsh formation is initiated due to sediment stabilization by halophytes (Zedler et al., 2008). Plants occur in distinct zones along a tidal inundation and elevation gradient (Davy, 2000; Rogel et al., 2000; Rogel et al., 2001; Bockelmann et al., 2002; Costa et al., 2003; Ursino et al., 2004). They are dynamic systems that respond to varied environmental conditions (Adam, 2000). Flooding regime, salinity and groundwater level are considered the most important factors affecting plant distribution and zonation in salt marshes (Chapman, 1974). Research has been conducted on salt marsh zonation and explanations have been based on the effect of numerous abiotic factors (Streever and Genders, 1997), but primarily relating to tidal inundation, soil salinity and waterlogging (Clarke and Hannon 1970; Ranwell 1972; Vince and Snow, 1984; Adam, 1990; Pennings and Callaway, 1992). Ungar (1998) proposed that halophyte distribution in the lower intertidal zones is governed by physico-chemical factors (predominantly salinity), while biotic competitive interactions control the upper supratidal zones (areas of periodic freshwater inputs and therefore lower salinity). Other authors have recognized other factors that are involved in establishing zonation in salt marsh, such as topography or elevation (Adam, 1990), oxygen, nutrient availability, grazing (Bertness and Ellison, 1987; Mulder et al., 1996), disturbance, such as flooding, drought, wrack deposition (Clarke and Hannon 1971; Zedler et al., 1995; Zedler et al., 2006; Minchinton, 2002), weather patterns (Pielou and Routledge, 1976) and dispersal (Clarke and Hannon, 1971).

Only certain South African estuaries have salt marshes, which are associated with specific environmental conditions within each system (Adams et al., 1992). South African salt marshes have been divided into subtidal, intertidal and supratidal habitats due to this zonation pattern (Adams et al., 1992; Bornman, 2002). The supratidal salt marsh occurs in areas that only get flooded occasionally and is located above the normal spring tide. Hygrophillous grasses such as Sporobolus virginicus, Stenotaphrum secundatum and Cynodon dactylon are found co-existing with Sarcocornia pillansii, crassifolium and Plantago crassifolia (Day, 1981; O‘ Callaghan, 1994; Adams et al., 1992), including the rush species Juncus kraussii Hochst and Juncus acutus L. The grasses are often submerged during high spring tides and form borders between estuarine and terrestrial vegetation. Sarcocornia decumbens, Limonium scabrum and Bassia diffusa occur in a mixed mosaic from the extreme high water spring tide level to the mean high water spring level, known as the upper intertidal zone (Day, 1981; O‘ Callaghan, 1994). Sarcocornia tegetaria and Triglochin bulbosa follow, occurring between the mean high water spring level and the mean high water neap (Day, 1981; O‘ Callaghan, 1994). Below this zone

10 to the zone located below mean sea level is Spartina maritima followed by Zostera capensis, which grows below mean sea level.

Temperate salt marshes are considered one of the most productive ecosystems in the world (Boorman, 1999). They have considerable conservation value because the loss of salt marsh would result in degraded salt pans. These are more easily eroded by water and wind (Bornman et al., 2008). This reduces the function of salt marshes as buffers to sea level rise and storm surges, including the ability of rhizomes to bind the sediment thus maintaining the stability of the shoreline (Davidson and Hughes, 1998). Temperate salt marshes are also home to endemic and rare plants, including migratory water birds (Costa et al., 1987), providing food sources for nekton, avian species and other fauna (van Niekerk et al., 2009; Morton et al., 1987; Adams et al., 1999; Minello 2000; Greenwood, 2008). Salt marsh is a major contributor to shoreline stabilization and transforms and exports nutrients into the marine environment. Salt marshes assist with flood attenuation, improve water quality and contribute significantly to coastal food systems. They provide numerous other ecosystem services, such as their capacity to act as a sink for pollutants, which are retained in an immobile state (Kelly et al., 1998), carbon sequestration due to high net primary production, ground water recharge, sediment transport, sediment and bank stabilization, shelter and food for various estuarine and marine biota, and providing estuarine ecosystems with inorganic and organic nutrients (Long and Mason, 1983; Adam, 1990; Streever and Genders, 1998; Adams et al., 1999; Zedler et al., 2001; Laegdsgaard, 2006; Zedler et al., 2008; Greenwood, 2008).

2.2.1.1 Selected macrophyte species: Supratidal salt marsh habitat Juncus kraussii Hochst J. kraussii has a perennial life-cycle and grows in the supratidal zone. It occurs along estuary banks near the coast in brackish areas and where irregular inundation may occur, dominating salt marshes where salinity is low and freshwater is regularly available (Adams et al., 1999). The plant reproduces asexually by spreading underground through rhizomes, or sexually via seeds. J. kraussii has a paniculate inflorescence with flowers in clusters of 4-6 (occasionally 2) and 15-50 clusters per inflorescence (Cook, 2004). Flowering occurs between October/November and February (spring to summer) and capsules ripen in mid-summer (Jones and Richards, 1954; Muir, 2000; Brown and Brooks, 2002). The ideal salinity range for J. kraussii is < 20 ppt (Adams et al., 1999). They are known to grow in a pH range of 5.4 to 6.6 and a redox potential range of -115 to +200 mV (Clarke and Jacoby, 1994). A study in the Knysna Estuary, South Africa, found that the sediment organic matter content ranged from 45 - 57 % and the moisture content from 78 - 83 % (Muir, 2000). Higher temperatures tend to stimulate germination and optimum germination is in spring (Greenwood, 2008).

Juncus acutus L. J. acutus is a perennial plant that grows in the suptratidal zone of estuaries, occurring along the coast and inland. New appear in spring and die-back occurs in winter, but the appearance of the tufts changes very little (Jones and Richards, 1954). They reproduce sexually via seeds and asexually by spreading underground rhizomes (DNC, 2006) but expansion is largely by seeds. The paniculate inflorescence consists of numerous clusters of 1-6 sessile flowers. Flowering occurs mostly in spring and summer, but can occur throughout the year. The capsules ripen by mid-summer, with inflorescences producing flowers at different stages (Jones and Richards, 1954; Brown and Brooks, 2002; Greenwood, 2008). Each can produce

11 up to 200 seeds and one plant can produce 4 000 seeds (DNC, 2006) with capsules remaining on the plants for long periods (Jones and Richards, 1954). Seeds usually germinate the following spring and summer (Jones and Richards, 1954). This species exploits areas where salinity and inundation levels are moderate, and has the potential to outcompete J. kraussii where salinity is lower (Auld and Medd, 1987; Greenwood, 2008). Salinity of ≥10 ppt is detrimental to growth, while high temperatures stimulate germination and optimum germination is in spring (Greenwood, 2008). It has been found to grow in a pH of 5.9 – 8.6 with sediments rich in organic matter and moisture content but is intolerant of permanent waterlogging (Jones and Richards, 1954).

Sporobolus virginicus (L.) Kunth.

S. virginicus is a rhizomatous perennial grass that grows to heights of 15 - 40 cm and forms thick mats. It has a moderate growth rate and growth occurs most of the year. This grass species inhabits the supratidal areas of estuaries and tolerates highly saline sediments and frequent inundation. The inflorescence is known as a panicle, which flowers mainly from November to May (late spring to late autumn). Peak flowering is usually in summer and autumn, but they can flower throughout the year. Seeds are therefore produced several times throughout the year and are mostly not viable (Duvauchelle and Magee, 2007). Summer is also the peak period for fruit and seeds but production is usually low (Eleuterius and Caldwell, 1984). Vegetative reproduction via rhizomes is considered to be the most successful, although propagation by seeds is possible (Leithead et al., 1976). The ideal salinity range for S. virginicus is 1 - 13 ppt but it has been found growing in salinity levels of 28 – 34 ppt (Breen et al., 1977; Marcum and Murdoch, 1992; Naidoo and Naidoo, 1998; 1999; Muir, 2000). The species can withstand long periods of waterlogging (Breen et al., 1977; Naidoo and Mundree, 1993) and is usually found in sediment pH levels of 6 – 8 (Eleuterius and Caldwell, 1984). A study in the Knysna Estuary, South Africa, found that the sediment organic matter content ranged from 4 - 58 % and the moisture content from 6 - 23 %, while pH ranged from 4.8 - 9 and redox potential from 109 – 352 mV in the plant‘s habitat (Muir, 2000).

2.2.1.2 Selected macrophyte species: Intertidal salt marsh habitat Sarcocornia decumbens (Tölken) A.J.Scott S. decumbens is a perennial shrub that is usually decumbent, forming extensive mats, but can reach 50 cm. The main branches are woody, whilst the segments are cylindrical to obconical and succulent. This species grows in the upper to middle zones of salt marshes, just below the Sarcocornia pillansii zone. According to (Day (1981) and O‘ Callaghan (1994) it grows in the upper intertidal zone. It is usually inundated by spring tides (O‘Callaghan, 1992; Steffen et al., unpublished). The inflorescence is a spike and is usually terminal, with 2–24 fertile segments. They usually contain 3-9 flowers per cyme and mature from the bottom upwards. During the fruiting phase the is corky. Flowering occurs from January to June (O‘Callaghan, 1992) or to July (Steffen et al., unpublished). The ideal salinity range for S. decumbens is 28 - 34 ppt (O‘ Callaghan, 1992), while Davy (2001) suggests that sediments are usually well oxidized with high oxidizing redox potentials. A study in the Knysna Estuary, South Africa, found that the sediment organic matter content ranged from 35 - 41 % and the moisture content from 46 - 52 % (Muir, 2000).

12 Salicornia meyeriana Moss S. meyeriana is an upright subshrub that reaches heights of 40 cm. It occupies the intertidal zone and has an annual life-cycle, dying back during late summer to early autumn when it turns red. Its life-cycle is well defined, with discrete generations (Silva, 2006). The inflorescence is a spike, where the hermaphrodite flowers aggregate in dense, thick spike-shaped thyrses (Kadereit et al., 2006). The number of fertile segments per spike is variable in Salicornia L. showing discontinuities that result in 2 – 30 fertile segments per spike. Six seeds per fertile segment are usually produced. Flowering occurs from February to May in late summer and autumn. The fruit are succulent and small, containing a single seed (Ungar, 1979; Castroviejo, 1990; Silva, 2000; Davy, 2001). Salicornia species are highly salt tolerant (Rozema et al., 1985) and the optimal growth usually occurs at external salinity levels equivalent to less than half that of seawater. Salicornia species are also extremely tolerant of regular flooding and of toxic reduced substances caused by chemical transitions at low sediment redox potentials (Davy, 2001). They are often associated with highly alkaline sediments (Davy, 2001), which are usually representative of high pH.

Sarcocornia tegetaria S. Steffen, Mucina & G. Kadereit S. tegetaria (Sarcocornia perennis (Miller) A.J. Scott.) has a perennial life-cycle and is a prostrate to decumbent subshrub. Plants form large dense mats which can attain heights of 20 cm. The leaves are grey-green to reddish and the margins are truncate (O‘Callaghan, 1992; Steffen et al., 2007). It grows in the middle to lower zones of salt marshes where it is regularly flooded by tidal action and is endemic to southern Africa (O‘Callaghan, 1992; Steffen et al., 2007). Steffen et al. (2007) hypothesise that S. tegetaria frequently hybridizes with other species of the genus, for example S. pillansii and S. natalensis, including but less frequently with, S. capensis and S. littorea. The inflorescence is a spike containing 2–22 fertile segments that taper at the end. Each fertile segment usually contains three (rarely four or five) flowers, which mature from the bottom upwards and produce six seeds. Flowering occurs from January to June (O‘Callaghan, 1992) or to July (Steffen et al., unpublished) in late summer and winter. During the fruiting phase the perianth turns corky. Effective reproduction is by seeds, which are abundantly produced under favourable conditions (O‘Callaghan, 1992; Davy, 2006). The ideal salinity range for S. tegetaria is 0 - 15 ppt (Adams and Bate, 1994a). A study in the Knysna Estuary, South Africa, found that the sediment organic matter content ranged from 35 - 41 % and the moisture content from 46 - 52 %, while pH ranged from 6.6 – 7.1 and redox potential from +283 to +495 mV (Muir, 2000).

2.2.2 Reeds and sedges

The reeds and sedges are a peripheral community located on the estuary banks which either grow in soft intertidal or shallow subtidal substrates with their photosynthetic portions partially and/or periodically immersed (Adams et al., 1999). This community usually occurs in the brackish upper reaches of estuaries, but can extend into the intertidal zone where they are inundated with saline water (Adams and Bate, 1999a). The reeds and sedges consist of species that grow in brackish waters and areas where occasional flooding occurs. Schoenoplectus littoralis (Schrad.) Palla; Typha capensis (Rohrb.) N.E. Br.; Cyperus laevigatus L. and Schoenus nigricans L grow in the fresher waters. Phragmites australis (Cav.) Trin ex Steud and Schoenoplectus triqueter (L.) Palla are found in the brackish areas, whereas Bolboschoenus maritimus (L.) Palla and J. kraussii are infrequently inundated in the suptratidal salt marsh zone. P. australis and B. maritimus

13 can form stands in marine or tide dominated estuaries where salinity is high but only where small freshwater streams flow and groundwater seepage occurs (Adams and Bate, 1999a; Muir, 2000).

Reeds and sedges are an important habitat for several reasons. For example, they stabilize banks and reduce erosion, prevent eutrophication through denitrification, provide detritus for nutrient cycling as well as providing a substratum for periphyton and bacteria (Riddin and Adams, 2008a). Reed stands are important for wintering, foraging, refuge and breeding of migrant birds (Haslam, 1971; Kassas, 2002; Eid et al., 2010), including numerous invertebrates and fish species; and are considered an important resource due to their high productivity (Westlake, 1963; Haslam, 1971). Reed detritus provides food for mollusks, crustaceans, and aquatic insects (Gucker, 2008). Whitfield (1980) found that the Phragmites reeds were critical as a detritus source for fish in the Mhlanga Estuary, South Africa. Reeds also have an application in (food production or weed control) and nature conservation (Güsewell and Klötzli, 2000). According to Brix and Schierup (1989), the aerenchyma of P. australis enhances oxyen availability in the sediment and is important in balancing nitrogen, phosphorous and silica (Hocking, 1989; cited in Eid, 2010). In a reed dominated small freshwater tidal marsh, reed decomposition contributed to the export of dissolved silica by 40 % thereby enhancing the capacity of salt marshes in recycling silica (Struyf et al., 2007). Emergent macrophytes have also been shown to accumulate heavy metals for the purposes of wastewater treatment (Rai, 1995).

2.2.2.1 Selected macrophyte species: Reed and sedge habitat Phragmites australis (Cav.) Trin ex Steud P. australis is a perennial reed that spreads asexually through stolons, rhizomes and sods or sexually via seeds (Gucker, 2008). Plants grow in and along the margins of estuaries, lakes and streams in permanently inundated waters, the intertidal zone or the supratidal zone. The inflorescence is a feathery panicle with numerous spikelets containing 1-3-8-10 flowers (Gucker, 2008). Flowering occurs from December to June (mid-summer to early winter) and fruits are released during winter and spring (Haslam, 1972; Auld and Medd, 1987; Boedeltjie et al., 2004). Many studies indicate that the production of viable seeds is variable (Kettenring and Whigham, 2009; Baldwin et al., 2010; Kettenring et al., 2010). Some studies have estimated that 350-800 viable seeds could be produced per inflorescence, compared to 500-2 000 seeds per shoot (Gucker, 2008). In southwestern Japan, seed set averaged 9.7 % and ranged from 0.1 - 59.6% for 12 common reed populations (Ishii, 2002). It is known to have a competitive advantage in waterlogged and nutrient enriched sediments (Haslam, 1970; Human, 2010). The ideal salinity range is 18 - 30 ppt, while reduced growth has been observed at >15 ppt and inhibited growth at 20 ppt over a two week period. Die-back was observed at 30 ppt during an inundation period of 94 days (Adams and Bate, 1999; Benfield, 1994; Lissner and Schierup, 1997). It is known to grow in a wide range of sediment pH, from 2.9 – 9.2 (Duke, 1978; 1979; Gucker, 2008). The main regulatory environmental factor is considered to be water depth (Grime, 1979; Haslam, 1971; Lieffers and Shay, 1981; Grace, 1989; Hellings and Gallagher, 1992) with ideal water depths ranging from 0 – 2 m (Adams and Riddin, 2007).

14 Bolboschoenus maritimus (L.) Palla B. maritimus (Schoenoplectus maritimus, Scirpus maritimus) is a perennial sedge that reproduces asexually through underground rhizomes and sexually via seeds. Plants inhabit estuaries, seasonal and permanent wetlands and marshy environments along the outer zone in fresh and brackish tidal areas (Clevering et al., 1996). Plants can grow where flooding and depths are moderate (Espinar et al., 2005) and are adaptable to saline and non-saline environments. They survive periods of drought by the rhizomes remaining dormant. The inflorescence is a solitary spike with 1-10-(-20) ovoid spikelets and numerous flowers. Flowering occurs in spring and summer from October to March (Dykyjová, 1986; Wilman, 2006). Each achene produces a single seed and achenes overwinter in the sediment, but occasionally in the spikelets and gradually shed until spring. However, germination is relatively rare in mature stands. Although achenes germinate poorly at maturity, 97 % germination can be obtained two months later and achenes can remain viable for two years (Kantrud, 1996). In Mediterranean climates, sprouting of below-ground parts starts at the onset of the wet season (autumn). During winter the shoots emerge through the water column and fruiting/seeding occurs through the spring (Espinar et al., 2004, 2005; Diggory and Parker, 2010). Plants then die-back in the dry season (Espinar et al., 2004, 2005; Wilman, 2006). In colder climates, vegetative growth commences in the spring through to summer (Dykyjová, 1986; Clevering and van Gulik, 1997). Seeds tend to be heavier while culms are longer but fewer in tidal areas compared to non-tidal areas (Clevering and van Gulik, 1990 cited in Clevering and van Gulik, 1997). The ideal salinity range is 18 - 30 ppt (Adams and Riddin, 2007), but is most successful in freshwater or salinity <10 ppt, although salinity tolerances of >45 ppt have been observed in mature stands, particularly when occasional freshwater inputs occur (Hellings and Gallagher, 1992; Lissner and Schierup, 1997; Chambers et al., 1998; Greenwood, 2008). Water salinity from 11.5 - 15.4 ppt has been shown to negatively effect growth, corm sprouting and achene germination (Kantrud, 1996). Ideal water depth ranges from 0 – 0.2 m. An increase in depth from 44 to 80 cm has been observed to reduce growth (Lieffers and Shay, 1981; Coops et al., 1996; Adams and Riddin, 2007).

2.2.3 Submerged macrophytes

Submerged macrophytes are flowering plants or angiosperms that are anchored in soft subtidal or low intertidal substrata and adapted to be completely submersed for most states of the tide (Adams, 1994). These plants are primary colonizers of mudflats and sandflats and grow in varying salinity, ranging from polyhaline (>30 ppt) to fresh (0 ppt) (Day, 1981). Most submerged macrophytes inhabiting estuaries appear to occur in salinity levels between 10 and 20 ppt (Adams and Riddin, 2007). Fresh and brackish waters are usually occupied by species such as Chara spp, Ruppia spp, Potamogeton spp, Ceratophyllum spp, Myriophyllum alterniflorum and Elodea canadensis (Haramis and Carter, 1983; Collins et al., 1987). Common temperate macrophytes are Zostera, Halophila, Ruppia, Potamogeton and Zannichellia (Day, 1981). Zostera capensis is usually associated with the marine end of POEs (Adams et al., 1992) in the intertidal and shallow subtidal zone, while Halophila ovalis is frequently associated with Z. capensis because of its opportunistic nature (Day, 1981). Chara vulgaris, Ruppia cirrhosa, Zannichellia and Potamogeton pectinatus are the most common in less saline or brackish South African estuaries (Howard-Williams, 1980; Lubke and van Wijk, 1988; Adams, 1994). Although Chara is a macroalga, it was included as a submerged macrophyte because it inhabits the same habitat and performs the same habitat function as R. cirrhosa (Riddin and Adams, 2008a). R. cirrhosa is common in TOCEs where salinity fluctuates as it can tolerate high salinity (2 - 40 ppt). It also grows in the calm brackish upper reaches of

15 POEs (Adams et al., 1992; Adams, 1994; Adams and Bate, 1994b/c; Adams et al., 1999). When exposed due to the estuary mouth breaching, submerged macrophytes die-back rapidly (Adams and Bate, 1994c; Riddin and Adams, 2008a).

Submerged macrophytes have been recognized for their high productivity, which is transferred to secondary consumers including herbivores, detritivores and microorganisms. They provide substrate for epiphytic organisms and the physical structure of these underwater meadows provides shelter and refuge for various marine and estuarine biota (Touchette, 2007). For example, in the East Kleinemonde, West Kleinemonde and Kasouga estuaries in South Africa maximum abundance and biomass of Exosphaeroma hylocoetes, an estuary isopod, was recorded in the middle reaches and was attributed to the presence of submerged macrophytes (Henninger et al., 2008). Aquatic macrophytes and associated invertebrates are an important food source for fishes (Whitfield, 1984; Li et al., 2010; Sheppard, 2010), while they act as a refuge for herbivorous zooplankton against fish predation (van Donk and van de Bund, 2001). Lammens (1989) writes that one of the major factors structuring the fish community of shallow eutrophic lakes is the occurrence of submerged vegetation (van Donk and van de Bund, 2001). It is also an important food source for avifauna. For example, coots (Fulica atra L.) consumed 20 % of the maximum biomass of Ruppia (100 g m−2) during the period from September to November in a brackish pond on Texel, The Netherlands (Verhoeven, 1979). Submerged macrophytes oxygenate the water column and increase the depth of the oxidized microzone at the sediment surface and maintain good water quality (Adams et al., 1999). Seagrass meadows reduce current velocity thereby stabilizing the sediment and protecting the root- complex from scouring (Fonseca et al., 1982) and prevent sediment re-suspension (Adams et al., 1999). Macrophyte beds are critical in the stabilization of clear water in shallow, mesotrophic and eutrophic lakes by affecting the planktonic food web (van Donk and van de Bund, 2001). Chara beds act as nutrient sinks which restrict the availability of nutrients in phytoplankton and the incidence of eutrophication (Kufel and Kufel, 2001 cited in Coops, 2002), while van Donk and van de Bund (2001) found that they have a stronger influence on water transparency compared to other submerged macrophytes. Their ability to exchange dissolved ions from their thin cuticles with the surrounding water aids ecological processes, such as nutrient cycling (Adams and Riddin, 2007). Submerged macrophytes have also been shown to accumulate heavy metals and are sometimes used for the purposes of wastewater treatment (Rai et al., 1995).

2.2.3.1 Selected macrophyte species: Submerged habitat Ruppia cirhosa (Petagna) Grande

R. cirrhosa is a facultative annual or perennial aquatic macrophyte found in estuaries anchored to the sediment underwater by rhizomes. They grow in a few centimeters to two metres of saline to brackish waters and are adapted to a wide range of salinity, depth, light and temperature (Gesti, 2005). The inflorescence is a spike with two tiny bisexual flowers. Numerous flowers develop within five to six weeks underwater, after the onset of spring growth, and the peduncle length is determined by water depth (Verhoeven, 1979). The plants are self or cross pollinated, but the former enhances fruit production. Drupes are formed two weeks after flowering (Verhoeven, 1979), each containing one seed that can remain viable in sediments for up to three years (Kantrud, 1991). They are common in TOCEs characterized by fluctuating salinity but are also found in the brackish upper reaches of POEs (Adams et al., 1999). Ideal salinity and light requirements range from 0 - 30

16 ppt and 17.5 - 42.5 ppm suspended sediments respectively (Kantrud, 1991; Adams and Riddin, 2007). Although they can withstand fluctuating salinity levels of up to 75 ppt, low salinity (<20 ppt) is important for seed germination (Adams and Riddin, 2007). Verhoeven (1979) reports on pH levels of 7.4 - 10.4 for waters in which Ruppia taxa grew and that turbidity generally reduces Ruppia cover. A maximum water depth of 1.5 m resulted in healthy R. cirrhosa beds, although greater depths (7 m) have been observed (Luther, 1951 cited in Verhoeven, 1979).

Chara vulgaris L.

C. vulgaris is an aquatic macroalgae that grows completely submerged in water a few centimeters to 2 m deep, although some charophytes have been found in depths of 27 m (Coops, 2002). They are anchored to the sediment by rhizoids. C. vulgaris usually inhabits still, clear water that is fresh, alkaline or brackish and tends to die when there is eutrophication and an increase in turbidity. They are monoecious or dioecious plants that reproduce by means of numerous oogonia (female organs) and antheridia (male organs). The oogonia are black when fertilized and develop a white calcium layer around the oogonia called a gyrogonite, which is a highly resistant survival organ (Soulié-Marsh, 2007). The oogonia drop from the parent plant and germinate to produce a new plant. They can remain dormant in the sediment for years, survive long periods (up to decades) of desiccation and germinate when favourable conditions return, such as increased water depth, good light, low turbidity and warmer temperatures (22°C) (de Winton, 2000; Coops, 2002; Cook, 2004; de Winton, 2004; Kalin and Smith; 2005; Sederias and Colman, 2007). Experiments by Kalin and Smith (2005) showed that fluctuating redox potential is a likely trigger in the germination of C. vulgaris oospores, while germination was higher under normal light conditions compared to the dark. Reproduction in dioecious perennial charophytes is stimulated by warmer temperatures while vegetative growth and expansion occurs early in spring (Casanova and Brock, 1999). In the East Kleinemonde Estuary, a small South African TOCE, the re-establishment of submerged species is best under low salinity (<20 ppt) (Riddin and Adams, 2008b).

17

Figure 2.2: Macrophyte species selected for the study.

18 2.3 EFFECT OF ENVIRONMENTAL CONDITIONS ON THE PHENOLOGY OF MACROPHYTES

The reproduction, germination and development of estuarine macrophytes is dependent on numerous physiological needs, broadly related to either a sufficient input of energy and vital substances (e.g. water, oxygen, light, salt ions, macro-nutrients and micro-nutrients) or to the limitation of stressful factors (e.g. soil waterlogging, sediment toxicity, sudden temperature changes, inter and intra-specific competition) (Silvestri, 2006). Biotic and abiotic factors both play a significant role in determining biomass production and reproductive effort in estuarine macrophytes (Ungar, 1987a/b). Several environmental factors acting synergistically may influence their survival, for example sediment salinity, groundwater, sedimentation and scouring by tidal movement, rainfall and other environmental variables (Ungar, 1987a/b). Other factors such as competition, degree of waterlogging, and nutrients status have also been shown to regulate biomass production (Mahall and Park 1976a/b; Pennings and Callaway, 1992; Scarton, 2002; Pennings et al., 2005). Elevation has been regarded as the most important environmental factor affecting sediment salinity, aeration, water depth and inundation periods (Armstrong et al., 1985; Wolters et al., 2008). All these factors act together as stressors, affecting plant growth and reproduction, while shifting community dynamics (Greenwood, 2008).

The variable and stressful nature of estuarine environments requires plants to be plastic in both their growth and reproductive phenology to ensure their persistence. Environmentally-induced phenotypic variation or phenotypic plasticity in plants is often considered to be a functional response that maximizes fitness in variable environments (Chapin et al., 1987). For example, Pucchinellia maritima has a variable breeding system as an adaptation mechanism to the harsh salt marsh environment by combining clonal spread and sexual reproduction. It also shows considerable phenotypic plasticity by adopting a mat-like growth form under intense grazing, compared to pioneer forms that are as tall as 50 cm. Mature plants show adaption through genetic and phenotypic variation by growing taller and producing copious seeds (Packham and Willis, 1997). Understanding these responses is essential to determine the limitations that estuarine macrophytes might display in relation to fluctuating environmental conditions that often occur in TOCEs so as to inform the management practices of these systems which are modified by anthropogenic impacts.

Although there are numerous physico-chemical factors that act synergistically together to affect macrophyte production and reproduction, the following variables have been considered in this study: climate, salinity, water regime, tidal exchange, redox potential, pH, sediment organic matter, sediment moisture content and light, turbidity and temperature. The sections below discuss these factors in relation to both the growth and reproductive phenology of estuarine macrophytes in more detail.

2.3.1 Climate: Seasonal changes in temperature, rainfall, irradiance and photoperiod Climate has been regarded as the dominant environmental factor controlling the broad scale distribution of vegetation due to the geographic gradients of rainfall and temperature (Nakamura et al. 2007; Zhu et al. 2007 cited in Ji et al., 2009). Regional climate is important in determining the location and type of salt marsh, for example, tropical latitude salt marshes have a distinctly different species composition than those of the temperate latitudes. Temperature and rainfall are the most likely causes of these patterns due to their effect on salinity (Laegdsgaard, 2006). Wolff and Jefferies (1987) observed that in the Canadian Arctic, where the

19 growing season is only three months long and conditions are severe, Salicornia plants reach heights of 1–10 cm and produce simple branches. In comparison, plants growing in the temperate latitudes, where the growing season is seven to eight months long, plants reach heights of 40 cm and produce multiple branches (Davy, 2001). Dreyer et al. (2006) found that the active growing period of Oxalis species in the Western Cape region of South Africa coincided with the peak rainfall period and that the start of flowering was dependant on average daily temperatures declining and rainfall increasing.

Seasonal changes in temperature have a major influence on seed production and germination, including the activity, growth and distribution of salt marsh plants (Packham and Willis, 1997). In Mediterranean climates, plants that dominate salt marshes grow best in the warm season (Zedler et al., 2000) and according to Onaindia et al. (2001) summer is the period for the optimum vegetative development of halophytic species. Seasonal growth patterns in estuarine macrophytes have been recorded in many salt marsh environments, while spring growth usually occurs in mediterannean-type ecosystems (Kummerow, 1983 cited in Pierce, 1984). Congdon and McComb (1980) investigated seasonal changes in J. kraussii and, although new culms were produced throughout the year, there was evidence for high growth in the warm season. Maximum biomass accumulated during summer suggesting a warm season growth pattern. Flowering occurred in spring, as temperature is also one of the environmental factors controlling several reproductive events, such as the initiation of flowering and the germination of seeds (Santamaria and van Vierssen, 1997).

In Britain, the growth of Sarcocornia perennis shows a seasonal trend with buds that develop into new branches occurring in spring. Flowering takes place from August to September in summer and autumn and seeds ripen in October (autumn). Although plants remain green in winter they are moribund (Davy, 2006). Salicornia species have a wide climatic range which corresponds with the 10°C July isotherm. Here they are confined to low, south-facing slopes, where temperatures are up to 7°C higher than the north-facing aspect (Jefferies et al., 1983) and relatively high daily radiant energy occurs during the growing season (Davy, 2001). They also flower during summer and autumn; and then germinate during the spring. In Spain, the biomass of S. perennis ssp. alpini showed distinct seasonal patterns of growth and decay, which is a characteristic of temperate zone coastal marshes where evapo-transpiration is high during summer (Curcó et al., 2002; Scarton et al., 2002; Palamo, 2009). The growing season of Salicornia fruticosa is from spring to autumn (Scarton et al., 2002). Although Spartina alterniflora flowers seasonally, the flowering date varies throughout its geographic distribution from summer to autumn. Eleuterius and Caldwell (1984) found that the flowering of salt marsh plants in Mississippi was influenced by temperatures and peak flowering was in summer, while Seneca (1974) explained earlier flowering in part due to a shorter photoperiod, which is associated with spring and summer. Photoreceptor proteins enable flowering plants to sense seasonal changes in night length (photoperiod) which is a signal to flower (Dutta, 1979).

Zostera growth is seasonal and closely associated with temperature. In Britain, growth occurs during spring and summer, while in Danish waters biomass also increases substantially during this period (Sand-Jensen and Borum, 1983). Sexual reproduction in Zostera species takes place from summer to autumn (Brown, 1990; Tubbs and Tubbs, 1983) and relatively high temperatures (>15°C) appear to be required for flowering and seed germination. Perennial populations remain dormant as rhizomes during the winter and produce new leaves in

20 the spring, while annual populations die-back in winter. They also demonstrate considerable morphological plasticity depending on environmental conditions (Davison and Hughes, 1998).

In Lake George, America, Stross (1979) recorded that the growth of the submerged species, Nitella flexilis (L.) Ag., shoots were most active at 18°C during summer (June) and that the affect of temperature on growth appears to be general for other charophytes and macrophytes. In Lake Pontchartrain, America, L. s.l. showed a seasonal growth and reproductive pattern with vegetative growth occurring in summer and high water temperatures instigating two seasonal peaks of biomass and flowering (Cho and Porrier, 2005).

Research on the impact of several environmental factors on the morphology of Chara vulgaris demonstrated that they had different seasonal patterns and some of the morphological characteristics had significant fluctuations, indicating differences in their seasonality (Hu et al., 2008). The emergence of sex organs and oospore maturation of C. vulgaris was positively affected by irradiance and photoperiod. Further, that plant growth and reproduction acclimated morphologically and physiologically to different irradiance levels and photoperiods and that the differences can partly explain the broad geographic distribution of C. vulgaris (Wang et al., 2009).

Salt marsh plants usually germinate in spring (Packham and Willis, 1997; Davy, 2001). Many temperate salt marsh species require chilling to overcome dormancy followed by cool temperatures for germination. This would target the timing of germination to early spring (Baskin and Baskin, 1998). The timing for germination in salt marshes may be infrequent and brief, limited to perids of rain, between tides and at the right time of the year (Fenner and Thompson, 2005). Egan and Ungar (1999) investigated the influence of temperature and seasonal changes on Salicornia europeae and Atriplex prostrata and found that both species germinated in spring and early summer when salinity was lower due to spring rains. The annual A. prostrata completed its lifecycle in seven months, from April to November during spring and autumn. Further, that high summer temperatures inhibited germination, especially when salinity was high. In the Olifants Estuary, South Africa, high rainfall initiated germination due to a decrease in salinity during late spring (Bornman, 2002). Numerous salt marsh plants germinate in spring after rains, coinciding with mild temperatures and reduced soil salinity (Allison, 1996). Temperature fluctuation is also known to initiate germination in many marsh species, for example P. australis and Typha latifolius (Lombardi et al., 1997; Ekstam et al. 1999, Fenner and Thompson, 2005).

Thompson and Grime (1983) showed that seed germination of 42 % of the 66 herbaceous species studied, was stimulated by alternating temperatures, and the capacity to respond to fluctuating temperatures was high in the wetland species. The ecological explanation for this requirement is that during spring, as favourable conditions for germination of wetland plants are created by falling water levels and rising temperature, shallow water or bare mud experiences large temperature alternations (Fenner and Thompson, 2005). Typha orientalis and P. australis both respond to an increase in the amplitude of alternating temperatures, but Typha only responds to these alternations at low temperatures (Fenner and Thompson, 2005). The low-temperature limit for seed germination is unknown, but germination in many species may be prevented by freezing only (Fenner and

21 Thompson, 2005). For example, Grime et al. (1981) found that many herb species germinated at temperatures as low as 5°C. This suggests that in dynamic environments, such as TOCEs, the possibility for germination to occur outside of the typical seasonal germination period may be possible.

It is clear from the research cited above that seasonal change, with the associated temperature, rainfall and irradiance fluctuations trigger changes in both the growth and reproductive phenology of estuarine macrophytes (Thompson and Grime, 1983; Packham and Willis, 1997; Zedler et al., 2000; Laegdsgaard, 2006). It is anticipated in POEs that these seasonal dynamics will be a major determinant in the annual growth and reproductive phenology of macrophytes. In comparison, because TOCEs are dynamic systems that may experience conditions which do not coincide with the preferred seasonal growth or germination period, it is anticipated that these seasonal ‗triggers‘ may be less important than the condition of the estuary mouth. For example, inundation of the supratidal and intertidal zones due to a closed estuary mouth may occur during spring and summer when germination, vegetative growth and flowering are generally at their peak (Adams and Bate, 1994b; Packham and Willis, 1997; Baskin and Baskin, 1998; Davy, 2001; Riddin and Adams, 2008b). This will prevent salt marsh plants from producing viable seeds and replenishing seed banks. However, due to the plastic nature of salt marsh species, vegetative growth or flowering may occur outside of the normal season or later on in the season over a shorter period.

2.3.2 Salinity Salinity is considered the primary stressor in salt marshes and predominantly responsible for changing vegetation patterns (Ungar, 1978; Armstrong et al., 1985; Bertness and Shumway, 1993). Salinity stress affects germination, seedling establishment, reproduction and productivity (Chapman, 1974; Ungar, 1978; O‘Callaghan, 1992) and reduces water potential, causing ion imbalance or inducing toxicity (Greenway and Munns, 1980; Ashraf, 2004). Depending on their reaction to saline environments, two broad categories of plants have been recognised, namely halophytes (salt tolerant) and glycophytes (salt sensitive) (Waisel, 1972; Matsumura et al., 1998; Jordan et al., 2002 cited in Greenwood, 2008).

Salinity changes are a primary determinant of macrophyte species composition (Adams and Riddin, 2007). The vertical and horizontal distribution of salt marsh is controlled by sediment salinity, which is influenced by groundwater seepage, run off, rainfall, evaporation and tidal exchange. The germination and productivity of salt marshes is adversely affected when salinity is increased (O‘Callaghan, 1992; Adams and Ngesi, 2002). Salinity can fluctuate extensively due to changes in tidal action, evapo-transpiration, precipitation and availability of fresh groundwater. These fluctuations require high physiological plasticity and result in strong phenotypic variation (Kadereit, 2007). Research by Aldous (2003) on the potential for the use of S. virginicus in saline areas concluded that shoot growth had considerable morphological plasticity relative to high salinity levels ranging from 18 - 24 ppt, although Breen (1977) demonstrated that the tolerance range is as much as 30 ppt. Marcum and Murdoch (1992) recorded a linear increase in roots in response to salinity which produced a high root/shoot ratio at a salinity of 450 mM NaCl.

22 Salt marsh plants predominate at salinity values ranging between 10 and 35 ppt (Chapman 1960, Day 1981, Riddin and Adams, 2008a) although certain species are more sensitive to higher levels than others (Snow and Vince, 1984). Ungar et al. (1979), for example, reported that the distribution of Salicornia europaea on an inland salt pan at Rittman, Ohio, was associated with sediment salinity and competition. Scarton (2002) observed that S. europaea was very competitive in highly saline soils because it showed optimal growth at salinity levels that were too high for other species. Various studies have found that Salicornia is better at tolerating hypersaline soils than other marsh species, grows faster, accumulates more biomass and occupies more cover when soil salinity is high (Mahall and Roderic, 1976; Pearcy and Ustin, 1984; Allison, 1996; Zedler and Beare, 1986 cited in Zedler, 2003). Salicorniodieae often form monospecific stands where salinity becomes toxic for other plants (Davy et al., 1990; Freitag et al., 2001; Kadereit, 2006). The submerged macrophyte, R. cirrhosa, can tolerate a high salinity of 42 ppt, however hypersaline conditions over the long term are usually detrimental to salt marsh plants (Riddin and Adams, 2008a). Myriophyllum spicatum prefers less saline conditions and has a salinity tolerance range of less than 13.3 ppt. As a result of increased freshwater input into the permanent marshes of the Camargue (southern France) it has displaced Potamogeton pectinatus, which prefers more saline conditions, namely 0 – 6 mg L-1 (van Wijck et al., 1994) or 5 - 15 ppt (Adams, 1994).

Freshwater impoundments reduce downstream flow and can cause adverse affects, such as an increase in the salinity of estuaries, creating hypersaline conditions. A reduction in freshwater can threaten estuary ecosystems as the flora and fauna are adapted to varying freshwater inputs and associated salinity levels (Adams et al., 1992). Impoundments on the Great Brak Estuary have decreased the occurrence and cover of Cotula coronopifolia since 1989 due to increased salinity because this species is usually associated with brackish conditions (Adams and Ngesi, 2002). The Orange River Estuary has been affected by the construction of 23 major dams on the Orange River. Seawater penetration into the estuary was the result of low flows during 2004 and 2005, which caused the less salt tolerant species, such as P. australis and Schoenoplectus scirpoideus, to die-back in the lower reaches (Bornman and Adams, 2010). Large watershed scale dams and drought in Texas resulted in hypersalinity and the loss of biota and invasions by stenohaline species (Copeland 1966; Hoese 1967 cited in Montagna et al., 2002). In the Kromme Estuary, the cover and biomass of Zostera capensis, which is tolerant of high salinity (2 - 40 ppt), has increased due to reduced flooding and increased sediment stability caused by the construction of a second dam on the river (Adams et al., 1992; Adams, 1994). Consequently, salinity levels in TOCEs may be increased where impoundments are constructed upstream of the estuary because freshwater input is curtailed. This may result in hypersaline conditions, which in turn cause macrophyte community structure to be modified to the detriment of ecological processes associated with them e.g. nutrient and food supply.

Numerous research findings indicate that high salinity reduces vegetative production. Congdon and McComb (1980) suggested that high salinity may have induced senescence of J. kraussii plants thereby reducing biomass. A study by Charpentier et al. (2009) indicated that increasing salinity (from 0 – 18 g/l NaCl) had a negative effect on the number of leaves per Juncus gerardi seedlings after 110 days of experiment. S. perennis can tolerate high salinity (Rubio-Casal et al., 2003) while S. tegetaria shows signs of deterioration at ≥35 ppt (Adams and Bate, 1994a) and the species show substantial phenotypic plasticity (Silva et al., 2006). Silva et al. (2006) found that two populations of Salicornia ramosissima differed significantly in their growth responses due

23 to the differences in salinity of the two sampling sites. The one site was inundated daily with mean salinity values of 11 ppt. The other site was not inundated due to being positioned 10 m from the margin of the channel and experienced mean salinity values of 34 ppt. The greenhouse findings showed that growth was greatest after treatments of freshwater and generally decreased with increasing salinity. Small TOCEs are more sensitive to changes associated with the volume of freshwater input due to the impact on salinity. Where there is a lack of freshwater input, high salinity values (>45 ppt) have caused impoverishment of the estuarine flora (Adams et al., 1992). Experiments by Hellings and Gallagher (1992) recorded that the biomass, density and height of P. australis was negatively affected by increasing salinity over the range from 0 - 30 ppt. In Denmark, P. australis die-back was observed in the lower fringe of stands where the soil water salinity was >15 ppt within the rooting depth (Lissner and Schierup, 1996). Chambers et al. (1998) found that nitrogen uptake rates of P. australis at a salinity of 20 and 30 ppt decreased significantly compared to the freshwater treatment. After a storm breach in the East Kleinemonde Estuary, the salinity increased significantly from 21.9 ppt to 31 ppt, including the water level. The supratidal salt marsh and the reeds and sedges cover declined significantly by 15 and 19.7 % respectively. The intertidal salt marsh and submerged macrophytes were also reduced (Riddin and Adams, 2010).

High salinity is decisive during the reproductive period reducing both reproductive output (flowering and seeds) and seed germination (Ungar, 1962; Chapman, 1974; Ungar, 1977; Ungar, 1978; Ungar; 1998; Riddin and Adams, 2008a). Seed germination in halophyte grasses is usually delayed under high salinity (Gulzar and Khan, 2001). In salt marshes with a mediterranean-type climate, the germination of most species is cued to variations in salinity, not temperature (Noe, 2002). Although the seeds are not permanently affected by storage under waterlogged saline conditions, young plants are influenced by salinity while older plants are more tolerant (Ungar et al., 1977). The spread of Phragmites in tidal wetlands is constrained by salinity, and invasion by seeds and rhizomes are therefore restricted to areas where the salinity is <10 % (Chambers et al, 2003). Both J. kraussii and J. acutus achieved high final percentage germination in salinity levels ≤15 ppt, and seed viability was not compromised by salinity (Greenwood, 2008), whereas S. tegetaria achieved 82 % germination success at 0 ppt compared to 57 % at 35 ppt in seeds recovered from the seed bank (Riddin and Adams, 2009).

The flowering phenology of salt marsh plants in relation to salinity has been researched by a few authors. Flowering in Sarcocornia quinqueflora is triggered by an increase in salinity during late summer (Clarke and Hannon, 1970). Salicornia dolichostachya (‗S. oliveri Moss‘) did not flower and showed poor growth when grown without salt (Davey, 2001). In Delaware, P. australis flowered at salinity nearing 65 ppt (unpublished results of Mills and Gallagher in Hellings and Gallagher, 1992). A study by Charpentier et al. (2009) in the Rhone Delta, France, observed that high salinity negatively affected inflorescence production in Juncus gerardi. Inflorescence biomass was significantly greater in high salinity levels (12-18 g L-1 NaCl) in water depths of 0-5 cm compared to 20 cm, but did not differ in the various water depth treatments when grown in a low salinity range (0-6 g L-1 NaCl). The number of both flowering and vegetative shoots decreased with increasing salinity (0-18 g L-1 NaCl), from 144 - 139 and 242 - 72 respectively, while fruit maturation was faster under saline conditions. The reproduction of female Chara aspera plants in fresh and brackish waters produced 70 % more gametangia at 0 than at 10 ppt, while 57 % developed into oospores in freshwater compared to 8 % in brackish waters. Similarly, male plants from freshwater habitats had a higher fertility when compared to those from

24 brackish water habitats (Blindow et al., 2009). To minimise the stress of acquiring water and nutrients, most salt marsh plants flower and set seeds infrequently (Huiskes et al., 1995); and during times when they are less likely to experience high salinity or inundation stress (Laegdsgaard, 2006). Leiffers and Shay (1982) found that B. maritimus did not flower in dry and highly saline sites. Deegan et al. (2005) recorded that the reproduction of S. triqueter was significantly reduced at a salinity of 10 ppt, and research by Coops and Smit (1991) demonstrated that salinity levels of 5.8 ppt stimulated spikelet and achene formation in B. maritimus.

TOCEs experience fluctuating salinity regimes as a result of changing mouth conditions. For example, during the closed mouth condition, hypersalinity is not uncommon and salinity usually ranges from 15 - 23 ppt or > 25 ppt in the East Kleinemonde Estaury. In comparison, during the open mouth phase the salinity usually ranges from 37 ppt in the lower and middle reaches to 1 ppt in the upper reaches (van Niekerk et al., 2008). Although intertidal salt marsh can tolerate fluctuating salinity (Adams and Bate, 1994a), a protracted period of closed mouth conditions with no freshwater inflow can cause the estuary to become hypersaline (Riddin and Adams, 2010). As demonstrated in the literature above, these conditions are detrimental to several macrophyte plants in terms of their growth and flowering phenology (O‘Callaghan, 1992; Adams et al., 1992; Chambers et al., 1998; Adams and Ngesi, 2002 Silva et al., 2006; Charpentier et al., 2009; Bornman and Adams, 2010; Riddin and Adams, 2010). In comparison, drought conditions in the Kowie Estuary would result in the reduction of freshwater inflow along drainage areas e.g. valley troughs causing an increase in salinity that is detrimental to macrophytes, such as P. australis and B. maritimus (Leiffers and Shay, 1982; Deegan et al., 2005; Greenwood, 2008; Bornman and Adams, 2010), growing in these zones.

2.3.3 Water regime Salt marsh plants vary in their adaptability and tolerance to a particular water regime and interspecific differences in plant responses to fluctuating water level are fundamental to habitat differentiation in wetlands (Noe, 2002; Olff, et al., 1988; Mitsch and Gosselink, 1993). Under favourable conditions, and assuming other environmental factors are not limiting, certain species will grow vigorously and flower profusely. Under unfavourable conditions, certain species will not grow and flowering is improbable or even absent (Roberts et al., 2000). Fluctuating water levels are a component of the water regime which is described by the water depth including the rate, amplitude and timing of flooding or draw-down events (Deegan et al., 2007). In POEs the water regime is tidal or cyclical and comparatively stable compared to TOCEs, where the water regime fluctuates in relation to estuary mouth closing forces (wave energy and sediment availability) and opening forces, such as freshwater flooding and/or storm surges (van Niekerk et al., 2008). Reduced freshwater supply due to drought or impoundments can increase the frequency and duration of mouth closure, and can result in either a decrease or increase in water levels. These modifications to drainage and hydrology can be devastating to salt marshes and can cause loss of the habitat (Laegdsgaard, 2006). Changes in water level and salinity are the primary factors causing modifications in macrophyte occurrence, composition and distribution (Adams, 1994; Riddin and Adams, 2007; Riddin and Adams, 2008a) and the importance of water depth in regulating these parameters has been widely acknowledged (Spence, 1982; Grace, 1987; 1989; Casanova and Brock, 1990; Brock and Casanova, 1991; Papastergiadou and Babalonas, 1992; Casanova, 1994; van der Valk et al., 1994 cited in Vretare, 2002). Alteration of the water regime has implications for the stability of emergent macrophyte communities and the functioning of the estuary (Riddin and Adams, 2008a).

25

Dam construction in the catchment areas of estuaries reduces freshwater input and can increase the frequency and duration of mouth closure in the TOCE. For example, in the Great Brak Estuary (34° 03'S, 22° 14'E), extended mouth closure caused water levels to rise due to localized freshwater run-off, which led to prolonged inundation of the salt marsh (Adams and Bate, 1994a). Dam construction can also reduce the flood requirements of downstream salt marshes. A study investigating the flood requirements of salt marsh plants in the Olifants Estuary (Bornman et al., 2002) was conducted due to a proposed dam construction. The recommendation was to ensure a single release of water, maintained for several days during the spring tidal cycle. This would allow a rise in the water table sufficient enough to cover the supratidal salt marsh with low salinity water and decrease the depth to groundwater in order to prevent the loss of salt marsh.

Inundation is known to reduce the productivity of salt marsh vegetation. Inundation can inhibit leaf growth, stem extension and photosynthesis or promote growth extension, increase senescence and reduce plant productivity (Jackson and Drew, 1984). The elimination of emergent vegetation is often caused by a prolonged increase in water level of more than one year (van der Valk, 1996). The effect of a wide range of inundation conditions on S. tegetaria growth was investigated by Adams and Bate (1994a) and results revealed that weekly stem elongation and branch production was significantly affected by submergence. Completely submerged plants did not grow significantly at ≥35 ppt. Plants showed the best results in saturated substrates maintained by regular tidal inundation, which is typical of their habitat vis a vis the interidal zone and that extended flooding can result in marsh die-back. Submergence of S. tegetaria is also worse during the growing season i.e. spring/summer when temperatures are more conducive to vegetative production (Olff et al., 1988; Adams and Bate, 1994b; van Eck, 2006). Shoot elongation that restores contact with the air is an adaptive response in many flood resistant plants, which can restore or even accelerate growth and can stimulate flowering and seed production (Blom, 1990). When estuaries are closed or tidally blocked, water levels rise as a result of localised freshwater run-off, leading to the inundation of salt marshes for extended periods (Laegdsgaad, 2006). Riddin and Adams (2008) recorded a high water level in the East Kleinemonde Estuary due to an extended period of mouth closure, which caused a significant decline in plant cover after three months of flooding. Experiments by Deegan et al. (2007) discovered that the biomass of two emergent macrophytes was influenced by the amplitude of cyclic water level fluctuations. Typha domingensis biomass declined by 52 % under amplitudes of ±45 cm and the highest biomass recordings in P. australis was recorded under the ±30 cm amplitude treatments. The study by Charpentier et al. (2009) in the Rhone Delta, France, observed that water level was significantly correlated with the cover of Juncus gerardi, with water depths varying from 3 - 17 cm and cover varying from 0 - 25 %. Water level fluctuations may result in the drainage of shallow water bodies or TOCEs, which results in the desiccation of submerged plants (Adams and Riddin, 2008a/b). For example, a 5 hr exposure period is fatal to R. cirrhosa (Adams and Bate, 1994c). As a result, hydrological dynamics can influence biomass fluctuations. Using data from northern Spain, Fernández-Aláez et al. (2002) confirmed the association between charophyte biomass patterns and hydrological dynamics in lakes with strong water-level fluctuations. Similarly, Riddin and Adams (2008a) demonstrated the changes in both emergent and submerged macrophytes in response to fluctuating water levels in the East Kleinemonde Estuary, a small TOCE off the eastern coastline of South Africa.

26 Water depth fluctuates temporally and spatially in wetlands and littoral zones (Rea and Ganf, 1994), which influences the spatial and temporal variation of plant assemblages due to the affects of flooding, erosion, desiccation, sediment stability and turbulence (Chambers et al., 1991, Biggs, 1996; French and Chambers, 1996 cited in Bunn and Arthington, 2002). Spatial and temporal variation in seedling composition of riverplain forests result from the interaction of emergence phenology with the occurrence of environmental disturbance by flooding, and the ability to increase rapidly in numbers during favourable periods (Blom, 1990). Flooding can trigger seed germination of submerged and some emergent species, while waterlogged conditions are suitable for the emergence of both (Boedeltje et al., 2002). Changes in rates of water level fluctuation and disturbance frequency and intensity can affect seedling survival, as well as plant growth rates (Sand-Jensen and Madsen, 1992; Froend and McComb, 1994; Rea and Ganf, 1994). As a result, seasonal maximum water depth and the predictability, duration and frequency of the high water levels affect emergent macrophytes (Ostendorp, 1991). These parameters affect emergent macrophytes differently at different stages of the species life-cycle, for example seed germination, seedling establishment and vegetative and sexual reproduction. Seed germination of Salicornia bigelovii was promoted during autumn inundation and subsequent rapid expansion into bare areas during the winter and following spring (Alexander and Dunton, 2002). These findings demonstrated the importance of the timing and quantity of freshwater inundation in dictating the response of halophytes to precipitation and inflow. Arthrocnemum, Suaeda, Triglochin, Juncus and Casuarina were all able to germinate in waterlogged conditions, but submergence by 5 cm of water retarded and reduced germination (Clarke and Hannon, 1970). Similarly, Britton and Brock (1994) indicated the importance of season of inundation and its affect on the germination of wetland species, with highest germination and species diversity during spring and autumn, and the lowest germination during summer. In contrast, submerged species require inundation for germination to take place. Ruppia seeds will only germinate when covered with water (Kantrud, 1991) and under high water levels they can rapidly complete their life-cycles but will dessicate and die when exposed for a very brief period (Adams and Bate, 1994b/c; Riddin and Adams, 2008a).

Net photosynthesis and oxygen supply to below ground parts is a direct consequence of water depth (Vretare, 2002). The indirect influences are changes to the sediment characteristics, such as redox potential, nutrient status and organic matter content (Grace, 1988; Vretare, 2002). Morphological plasticity within emergent plants, such as biomass allocation to different parts of the plant, changes in stem height and diameter (Lieffers and Shay, 1981; Stevenson and Lee, 1987; Coops et al., 1996; Blanch et al., 1999; Squires and van der Valk, 1992 cited in Vretare, 2002; Vretare, 2002) is therefore critical for adapting to fluctuating water levels. Clevering and Hundscheid (1998) demonstrated both plastic and non-plastic responses of Bolboschoenus maritimus to changing water depths. The proportional allocation of dry matter to stems increased at the expense of roots and rhizome spacers with increasing water depth. Morphological plasticity and biomass allocation of ramets allowed a fast emergence from the water. Vretare et al. (2001) demonstrated phenotypic plasticity in response to increased water depth in P. australis as the plant allocated more resources to stem weight by producing less but taller stems.

In less predictable, drier climates the depth, duration and season (month) of flooding affects the germination and establishment of wetland plants, including the completion of life-cycles through to sexual or asexual reproduction (Casanova and Brock, 2000; Warwick and Brock, 2003), which may be said for TOCEs as well.

27 Experiments by Warwick and Brock (2003), which follow the general idea of Casanova and Brock (2000), found that most of the wetland species that germinated and established during summer were able to flower and set seed. In comparison, those that grew during autumn flowered after 16 weeks with biomass accumulation significantly reduced. For submerged plants, flooding duration must be long enough for reproductive organs to develop and seeds to mature. Amphibious plants, in particular those that adapt morphologically to the changing water levels, developed inflorescences under a wide range of experimental conditions, namely submerged, damp or a portion of the life-cycle submerged. They also found a significant correlation between plant biomass and the number of reproductive units, which demonstrates that seed production is often a function of biomass production in wetland species. Bliss and Zedler (1998) have also demonstrated similar patterns in that the duration of inundation determined species diversity and density of seedlings germinating from vernal pool seed banks. Similarly, Stockey and Hunt (1992) found that the germination of wetland seedlings was dramatically higher in fluctuating water level conditions compared to constant wet or flooded conditions. The germination of many wetland plants is also determined by the interactions between the duration of soil saturation (Kuhn and Zedler, 1997) or inundation (Baldwin et al., 1996) and soil salinity (Noe and Zedler, 2002).

Many plants are able to avoid the adverse effects of submergence through the timing of important life-cycle events. For example the annual Chenopodium rubrum completes its life-cycle rapidly during the short period of water recession between two floods, producing seeds that will survive the next flood (Blom, 1999). In ephemeral habitats, annual Ruppia taxa are dependent on high fecundity to increase their probability of survival. Their reproductive strategy allows for rapid development, early maturity, and the allocation of energy into numerous small propagules when inundated (Kantrud, 1991). Studies by Froend and McComb (2003) observed changes in the phenology of Baumea articulata and Typha orientalis, two wetland species, due to changes in water level. Inflorescence density and standing biomass varied along the water gradient with peak values generally occurring at intermediate water depths. There was also a shift in phenology (flowering, seed production and new leaf growth) with increasing mean water depth and nutrient status. Charpentier et al. (2009) found that at water depths of 0 - 5 cm Juncus gerardi produced more flowering shoots (127 - 121 respectively) compared to 10 - 20 cm water depths (47 - 67 respectively). The height of vegetative and flowering shoots was also significantly taller in the deeper waters. Capsule number per inflorescence was not significantly affected, but fruit maturation was influenced by increasing water depths (0-20 cm). If water levels are too high and Salicornieae plants are submerged, the flowering period can be significantly reduced or favourable flowering conditions may even occur outside of the normal flowering cycle (O‘Callaghan, 1992). Clarke and Hannon (1970) found that saturated sediments and waterlogging slightly above the sediment surface hastened the onset of flowering and produced no adverse effects in S. virginicus. In contrast, Deegan and Harrington (2004) found that seed output was greatly reduced in Schoenoplectus triqueter in the Shannon Estuary in part due to inundation. Further, long periods of inundation (12 hrs) significantly reduced seed production compared to shorter periods (6 hrs) in an experimental study (Deegan, 2000 cited in Deegan and Hannon, 2004). For submerged species it has been demonstrated that the decline in water levels stimulates oospore production (Kautsky, 1990; Casanova and Brock, 1996; Asaeda et al., 2007) and that the high abundance of sexual propagules are probably due to persistent seed bank reserves brought about by adverse environmental conditions, such as water level fluctuations, which are typical of TOCEs (Riddin and Adams, 2008a).

28 The frequency of inundation in TOCEs can also influence macrophyte phenology as it can change suddenly and dramatically as a result of drought, increased rainfall, overwash events or dam construction. These fluctuations can occur for short or long periods. In comparison, POEs experience predictable daily tidal inundation in the intertidal zone. Lenssen et al. (2004) demonstrated the affect of inundation frequency on the anatomical and eco-physiological traits of Ranunculus reptans. The frequency of inundation may also influence the balance between sexual and vegetative reproduction (Mony et al., 2010). Desclaux and Roumet (1996) observed that plant emergence initiated flowering by triggering a signal that causes bud formation. Consequently, increased inundation frequency and duration may delay flowering and represent an escape strategy under such stressful growing conditions (Levitt, 1980 cited in Mony et al., 2010). In the Netherlands, the flowering of two perennial Rumex species was postponed to later in the season or even to the following year (van der Sman et al., 1993) when conditions were more favourable, while the submerged macrophyte, Ruppia cirrhosa, is known to change from a sexual to an asexual reproductive strategy during long periods of inundation i.e. favourable conditions (Gesti et al., 2005). Reserach by van der Sman et al. (1993) found that duration, time and frequency of inundation substantially influenced growth, onset of flowering and seed production in Rumex maritimus and Chenopodium album, two salt marsh and riverine floodplain species. R. maritimus, for example, produced incompletely ripened fruit and C. album produced very little seed due to flooding.

Freshwater wetlands and riverine floodplains are dynamic habitats (Dawson, 1988; Wiegleb, 1988; Casanova, 1994; Casanova and Brock, 2000), and can be equated to TOCEs due to the fluctuating water levels and associated physico-chemical conditions. These fluctuations can occur unpredictably from high water level during closed mouth conditions compared to low water levels and exposure during open mouth conditions after a breaching event. The presence and abundance of plants can change in response to disturbance or fluctuating environmental conditions. The ability to colonize bare space and to re-establish after local extirpation is critical for the long-term persistence of plants in such dynamic habitats (Keddy and Reznicek, 1982; Welling et al., 1988). Capers (2003) found that submerged plants in Whalebone Cove, a tidal freshwater wetland off the Connecticut River, relied primarily on vegetative propagules and vegetative growth to colonize newly available space. However, sexual reproduction complemented vegetative reproduction for Potamogeton pusillus, Elodea nuttallii and Vallisneria americana Michx. These species were most abundant in the seed bank and among the most abundant in the community each year, which suggests that using multiple reproductive strategies may provide a competitive advantage. Every submerged species, apart from the annual Nitella flexilis L., was found to use two colonization strategies (Capers, 2003). Similarly, Myriophyllum variifolium J. Hooker grows in both temporary and permanent habitats in a shallow Australian lake. Plants in both environments are different in growth form, yet produce both asexual and sexual propagules. This flexibility of life-cycle pattern and plasticity of growth form boosts survival in these widely fluctuating environments (Brock, 1991). According to Davis and Stevenson (2007) it is a reproductive and survival strategy of saline lake macrophytes to produce abundant resistant seeds and oogonia within as little as a six week flooding period. Where macrophytes are under less stress, such as in fresher and more permanent systems, they invest less in reproduction and preservation. Similarly, Casanova (1994) found that with decreased water depth sexual reproduction was stimulated and more female than male Chara australis shoots developed while growth increased with increasing water depth. This confirmed their morphological and reproductive plasticity in response to water level fluctuations.

29 2.3.4 Tidal exchange Tidal exchange is very important in the renewal of water within an estuary, as it removes potentially stagnant water and provides input of sea water, nutrients, and sediment, which are important for the ecological health of estuaries. The varied and diverse nature of salt marshes is strongly influenced by the tidal regime because it affects the physico-chemical factors of the estuarine environment, for example salinity, redox potential, and nutrient concentration (Silva, 2006; Whitfield and Bate, 2007). For example, the extent of tidal flushing is important in determining how much of the inorganic and organic nutrients are released from salt marshes to the water column (Riddin and Adams, 2008a). One of the major factors affecting phenotypic plasticity in salt marsh plants is their reaction to this fluctuating environment (Silva, 2006). Increasing tidal flow increases depth and duration of inundation of salt marshes and has a major impact on species composition (Long and Mason, 1983; Greenwood, 2008). Tidal range and tidal exchange is therefore important in the population biology of halophytes and salt marsh morphology (Long and Mason, 1983; Silvestri, 2005; Silva, 2006).

The South African coast is semidiurnal (12.42 hr period) due to the lunar influence and diurnal (24 hr cycle) due to the sun‘s influence, which generates succeeding tides of variable amplitudes. Tidal height variation is limited due to the relatively straight South African coastline, with a height of ~1.0 m to ~0.25 m above mean sea level (amsl) during spring and neap high tides (Whitfield and Bate, 2007). Tidal amplitude is influenced by the state of the mouth in TOCEs and is frequently reduced due to the constricting effect of the mouth. When the mouth is open, tidal variation can either be a few centimetres in a small ‗perched‘ estuary or >1.5 m after a breaching or flood event. Although no tidal variation occurs during the closed mouth state, water level in the estuary may still fluctuate due to river inflow, overwash events, evaporation and/or variations in the berm height (van Niekerk et al., 2002). During high flow periods, the average tidal amplitude in POEs varies between ~1.0 m (spring tide) and ~0.3 m (neap tide), but this can be substantially reduced during low flow periods (Schumann et al., 1999; Whitfield and Bate, 2007).

South Africa is a wave-dominated coast due to the characteristically low tidal ranges and high wave energy (Cooper, 2001). The approximately 250 functional estuaries are predominantly microtidal systems and due to the strong wave action and high sediment availability, more than 90 % have restricted tidal inlets (Whitfield, 1992), which reduces tidal exchange. In larger TOCEs (>150 ha), tidal flow usually maintains an open mouth state after run-off decreases during the low flow season. POEs, estuarine bays and some river mouths are examples of such systems. Tidal exchange is largely a function of the state of the tide and mouth dimensions i.e. width and depth (Taljaard et al., 2009). Tidal exchange is greater in POEs compared to TOCEs, where tidal exchange is often absent due to the presence of a sand bar at the mouth. The tidal prism is partly responsible for the mouth condition of TOCEs (Reddering, 1988; Cooper, 2001), while large POEs are able to maintain an open mouth due to strong tidal forces and a large tidal prism (Whitfield, 1992). Periods of high river inflow, however, reduce the tidal effect and tidal exchange within these estuaries. In medium sized TOCEs (<150 ha), tidal flow usually maintains open mouth conditions during spring tides but frequently the mouth closes during neap tides, e.g. Great Brak and Seekoei (CSIR, 2003; DWAF, 2005). These systems have a small tidal prism which is unable to maintain open mouth conditions during periods of low river inflow (Whitfield and Bate, 2007). This has obvious consequences for the water regime and spatio-temporal distrubtion of macrophytes in TOCEs.

30

Tides are a significant factor influencing the health and functioning of salt marshes (Congdon and McComb, 1980; Adam, 1990; Clarke and Jacoby, 1994; Zedler et al., 2001; Adam, 2002; Greenwood, 2008). Seawater is forced into the estuary during the incoming (flood) tides and raises the water level. During the outflowing tides, water levels decrease more rapidly at the mouth than higher up the estuary, which can result in strong out flowing currents (Breen and McKenzie, 2001) that can scour the estuary mouth, maintaining an open state.

Regular tidal exchange, such as in POEs, generates well developed salt marshes, although they are present in some TOCEs. Salt marsh zonation is particularly well developed in estuaries with high tidal ranges, such as the Kowie and Bushmans estuaries, because plants are distributed in distinct zones along the tidal inundation and elevation gradient (Adams, 1991; Davy, 2000; Rogel et al., 2000; Rogel et al., 2001; Bockelmann et al., 2002; Costa et al., 2003; Ursino et al., 2004). The Bushmans Estuary, for example shows distinct zonation, with Spartina maritima in the intertidal zone and Zostera capensis at the low water mark. Sarcocornia species grow above the Spartina zone and are followed by Triglochin maritima. Atriplex sp. and Bassia diffusa usually indicate the upper limits of the salt marsh area (Adams, 1994; Adams et al., 1999). Spartina maritima only grows in POEs where there is sufficient tidal exchange (Adams and Bate, 1994b). Where there is little tidal affect, such as in TOCEs, plants occur in mosaics rather than distinct zonal bands (Adams and Riddin, 2007). When tidal exchange is restricted, the intertidal zonation patterns can be restricted (Jackson and Drew, 1984). Two types of salt marsh communities were identified by O‘Callaghan (1990) in the Cape estuaries of South Africa. Firstly, salt marsh that develops where tidal exchange predominates, namely in POEs; and secondly, those that develop where estuaries are predominantly closed, namely TOCEs.

During open and tidal conditions in the East Kleinemonde Estuary, a small South African TOCE, intertidal and supratidal salt marsh increased at a maximum monthly expansion rate of 25 and 33 % respectively (Riddin and Adams, 2008a). The tidal amplitude was elevated from 0.4 to 1.2 m and turbidity was high (up to 84 NTU) compared to the closed mouth conditions (up to 26 NTU). The high turbidity was due to the re-suspension of bottom sediments during tidal exchange. In contrast, during closed and non-tidal conditions the macrophyte cover can change dramatically and significantly decline due to high water level and high salinity. The intertidal species, Sarcocornia tegetaria, for example, prefers tidal inundation because it is adapted to saturated substrates maintained by regular tidal inundation (Adams and Bate, 1994a) and occurs relatively low along the elevation gradient of tidal salt marshes (Davy, 2001).

Tidal exchange can be inhibited by natural (e.g. a rocky sill, sand berm) or anthropogenic structures (e.g. piers, dams) (Greenwood, 2008; Ritter et al., 2008). One of the most prevalent anthropogenic alterations of estuarine ecosystems is restriction of tidal and freshwater exchange due to the construction of water control structures (Kennish, 2002). Tidal restriction reduces the scouring action of the tides and can shorten the duration that the mouth remains open in TOCEs, or increases sedimentation at the mouth of POEs. Flood mitigation structures in the Tomago Wetlands, Australia, were constructed to protect agricultural lands from inundation by tidal exchange. As a result, the tidal restrictions have degraded and reduced the functionality of the salt marsh (Hughes, 1998; MacDonald, 2001 cited in Greenwood, 2008). Zedler et al. (2001) concluded that biodiversity in the Los Penasquitos and San Elijos lagoons, California, had declined due to impaired tidal flow and extended

31 mouth closure. Both wetlands had their tidal flows restricted by a railroad bridge and, at their tidal inlets, a coastal highway. Further, those estuaries with a long history of good tidal flow had high species richness compared to those that had been closed to tidal flow for one or more lengthy periods. Reinstating the hydrological regime through restoration projects increases the velocity and low tide drainage, which may also dissipate floodwaters and reduce the impacts of future floods (Streever et al., 1996; Greenwood, 2008). Salinity increases occur with tidal exchange and a functional salt marsh can be reinstated (Streever and Genders, 1997; Greenwood, 2008).

A study by Ritter et al. (2008) found that water quality characteristics varied considerably with tidal restriction and could substantially influence patterns of species presence or absence. Water quality upstream of water control systems frequently has lower salinity, higher temperature and higher nutrients and suspended heavy metals concentrations (Sanzone and McElroy, 1998; Giannico and Souder, 2004 cited in Ritter et al., 2008). Habitat structure can also be modified, for example, tidal restriction reduced the availability of intertidal mudflats (Ritter et al., 2008). Restricted tidal exchange sites are also known to exhibit considerable daily fluctuations in salinity, temperature and dissolved oxygen (John Haskins, Elkhorn Slough National Estuarine Research Reserve, unpublished data cited in Ritter et al., 2008), including severe diel biogeochemical cycling (Beck and Bruland, 2000). The results from Ritter et al. (2008) suggest that a mosaic of tidal exchange regimes would improve total estuary-wide biodiversity.

Sedimentation due to tidal restriction is also elevated which increases the elevation of the salt marsh (Zedler et al., 2001). In the Tijuana Estuary, both sedimentation and salinity increases have occurred since the 16 years of mouth closure and tidal restriction, creating higher and drier marsh areas. This could be fatal to intertidal species sensitive to 5 - 10 cm differences in elevation and higher salinity. Removal of sediment build up can occur naturally during high rainfall periods. For example, a flood event in the Maitlands Estuary, South Africa, lead to improved tidal exchange across the sand bar which resulted in higher than average salinity and strong horizontal and vertical salinity gradients in the estuary (Whitfield and Bate, 2007). Tidal restriction in the - Haringvliet, the Netherlands, modified the estuary from brackish (2.5-3%o CI ), with a mean tidal range of 2 m, - to a semi-stagnant freshwater lake (<0.3%o CI ), with a mean pseudo-tidal range of 30 cm (Kuijpers, 1976; Ferguson and Wolff, 1983 cited in Clevering and van Gulik, 1997). Further hydrological changes included reduced sedimentation, a significant reduction in the extent of the intertidal mud flats and erosion of riverbanks due to the concentration of wave action. Previously, extensive mud flats and large stands of Scirpus lacustris L. ssp. tabernaemontani (C.C. Gmelin) Syme, Scirpus maritimus L. and Phragmites australis grew. After closure, the Scirpus dominated wetlands were almost completely removed and their extent reduced from approximately 500 ha to <1 ha in 1988 (Clevering and van Gulik, 1997).

In the Changjiang Estuary, east China, diking has partially restricted the tidal flow and changed the tidal fluid dynamics (Sun et al., 2003). Vegetation changed dramatically in relation to the change in elevation. Previously there was a distinct zonation pattern of Phragmites australis, Scirpus mariqueter (Tang et Wang) and Scirpus tabernaemontani (Gmel.). P. australis was dominant in the upper marsh, S. mariqueter was dominant at the low and medium elevations and S. tabernaemontani grew in a narrow band between the two. Subsequent to the tidal restriction, S. tabernaemontani invaded the S. mariqueter stands obscuring the zonation that was

32 previously apparent. The restoration of diked or impounded wetlands frequently occurs when tidal flow restrictions are removed and saltwater is re-introduced (Chambers et al., 1998).

Due to the mouth condition tidal exchange and movement can therefore vary in TOCEs compared with POEs where tidal exchange is daily and regular or cyclical in nature. Where tidal exchange is prevented or limited in TOCEs, for example during the closed mouth phase, the occurrence and distribution of macrophytes may be modified. During the open mouth condition, tidal exchange should encourage the growth of supratidal and intertidal salt marsh, whereas during the closed mouth condition the extent of the salt marsh vegetation may be restricted. In comparison, the regular tidal exchange in POEs is a constant that influences and maintains a well developed salt marsh.

2.3.5 Redox potential Redox potential is a measure of the tendency of chemical substances to lose or gain electrons and, as a consequence, be oxidized or reduced. Increasing tidal flow increases the depth and duration of flooding, including changing sedimentation, biochemical interactions and short-term nutrient cycling (Greenwood, 2008). When sediments are flooded, the flood layer restricts the movement of oxygen and other gases from the atmosphere to the sediment. The dissolved oxygen is rapidly consumed by plant roots and microbial decomposition of organic matter causes the sediment to become anaerobic (de Laune et al., 1987; Vaughan et al., 1996; Kyuma, 2004). Waterlogged conditions therefore cause soil oxygen depletion, restricting oxygen transport to roots; and impair root function in the low-redox sediment environment. Alternate electron acceptors become reduced and redox potential decreases causing potentially toxic compounds to accumulate and alter the sediment conditions (de Laune et al., 1987; Greenwood, 2008). Temperature, microbial activity, and the presence of various electron acceptors, such as oxygen, nitrate, manganese oxides, and iron oxides influence sediment redox potential (Ponnamperuma, 1972). For example, microbially mediated reactions are slower at cooler temperatures (Atlas and Bartha, 1987; Paul and Clark, 1996 cited in Vaughan et al., 1996), while greater carbon abundance leads to increased microbial activity and therefore more rapid reduction (Vaughan et al., 1996).

On submergence, the redox potential drops from +200 to -300 mV depending on the type of sediment and other factors, although the Eh in the first few millimeters of the topsoil remains high i.e. approximately +300 to +500 mV (Ponnamperuma, 1972). Patrick and Mahapatra (1968) categorised soil redox potential into four ranges, firstly aerated (+700 to +400 mV); secondly moderately reduced (+400 to +200 mV); thirdly reduced (+200 to -100 mV); and lastly highly reduced (-100 to – 300 mV) (cited in Kongchum, 2005).

Salt marsh soils are often anoxic just below the sediment surface due to high sediment organic matter and moisture content, which stimulates microbial activity. Lower intertidal areas and impounded marshes are particularly prone to anaerobic conditions. Redondo–Gomez et al. (2007) have shown increasing sediment redox potential up the marsh due to decreasing tidal submergence, while lower Eh seems to be one of several factors that reduces species richness at lower elevations in salt marshes (Brewer et al., 1997). Due to the anoxic conditions in tidal marshes, sediments are typically high in sulfur, forming sulfides that blacken the

33 sediment and emit an unpleasant odour, which diminishes the growth of macrophytes. Species richness is reduced, frequently to a single tolerant species, across the intertidal elevation range due to sulfides, microbial activity and inundation, particularly where periods of inundation and anoxic conditions are prolonged (Zedler et al., 2003).

With long periods of tidal inundation sulfidic conditions start developing (Long and Mason, 1983; Chambers et al., 1998; Mitsch and Gosselink, 2000), which is toxic to some plants. Sulfide interacts with salinity, with nitrogen uptake declining as salinity rises (Weisner, 1996; Chambers et al., 1998). Flooded organic substrates are generally reduced and may contain reduced phytotoxic substances, such as ferrous irons, sulfides and methane. Another consequence of reduced sediments is the changes in availability and/or concentrations of various nutrients required for macrophytes to function (Vaughan et al., 1996, Pezeshki, 2001). Emergent macrophytes adapt by oxidizing the substrate around their roots and developing an efficient oxygen transport system from their aerial parts to below ground roots (Weisner, 1996). However, in deep water, growth may be inhibited by sediment eutrophication and the consequent damage to root function and efficiency as a result of sediments containing toxic and reduced substances (Weisner, 1996). Factors that may negatively affect the performance of macrophytes at permanently inundated locations includes low redox conditions in the sediment due to a high organic matter content (Brock et al., 1987; van den Brink et al., 1995), and low internal oxygen availability (Yamasaki and Tange, 1981).

The affect of fluctuating redox potential on Spartina alternifolia and its ability to assimilate carbon was investigated by Pezeshki et al. (1989). The results demonstrated that photosynthesis changed very little in relation to a moderate reduction in redox potential but declined significantly at –200 mV. Under extreme anaerobic conditions and rapid Eh changes S. alternifolia is unable to cope despite the extensive system of aerenchyma. Rapid decline of Eh caused by tidal inundation can therefore cause a reduction in carbon fixation. These results were in accordance with another study along the Louisiana coast where growth differences in short and long forms of S. alternifolia was attributed to severely reduced sediments at the short-form site (de Laune et al., 1983).

The tolerance redox potential range can differ among macrophytes. Pezeshki et al. (1996) observed the photosynthetic response of Typha domingensis Pers. and Cladium jamaicense Crantz under a range of redox potential treatments (+500, +250, 0 and -200 mV). Stomatal conductance was significantly reduced at -200 mV and net photosynthesis declined significantly under reduced Eh conditions. At –200 mV, net photosynthesis declined by 71 % in T. domingensis compared to C. jamaicense, in which only respiration occurred. Under 0 and -200 mV soil Eh treatments, both species experienced a decline in total biomass but the reduction was significantly higher in T. domingensis. In a survey of 76 salt marshes in England, Sarcocornia perennis generally occupied sediments that had oxidizing redox potentials which demonstrated their preference for well aerated sediments (Davy, 2001). In south-west Spain at the Odiel marshes, transplant experiments and field measurements demonstrated that S. perennis was intolerant of persistently low sediment redox potentials (Davy, 2001). Castellanos et al. (1994) observed that S. perennis invaded the raised tussocks of Spartina maritima in areas of a lagoon where there was adequate drainage, and the sediments were always positive, ranging from +48 mV to +133 mV. In contrast, it did not invade similar tussocks in the neighbouring lagoon

34 where drainage was gradual and pooling occurred. This was because the parts of the tussocks dominated by Spartina and the tussocks in the ponded lagoon maintained low Eh conditions, ranging between −150 mV and −400 mV (Davy, 2001). S. virginicus, on the other hand, is able to tolerate moderately oxidised (+340 to +364 mV) to moderaly reduced (50 - 100 mV) conditions (Naidoo and Naidoo, 1992; Naidoo and Naidoo, 2000). Redondo-Gómez et al. (2007) showed that the failure of Sarcocornia perennis ssp. alpini to survive below its limit was because of its intolerance of increasingly hypoxic and reducing sediments.

Increasing the depth and duration of flooding can create anoxic and sulfidic conditions for emergent macrophytes. For example, investigations on the affect of Phragmites australis showed that sulfide concentrations >1 mM can cause stunted growth, callus blockage of gas-pathways and bud death (Armstrong et al., 1996). Salt marsh plants shunt oxygen to the roots to maintain aerobic respiration (Gries et al., 1990) under anoxic conditions, but when this mechanism is overwhelmed they shift to fermentative metabolism (Chambers et al., 1998). The activity of alcohol dehydrogenase is blocked under high sulfide concentrations, which inhibits fermentative metabolism, and results in decreased energy production, including reduced nitrogen uptake and plant growth (Koch et al., 1990). The nitrogen uptake in P. australis dropped to 0 under high sulfide treatments (Chambers et al., 1998). Maintenance of a positive oxygen (Weisner and Graneli, 1989) and carbon balance (Cizkova-Koncalova et al., 1992) is therefore key to the success or failure of P. australis under different soil and water regimes.

Submerged macrophytes are well adapted to reduced conditions because they are rooted in anoxic sediments, except for the thin oxygenated surface layer. Adaptations include gaseous exchange over the entire plant surface, nutrient uptake by the above-ground parts and the roots (van Wijk, 1989) and oxygen transport through an intercellular air-space system from above-ground parts to the roots (van Wijk et al., 1992). A study by van Wijk et al. (1992) showed that Potamogeton pectinatus had a limited tolerance of the reduced substances Fe2+ and S2-, as it was still able to grow but growth was generally poor. The results suggested that the relatively low biomass production observed on sediments rich in organic matter was due to anoxic conditions caused by the high quantities of organic matter present. The anoxic or reduced conditions would have allowed for the formation of reduced substances, such as Fe 2+, S2- and H²S, which probably limited the accumulation of biomass due to their phytotoxic nature (de Laune et al., 1983; Pezeshki, 2001). In contrast, Penhale and Wetzel (1983) found no reduction in root respiration of Zostera marina L. at high S2- concentrations. This shows that submerged macrophytes have different tolerance ranges because, in the case of Zostera marina L., an enhanced air-space system develops in reaction to anaerobic or reduced conditions. Further, that tolerance of Fe2+ and S2- under reduced conditions may be an important ecological adaptation, which may influence the concurrence between different submerged and emergent macrophyte species (van Wijk et al., 1992).

Not only does redox potential fluctuate with water depth, but it fluctuates with the seasons as well. Dwire et al. (2006) examined plant species distribution in three riparian plant communities, defined as wet, moist, and dry meadow, along a short topographic gradient. Sediment redox potential differed in the three communities and Eh dynamics were related with the seasonal fluctuations in water-table depth. Sediments were predominantly anaerobic in the wet meadow community, with Eh values ranging from -113 to +436 mV at 10 cm depth during

35 spring and summer; and were highly or moderately reduced at 25 cm depth throughout the year. Sediments were anaerobic during spring high flows and aerobic in summer and autumn during low flows in the moist meadows. In contrast, sediment conditions were principally aerobic at both depths throughout the year in the dry meadow community.

Submerged macrophytes, through their adaptation to aquatic environments, are able to transport oxygen from above-ground parts to the roots through an intercellular air-space system. This allows them to affect the redox potential of the rhizosphere (Sand-Jensen et al., 1982; Kemp and Murray, 1986). Sand-Jensen et al. (1982), for example, found that Zostera marina released oxygen from its roots but at lower levels than two stream species, Sparganium and Pomatogeton species, while three isoetid species had the highest releases. Charophytes, which grow close to the sediment in which they are anchored, may influence redox potential at the sediment/water interface and their growth may distribute oxygen to anaerobic bottom sediments. An oxygen gradient formed in this way may inhibit the release of iron-bound sediment phosphorus and promote nitrification/denitrification nitrogen losses (Lijklema, 1994; Weisner et al., 1994 cited in Kufel and Kufel, 2002) required for macrophyte growth (Kufel and Kufel, 2002). However, the oxygen release from roots may increase the nutrient uptake of other plants by increasing the aerobic generation of nutrients and the efficiency of uptake mechanisms (Morris, 1979 cited in Sand-Jensen et al., 1982).

Most researchers agree that changes in redox potential are a likely trigger in the germination of charophyte oospores (Kalin and Smith, 2005). Sediment anoxia is a prerequisite for oospore germination of charophytes (Forsberg, 1965; Matheson et al., 2005), and high redox potential of the benthic environment caused by photosynthesis of benthic algae causes low germination rates (van den Berg et al., 1999; Asaeda et al., 2007). Factors triggering the germination of C. vulgaris and Nitella flexilis L. were investigated in a series of three experiments by Kalin and Smith (2005). N. flexilis oospores did not germinate under constant redox values but all of the oospores germinated in a tube in which the redox of the agar declined to below 200 mV. Results suggested that both submerged plants were influenced by changing redox potential. Mateos-Naranjo et al. (2008) also demonstrated the importance of redox potential in the germination of Spartina alternifolia seeds because low final germination percentage and delayed germination was associated strongly with the negative redox potentials at the lowest sites in the Gulf of Cadiz salt marsh, Spain.

It is anticipated that sediment redox potential conditions in TOCEs will fluctuate more compared to that in POEs due to changing open or closed mouth conditions, which influence estuary water level and potential periods of protracted submergence of intertidal and supratidal macrophytes. Increased submergence will therefore reduce sediment redox potential. The submerged reductive sediments create greater oxygen stress compared to oxidative sediments because the oxygen demand is far greater. In addition, a combination of low redox potentials with the presence of natural toxins in the sediment may cause an even greater stress (Armstrong, 1967; Ernst, 1990 cited in van den Brink, 1995). Macrophyte phenology will be negatively influenced and die- back is likely to occur because many of the species are not tolerant of protracted inundation (> two months) and highly reduced sediments. Although Bolboschoenus maritimus and Phragmites australis may have a higher tolerance for reduced sediments, prolonged periods of inundation will create toxic conditions that will reduce plant vigour (de Laune et al., 1987; Vaughan et al., 1996, Pezeshki, 2001; Zedler et al., 2003; Greenwood,

36 2008). In comparison, POEs experience daily tidal inundation but it is not prolonged and as a result it is anticipated that redox potential will be less variable and sediments will be well aerated with moderately reduced to highly oxidized sediments (depending on the position along the elevation gradient). In the East Kleindemonde, the water column is expected to be generally well oxygenated, with dissolved oxygen levels not declining below 4 mg L-1. During protracted mouth closure periods, bottom waters (< 1m) may occasionally drop below 2 mg L-1 (van Niekerk et al., 2008)

2.3.6 Sediment organic matter Sediment organic matter is an essential component of macrophyte productivity as it contains all of the essential plant nutrients for growth. It consists of carbon and nutrients in the form of carbohydrates, proteins, fats and nucleic acids. The stable organic component (humus) has both a nutrient-holding capacity as it adsorbs and holds nutrients, and a moisture-holding capacity. Through the process of decomposition and the colloidal nature of humus, nutrients are released in a form available to plants and plant productivity is enhanced (Doerr, 1990; Lickacz and Penny, 2001).

It has been suggested that most South African estuaries lack sufficient sediment organic matter reserves, except perhaps in human impacted estuaries where nutrient enrichment occurs (Taljaard et al., 2009). The justification for this is that during seasonal high flows the smaller, shallow estuaries (e.g. the East Kleinemonde Estuary, a TOCE in the warm temperate region) are effectively flushed of accumulated organic matter. In comparison, the larger, deeper and wider estuaries require much higher river inflows for effective flushing, which results in an accumulation of benthic organic matter. During the closed mouth state however, the breakdown of organic matter is still important with regards to supporting macrophyte growth. During floods it is also expected that sediments and benthic particulate organic matter is re-suspended in the estuary through scouring, and subsequently transported onto the nearshore coastal shelf thereby reducing the organic matter stock (Taljaard, et al., 2009) in the estuary. Die-back of macrophytes during extended periods of high level inundation may subsequently become a significant source of organic matter to the estuary when tidal exchange is re-established, or when the system is flushed (Adams and Bate, 1994a; Adams et al., 1999; Taljaard et al., 2009).

Typical marsh sediment consists of 47 % sand, 27 % silt/clay and 26 % particle matter and organic matter (Greenwood, 2008). Organic matter is produced by numerous primary producers either connected to the sediment, such as benthic microalgae, macroalgae and rooted macrophytes, or in the pelagic areas, such as phytoplankton (Flindt et al., 1999). Sediment organic matter is therefore derived from bacteria or plankton and plant and animal detritus formed in situ and can be the result of macrophyte die-back due to elevated water levels (Adams and Bate, 1994b; Adams et al., 1999; Taljaard, et al., 2009). Clarke and Jacoby (1994) suggest that Juncus kraussii productivity appears to accumulate in the upper salt-marsh due to the higher organic matter and nutrient concentrations in the sediments compared with those in the lower salt marsh zone. Consequently, organic matter and nutrients in the upper or suptridal zone may provide a source to the lower intertidal zone where detritus depletion occurs due to microbial activity and frequent tidal flushing (Clarke 1986 cited in Clarke and Jacoby, 1994). This is reflected in the sediments occupied by Sarcocornia perennis where the organic matter content is generally low (Davy, 2006). Spartina alternifolia productivity is also known to

37 accumulate in sediments, although the proportion of organic material is less at lower latitudes (Bertness, 1999 cited in Adam, 1990). The accumulation of large amounts of organic matter due to annual decomposition of old reeds is common in stands of P. australis (Greenwood, 2008). According to Allanson and Baird (1999), during high spring tides, detrital plant material originating from intertidal salt marsh areas is deposited in the supratidal areas, which demonstrates the movement of organic matter or nutrients in an estuary, as proposed by Clarke and Jacoby (1994). Sediment organic matter is therefore a major source of nutrient supply for macrophytes (Flindt et al., 1999) and increases the water holding capacity of substrates, while low organic matter can lead to nutrient deficient sediments and low productivity (van Wijck, 1980; Mitsch and Gosselink, 2000; Bai et al., 2005). However, high organic matter may cause anoxic conditions, especially during the closed mouth condition of a TOCE (Adams and Riddin, 2007).

Salt marshes differ with regard to the tidal exchange of nutrients (Jordan, 1983). For example, the Great Sippewissett Marsh, Masachusettes, exports carbon and particulate nitrogen (Valiela et al., 1978), Flax Pond Marsh, New York, imports particulate organic carbon (Woodwell et al., 1977) while Carter Creek Marsh, Virginia, imports particulate nitrogen (Axelrad et al., 1976 cited in Jordan et al., 1983). Jordan et al. (1983) measured approximately 75 % organic matter in the Rhode River high marsh sediment, Chesapeake Bay, and that 7 % of the organic carbon produced in the high marsh is exported while 33 % accumulates in the marsh. Fluctuations of river inflow can influence the movement of organic matter into the estuary from the terrestrial landscape and thus some estuaries are dependant on locally produced plant material and marine imports (Kokkinn and Allanson, 1985 cited in Adams et al., 1999). In the Kowie Estuary, a temperate POE along the Eastern Cape coastline of South Africa, Kokkinn and Allanson (1985) found that the organic carbon produced by the salt marsh plants remained in the estuary and was utilized by the organisms living there. Consequently the amount of organic carbon import from the marine environment was substantially higher compared to that which was exported from the estuary (cited in Adams et al., 1999).

Sediment nutrients are linked to the surrounding water through submersed rooted macrophytes and this has potentially important implications for nutrient cycling within estuaries (Flindt et al., 1999). A study by Flindt (1994) compared the nutrient dynamics in bare bottom sediment versus eelgrass covered sediment. Results revealed that rooted macrophytes, Zostera marina, potentially enhanced denitrification thereby minimising the releases of ammonia to the water column, which reduce competition from macroalgae and phytoplankton. Low phosphate fluxes are also ensured through the superior phosphate adsorption capacity in the rhizosphere. Seagrasses are able to modify seasonal nutrient cycling by storing nitrogen during the growing season and then reallocating nutrients from old to new leaves before dying. In so doing, seagrasses keep nutrient dynamics very close to the rhizosphere promoting nutrient availability for macrophyte productivity (Flindt et al., 1999). Similarly, charophytes act as nutrient sinks by incorporating them in plant biomass, delivering oxygen to the reduced sediment/water interface creating favourable conditions for nitrification/denitrification (Lijklema, 1994) and form dense stands that restrict the return of nutrients already stored in bottom sediments (Kufel and Kufel, 2001).

The positive effect of organic matter on plant growth has been observed in submerged sediments (Sand- Jensen and Sondergaard, 1979; Kiorboe, 1980; Lillie and Barko, 1990). Growth was probably due to the

38 release of nitrogen and/or phosphorus during the mineralization of the organic matter (Barko et al., 1986). In the Ringkobing fjord, Denmark, total biomass of Ruppia cirrhosa, Chara globularis, Potamogeton pecinatus L., Myriophyllum spicatum and Ranunculus baudauti increased with increased organic matter, nitrogen, carbon, silt and clay content of the sediment substrate (Kiorboe, 1980). A study by Barko and Smart (1986) corroborated this by suggesting that the lack of multiple nutrient availability caused a decline in growth. Further, that diminished growth on sands was probably due to nutritional inadequacies because of the inherent infertility of this sediment type (Barko and Smart, 1983; 1986), because sand is generally low in organic matter content.

Field experiments on the growth of Schoenoplectus lacustris ssp. lacustris and Bolboschoenus maritimus was conducted at different sites in the Hargvliet Estuary, Netherlands, in fertilised and non-fertilised plots (Clevering and van Gulik, 1997). The fertilised plots were fertilized using pellets of 7.5 g of a slow release fertiliser per plant (Osmocote 17 % N, 1.6 % P205, and 8.7 % K, active for 9 months at 21°C), which were buried in the sediment together with the rhizomes or tubers. Growth was monitored during spring over a period of three years. The results showed that fertilisation had no significant effect on the number and dry weight of shoots. They concluded that it was unlikely that differences in soil fertility caused the differences in dry matter production of S. lacustris between sites and that the differences were not associated with differences in sediment nutrients, particle sizes and organic matter content. Van Wijck (1980) investigated the influence of sediment organic matter on the biomass production of Potagometon pectinatus using nine sediments ranging from 8 - 66 mg C g-1. Biomass production increased with increasing sediment organic matter content up to 26 mg C g-1 and was considered to be related to the nitrogen bound to the organic matter. This experiment also showed a significant correlation between sediment organic matter and moisture content. Pennings et al. (2002) demonstrated similar trends, in that different macrophytes respond differently to nutrient inputs thereby modifying plant competitive interactions and community structure. Troyo-Dieguez et al. (1994) found that the growth of Salicornia bigelovii in Baja California Sur, Mexico, was positively correlated with sediment organic matter content.

Aquatic plants have a differential growth response to varying sediments which may influence the species composition of aquatic macrophyte communities (Barko and Smart, 1983, 1986). Barko and Smart (1986) found that Hydrilla verticillata (L.f.) Royle appeared to be more sensitive to sediment composition than Myriophyllum spicatum L. because its root to shoot ratio on infertile sediments was lower, suggesting that the latter species has a competitive advantage where sediments contain low organic matter content or nutrients. This is because plants growing in infertile soils tend to increase the shoot to root ratio in the search for nutrients. As a result, those aquatic plants that can allocate resources to higher root formation should have a competitive advantage. The spatial distribution and species composition of these communities have been correlated to organic matter content in various studies (Pearsal, 1920; Mirsa 1938; Macan, 1977 cited in Barko and Smart, 1986) and is considered to have an influence on the succession of macrophyte vegetation. Modifications in sediment organic matter due to catchment disturbance, mechanical bottom sediment disturbance and autogenic processes can therefore impact aquatic macrophyte composition (Barko and Smart, 1986), including tidal flushing in POEs and the mouth condition of TOCEs.

39 Perez et al. (1994) demonstrated the growth plasticity of the Mediterranean seagrass Cymodocea nodosa (Ucria) Aschers in response to an existing gradient of nutrient supply in shallow water. The plants that grew in areas where phosphorus was limited allocated a greater proportion of the limiting resource to root development. This allowed plants to increase their ability to acquire phosphorus from the sediment at the expense of shoot growth and size. Once the nutrient shortage was ameliorated resources were relocated, which reduced the relative allocation to below-ground biomass thereby reversing the process in response to nutrient availability.

Seed germination and seedling establishment is also influenced by sediment organic matter content. The light- requiring mechanism of some salt marsh species, for example Scirpus robustus and Bolboschoenus maritimus, prevents germination of seeds which have been buried under dead organic matter (Dietert and Schontz, 1978; Wilman, 2006). Due to the moisture-holding capacity of organic matter (Doerr, 1990; Lickacz and Penny, 2001), seed germination and seedling establishment may be enhanced. It has been demonstrated that seedling establishment is encouraged under the canopy of parent plants due to the sediments containing higher quantities of organic matter (Bashan et al., 2000 cited in Fenner and Thompson, 2005), and as a result higher sediment moisture content and nutrients. Plants are thought to have a critical water content requirement for germination to occur (Hunter and Erickson, 1952 cited in Fenner and Thompson, 2005). Water is therefore crucial for germination and studies by Evans and Etherington (1990), for example, showed that none of the wetland species they researched, such as Juncus acutus and Typha latifolia, could germinate under low water potentials. Consequently, a reduction in organic matter content may reduce sediment moisture content, particularly in the suptratidal zone, thereby reducing the ability of seeds to germinate. Jarvis and Moore (2008) found that seed germination in the freshwater submerged macrophyte, Vallisneria americana Michx., was enhanced in sediments composed of ≤3 % organic matter content, while experiments by Hoover (1984) indicated delayed germination in sediments with >6 % organic matter (cited in Jarvis and Moore, 2008).

In TOCEs, the removal of organic matter during the open mouth phase due to floods, or the accumulation of organic matter which could result in anoxic conditions during the closed mouth phase, will therefore influence macrophyte phenology. Estuary mouth dynamics will therefore play a role in influencing organic matter content where a rapid reduction in these reserves may result from a breaching event. The closed mouth phase in contrast, will allow the gradual accumulation of debris and organic matter to take place resulting eventually in anoxic conditions that are detrimental to macrophyte growth. In comparison, POEs will most likely maintain a relatively stable level of organic matter due to the estuary mouth remaining open and the resultant daily tidal exchange.

2.3.7 Sediment moisture content Sediment moisture influences the distribution and growth of plants, soil aeration (mostly oxygen and nitrogen), soil microbial activity and the movement of nutrients from the sediment to the roots. It is that portion of water in the sediment that can be readily absorbed by plants and which carries dissolved nutrients for plant growth (Miller, 1992). Sediment moisture is held in the interstitial pore spaces of sediment particles and the interstitial pore size is dependent on the grain size of the sediment. Large sediment grains, such as in sandy sediments, can accumulate greater volumes of water, but due to their irregular shape the water is easily transferred through a hydraulic gradient (Miller, 1992; Shaw, 2007). The grain size not only affects water movement, but 40 the movement of nutrients dissolved in solution and water retention within the sediment. Clay particles develop a negative charge on their surfaces which holds both water and plant nutrients in the soil (Anderson and Kalff, 1988; Doerr, 1990).

Spatial and temporal variability in sediment moisture content is influenced by several factors, such as sediment properties, vegetation cover, tidal inundation, depth to water table, topography, temperature and rainfall; and these factors can vary locally and seasonally (Gomez-Plaza et al., 2001; Noe and Zedler, 2001; Shaw, 2007). Sediment properties, such as the amount of sediment organic matter and texture, also influence sediment moisture content because organic matter and clay/silt soils have a higher water holding capacity compared to sandy, organically depleted soils that have a high leaching capacity (Barko and Smart, 1986; Gomez-Plaza et al., 2001; Lickacz and Penny, 2001; Bai et al., 2005). Sediment moisture content is therefore often strongly paralleled with sediment organic matter (Barko and Smart, 1986). Although clay particles retain the greatest volumes of moisture compared to other sediment types, the moisture is often not available to plants due to the small particle size of the sediment (Gomez-Plaza et al., 2001).

The sediments of coastal salt marshes are characteristically moist while the groundwater or water table is generally in close proximity to the sediment surface (Bornman et al., 2004). Halophyte distribution is influenced by moisture amongst other environmental factors (Ungar, 1978). The growth of xerohalophytes (plants that have adapted morphologically to arid conditions) is generally limited by sediment moisture content (Ungar et al., 1979; Riehl and Ungar, 1982; Young and Nobel, 1986; Zedler et al., 1986) and in many systems is dependent on the depth to the water table (Richards and Caldwell, 1987; Cantero et al., 1998; Pan et al., 1998; Drabsch et al., 1999). This is because the position of the water table influences the sediment moisture profile above it (Hillel, 1971). Knowledge of the variations in sediment moisture is therefore important when interpreting macrophyte phenology, water use, productivity and seedling establishment, all of which can influence the distribution patterns of salt marsh plants (Bornman et al., 2004). In the Olifants Estuary, along the arid west coast of South Africa, Bornman et al. (2002) demonstrated the importance of groundwater in relation to soil moisture availability for the Sarcocornia pillansii population. The research found that the depth to the water table had the greatest influence on the distribution of S. pillansii. Pan et al. (1998) found that there were two distinct vegetation gradients in the lower alluvial fan of the Hutubi River, China. One was determined by the sediment moisture content (position relative to the water table) and the other determined by a salinity gradient, while the main spatial gradient that structured these halophytic communities was the depth to water table. Environmental gradients are chiefly influenced by sediment moisture and/or flooding regime and/or sediment salinity (Zedler et al., 1999; Álvarez-Rogel et al., 2000; 2001; Bouzillé et al., 2001; Denslow and Battaglia, 2002).

A study by Muir (2000) in the Knysna Estuary, South Africa, demonstrated the spatio-temporal nature of sediment moisture content (Ehrenfeld, 1997; Noe and Zedler, 2001). Moisture content decreased from the intertidal Triglochin / Spartina / Zostera zone to the supratidal habitat because the substrate in intertidal habitat was inundated on a diurnal tidal pattern. Sediment moisture content has been shown to decrease with increasing elevation (Redondo-Gómez et al., 2007). The sediment organic matter content was inversely proportional to the moisture content due to the fact that high moisture content assists in the rapid decay of

41 organic matter. Seasonal changes in moisture content were also evident and moisture content was one of the significant environmental factors affecting plant distribution in autumn, winter and spring. The decline in moisture availability due to a change in the tidal reach was considered to be the possible cause for a significant decline in the cover of Plantago crassifolia from spring to summer (Muir, 2000). Redondo-Gómez et al. (2007) also found that the failure of Sarcorcornia ssp. perennis to survive above its limit was associated with water stress (including hypersalinity) as moisture content decreases with increasing elevation in summer.

Plants can only absorb water from the sediment when their water potential is lower than the water potential in the surrounding environment, which means water in the plant is rarely in equilibrium with sediment water. The difference in water potential between the two depends on the evaporative demand, the water conducting properties of the sediment and the extent to which the plant can meet that demand (Smith and de Laune, 1984). Smith and de Laune (1984) demonstrated that water stress in Spartina alterniflora salt marsh can occur even at a relatively high sediment moisture content due to increased salinity as a result of moisture loss through evapotranspiration. The high substrate salinity in the interstitial water reduced photosynthetic activity through the effects of plant water stress. Although the sediment moisture content remained above saturation throughout the experiment, the osmotic potential was equivalent to being in the range between field capacity and permanent wilting.

Moisture induces seed germination and is a requirement for seedling establishment and plant growth. Experiments by Noe (2002) investigated the influence of varying moisture (and salinity) regimes on seed germination in various estuarine plants. Species responded to even small differences in moisture variations and 10 of the 11 species responded to temporal variations in the duration, amplitude and seasonal timing of high moisture sediments. In conclusion, germination was always greatest in the high moisture treatments, with Cotula and Hutchinsia species the most responsive. Noe and Zedler (2000) also demonstrated the importance of sediment moisture content on seed germination in seven upper intertidal salt marsh species. Three sediment moisture treatments, low (37.1 %), medium (45.5 %) and high (50.5%), were assessed with four salinity treatments, namely 34, 17, 8, and 0 ppt. Higher sediment moisture increased the final proportion of five germinating species including the germination speed of all seven species under investigation. Annual seedlings of species also segregate along spatial gradients of moisture content and salinity (Noe, 1999), while the timing variance associated with establishment is most likely due to temporal differences in soil moisture, soil salinity, photoperiod, or temperature and differential responses of species to these abiotic factors. The increase in soil moisture and the lowering of soil salinity are probably the most critical amongst these environmental factors in the timing of germination (Bornman, 2002; Noe and Zedler, 2001). Mateos-Naranjo et al. (2008) also demonstrated the importance of moisture content in the germination of S. alternifolia seeds because no germination occurred at the highest elevation in the Gulf of Cadiz salt marsh, Spain, where sediment moisture content was lowest.

Sediment moisture is important because it influences the spatial variability in sediment salinity (Bush, 2006; Noe and Zedler, 2001). Bolen (1964) discovered such variability in an inland salt marsh of Utah. The study indicated that during periods of low sediment moisture content, the surface salinity was far greater compared to the salinity deeper in the soil profile. Smaller salinity variations also occurred with depth in the wetter areas of

42 the salt marsh. Soils with low sediment moisture were approximately four times more saline in the upper 15 cm and about twice as saline in the 15 – 30 cm sampling level depth compared to soils with higher sediment moisture content. Similarly, a study by Bush (2006) in the Chihuahuan Desert salt marshes of North America, found that soils with low sediment moisture (~40 % versus ~60 % sediment moisture) had far greater surface salinity readings (~25 g kg-1 versus ~11 g kg-1 sediment salinity). Further, although salinity was the most significant factor influencing the growth of the endemic salt marsh plant Helianthus paradoxus early in the growing season, sediment moisture was the most important later in the growing season. Bush (2006) concluded that the removal and redistribution of salts in the soil profile due to sediment moisture was probably the reason for this. The fact that these plants tend to grow closer to the drainage channels during dry years compared to wet years may support this conclusion. Further, during the years with average precipitation, H. paradoxus germinates when sediment moisture is elevated and surface soil salinity is lowered in the winter months. It is therefore anticipated that during droughts or periods of low rainfall and high temperatures (typical of the study region), sediment moisture content will be reduced in both POEs and TOCEs. Consequently, sediment salinity will increase, particularly in the supratidal zone where inundation occurs infrequently in POEs or not at all in TOCEs depending on water level. These conditions may escalate sediment salinity to a point where macrophyte phenology is negatively influenced in both systems, as discussed in Section 2.3.2.

Sediment moisture content is important at different stages of macrophyte life-cycles, for example germination and seedling establishment; and the temporal changes in sediment moisture are important in influencing germination periods (Ehrenfeld et al., 1997; Noe and Zedler, 2001). Shaw (2007) found that moisture availability was more important at high salinity during seed germination because a higher number of seeds germinated under the inundation treatment compared to moist conditions only (Noe and Zedler, 2001). The experiments included three water regimes i.e. damp, waterlogged and inundated (with seeds floating in the water) and four salinity treatments i.e. 0, 15, 35 and 70 ppt (Shaw et al., 2008). Seeds germinate rapidly under freshwater when moisture requirements are met but the seedlings also require another two to three weeks of additional moisture to establish (Kuhn and Zedler, 1997; Alexander and Dunton, 2002). Adult plants can assist in seedling establishment by shading the sediment surface thereby reducing evaporation and retaining the sediment moisture for longer (Bertness, 1991; Callaway, 1994; Callaway and Subraw, 1994; Noe and Zedler, 2001; Shaw, 2007). They then require a reliable source of moisture to grow and mature, as do the adult plants but they are generally more tolerant of water stress and can persist through periods of drought and hypersalinity (Ungar, 1978; Fenner, 1985).

Sediment moisture content is therefore critical for seed germination, seedling establishment and macrophyte growth particularly so with reference to its influence on salinity levels and redox potential. Sediment moisture content in TOCEs is anticipated to be more important when water level is low and larger areas of supratidal and intertidal areas are exposed. This is because the germination of persistent seed banks will be encouraged if salinity is lowered by higher sediment moisture content induced by rainfall (Ungar, 1978; Ungar, 2001; Rubio- Casal et al., 2003; Bornman et al., 2002; 2004). However, the lack of rainfall, freshwater inflows and drought will be detrimental in this regard. In POEs, the same will apply as both supratidal and intertidal habitat is anticipated to remain comparatively stable in extent (m²). Reduced sediment moisture content and the resultant high sediment salinity in both these habitats, will also create stress if the ideal salinity range for macrophytes is surpassed. In contrast, during the closed mouth phase of TOCEs, protracted periods of inundation will result in

43 high sediment moisture content and therefore highly reduced sediments and low oxygen content which will negatively influence macrophyte phenology (Section 2.3.5).

2.3.8 pH Soil reaction, which is expressed as pH, refers to the acidity or alkalinity of water, and relates to the relative proportion of hydrogen (H+) and hydroxide (OH-) ions. Hydrogen ions predominate in waters with pH <7 (acidic) and hydroxyl ions predominate in waters with pH >7 (alkaline). In acidic environments, the major reaction influencing pH is the oxidation and reduction of iron. Most natural freshwaters have pH values ranging from 6.5 to 8.0, while marine water is between approximately 7.9 and 8.2 (DWAF, 1995). During periods of inundation, hydrogen is consumed by the reduction of iron resulting in an increase in pH. Once the water retreats, the system is oxidized and the pH will revert to the original value (Doerr, 1990; Millar, 1992; Bai, 2005; Holley, 2009; Radke, 2010). pH has also been shown to change in response to variations in salinity (Baldwin and Mendelssohn, 1998; Al-Busaidi and Cookson, 2003). For example in alkaline sediments, pH usually increases with an increase in salinity while in less alkaline sediments pH decreases with a salinity increase (Gupta et al., 1989; Lai and Stewart, 1990 cited in Al-Busaidi and Cookson, 2003).

The possible physiological consequences of acidic conditions are well documented in terrestrial plants and range from limited nutrient availability, hydrogen ion toxicity and toxicity of metals more soluble at higher pH levels, while submerged plants may experience the added reduction in the availability of photosynthetic carbon sources (Roelofs et al., 1984; Wetzel et al., 1985; Titus et al., 1990). At sediment pH levels below 5.5 or acidic conditions, microbial activity is lowered (Doerr, 1990; Millar, 1992; Bai, 2005; Holley, 2009; Radke, 2010). For example, Bai et al. (2005) observed that the contents and spatial distributions of sediment organic matter and total nitrogen varied greatly in wetland soils where the sediment pH values ranged from 8.1 to 9.6 due to its influence on microbial activity. Soil reaction (pH) has a direct influence on membrane function, cell regulation, respiration and ion solubility (Linthurst and Seneca, 1980; Larcher, 1995). The sediment pH affects the uptake of nutrients by plants indirectly through the solubility of ions and the activity of micro-organisms. The availability of several nutrients, especially phosphorous and the micronutrients, are affected by pH and become less available as it escalates. In submerged habitats, an increase in pH causes a decrease in CO2 concentrations, which is detrimental to photosynthesis. As a result, the photosynthetic capacity of submerged macrophytes becomes dependent on the use of HCO3; the efficiency in the use of this source of dissolved inorganic carbon varies greatly from one plant group to another resulting in a similarly diverse pattern of response to increasing pH (Invers et al., 1997).

Estuarine pH is influenced by both the river and the sea waters which flow into them. River water pH is generally a function of catchment characteristics, for example rivers draining Table Mountain quartzite are rich in humic acids and characterised by low pH levels (~4). Once the river water comes into contact with the sea water (pH of 7 to 8.5), the pH in estuarine water is usually within the range of 7 to 8.5, due to the strong buffering capacity of seawater (Whitfield and Bate, 2007). A review on the physico-chemical conditions of TOCEs in South Africa was undertaken, which revealed differing pH conditions during different phases of the mouth condition but in most estuaries pH generally ranged from 7 to 8.5 (Whitfield and Bate, 2007). In the East

44 Kleinemonde Estuary, the pH tends to range between 7.7 and 8.3 during all conditions of the estuary mouth, namely open, semi-closed and closed (van Niekerk et al., 2008).

The majority of plants grow in a range of pH from 5.5 to 8.3, but a sediment pH between 6 and 6.5 is considered the ideal (Doerr, 1990; Millar, 1992; Bai, 2005; Holley, 2009; Radke, 2010). Salt marsh plants have varying tolerance pH ranges, with Juncus effusus L. as low as 3.5 (McCorroy and Renou, 2003) and as high as 9 for Juncus acutus (Greenwood, 2008). Hroudová et al. (1999) found four morphological types of Bolboschoenus maritimus in Central Europe based on habitat conditions, with a high discriminate efficiency in their relationship to sediment chemistry. Morphological types 1 and 2 grew in freshwater habitats, frequently on acid (pH 5.5 - 6.8), nutrient poor substrates, while types 3 and 4 occurred in more alkaline (pH 7.5 - 8.2) and highly saline habitats. Linthurst and Blum (1980) observed that Spartina alterniflora growth was severely retarded at pH 4 when compared to pH 6 and 8 (cited in Linthurst and Seneca, 1980). The submerged macrophyte Ruppia cirhhosa is tolerant of a wide range of pH, ranging from 7.4 - 9.2 (Setchell, 1924), much like Phragmites australis, which has been found growing in very acidic soils, with a pH of 2.9 and 4.8; and alkaline soils, with the pH ranging from 8.2 to 9.2 (Duke, 1978; 1979; Gucker, 2008). Charophytes, in contrast, prefer a high alkaline pH (8.2) for favourable growth (James et al., 1995; van Nes et al., 1999).

A study by Alvarez-Rogel et al. (2000) in the Mediterranean salt marshes of south east Spain investigated the role of edaphic factors on the salt marsh plant zonation, one of which was pH. Through Canonical Correspondence Analysis, three of the four groups identified were associated with specific pH levels. Among the dry salt marsh species, two groups were identified. The first included stands of Lygeum spartum growing in soils with the lowest pH. The second included species that colonized high pH sediments, growing in a few of the Limonium zones and the halonitrophilous shrub zones, which was dominated by Suaeda vera. In the wet salt marshes, Schoenus nigricans grew in sediments with the highest pH. Among the edaphic variables that best explained the data was maximum pH (the other edaphic variables included maximum K+, maximum soil moisture, mean K+/Na+ ratio, mean Ca+2/Mg+2 ratio).

The influence of alkalinity, which is correlated with pH, as the main factor responsible for the distribution patterns of 106 submerged macrophytes in 82 Danish lakes was elucidated in a study undertaken by Vestergaard and Sand-Jensen (2000). Depending on the alkalinity of the lakes the dominant macrophyte forms varied. Lakes with low alkalinity and pH (pH 5.5) contained numerous isoetids and bryophytes, lakes of moderate alkalinity and pH (pH 7.5) had a range of elodeids (water plants) and vascular plants, whereas vascular plants of the elodeid growth form were predominant in lakes with high alkalinity or pH (pH >8.2). The alkalinity influenced the ability of the macrophytes to use bicarbonate and extract inorganic carbon for photosynthesis, which implied that the distribution had an eco-physiological foundation. Phenotypic plasticity in these submerged macrophytes was also demonstrated in their large variation related to the heterogenous environments studied. Other studies have also demonstrated a clear relationship between macrophyte distribution and pH (Pip, 1979; Catling et al., 1986; Arts et al., 1990; Riis et al., 2000; Riis and Bigg, 2001). Although the study was conducted in lakes, it demonstrates how pH in estuaries may influence the occurrence and distribution of submerged macrophytes. For example, if the estuary pH was not within the pH tolerance range of the selected study species, they would not occur. Further, if the tolerance pH range of submerged

45 macrophytes was modified this would affect macrophyte growth and reproductive phenology in estuaries. For example, catchment pollution (acidification) in a POE would be disadvantageous, or long periods of inundation and the resultant increase in pH in a TOCE would favour submerged macrophytes.

Lime addition can raise sediment pH to overcome some of the problems associated with acidity. In an experiment undertaken by González-Alcaraz (in press), the influence of liming and behaviour of Sarcocornia fruticosa to improve the sediment chemistry of metal polluted salt marsh soils was tested. Soils were taken from two polluted salt marshes (pH ~6.4 vs pH ~3.1) and 20 g kg-1 of lime was added, giving four treatments: neutral soil with/without liming and acidic soil with/without liming. The survival rates of Sarcocornia fruticosa were six times higher in the neutral, amended soil and three times higher in the acidic, amended soil, compared to the unamended treatments. Liming in both the sediments, which increased the sediment pH, enabled the expansion of S. fruticosa and enhanced the capacity of the plants to phytostabilise the metals in their roots. Davy (2001) writes about Salicornia species being invariably associated with alkaline soils (or high pH). Therefore the tolerance ranges of the Salicornieae macrophytes need to be met to enable appropriate phenological responses.

Spatio-temporal variability in sediment pH also occurs, which may influence macrophyte productivity. This was demonstrated in a study in the Knysna Estuary, South Africa, in which the supratidal habitat experienced higher pH values compared to the intertidal habitat (Muir, 2000). In contrast, Redondo-Gomez et al. (2007) found that sediment pH declined with increasing elevation. Wolaver et al. (1986) also found the latter in a mesohaline marsh in Virginia, with pH generally highest toward the tidal creek and lower in the high marsh, including that biological control of subsurface pH variability was evident in both zones. Further, it was found that there was a shift in control over short-term pH variability from tidal inundation to radiation with marsh surface height. Seasonal changes were observed in the Knysna Estuary, where sediment pH became more acidic in summer as opposed to more alkaline in winter, while the lower intertidal habitat remained at a pH of ~7. The affects of inundation in the intertidal zone may have buffered the pH, while the decay of organic matter in spring and the subsequent release of nitrates and sulphates, coupled with low moisture content, may have decreased the pH during summer (Zedler, 1996; Muir, 2000; Radke, 2010).

These fluctuations or changes in pH have the potential to influence macrophyte phenology in estuaries if the threshold pH range is not met. Sediment pH fluctuations will occur in response to both water level and salinity (Baldwin and Mendelssohn, 1998; Al-busaidi and Cookson, 2003). Studies in the East Kleinemonde Estuary have demonstrated that during all phases of the estuary mouth, the water column pH levels range between 7.7 and 8.3 (van Niekerk et al., 2008). This suggests fairly stable water column conditions, as with the Kowie Estuary which generally has a high pH (mean 8.2) and a high alkalinity due to the underlying shale and sandstone (Heydorn and Grindley, 1982; Cowley et al., 2003). An increase in pH to the higher range in the East Kleinemonde Estuary will favour Chara vulgaris phenology as charophytes prefer a high alkaline pH (8.2) for favourable growth (James et al., 1995; van Nes et al., 1999). In contrast, sediment pH may become increasingly acidic in the upper salt marsh due to organic matter breakdown (Muir, 2000), particularly in summer when sediment moisture content is low. If not flushed by freshwater inflow, these conditions may

46 negatively influence macrophyte growth and reproductive phenology or favour some species over others e.g. Juncus species which can tolerate lower pH conditions.

2.3.9 Light, turbidity and temperature Turbidity is a measure of water clarity, which expresses the degree to which light is scattered and absorbed by molecules and particles in the water column. Turbidity is the result of coloured dissolved organic matter and suspended particulate matter, such as silt, detritus and organisms, in the water column (Ward et al., 1998). In estuaries, turbidity and light availability is determined by the amount of suspended particulate matter carried in by freshwater inflow, and the density of phytoplankton and epiphytic algae, and the trophic state (Spence, 1982; Duarte and Kalff, 1990; Adams et al., 1999; Vestergaard and Sand-Jensen, 2000). Eutrophication is known to reduce light availability due to phytoplankton blooms (Crum and Bachman, 1973; Andersen, 1976, Jupp and Spence, 1977) and also indirectly due to epiphyte and filamentous algal growth that has a shading affect on the plants (Phillips et al., 1978). The turbidity levels in seawater entering estuaries along the warm temperate regions of South Africa are generally relatively low (<10 NTU) (Whitfield and Bate, 2007). Turbidity in tidal dominated estuaries, such as POEs, is naturally high because strong tidal currents re-suspend fine sediment and other particulate matter (Uncles et al., 2002). In comparison, TOCEs experience high turbidity during river floods or overwash events (Whitfield and Bate, 2007), but during the closed phase turbidity is generally low (Adams et al., 1999; Walker et al., 2001; Adams and Riddin, 2007; 2008b).

Submerged macrophytes have a minimum light requirement of 5 - 29 % of surface light (Duarte, 1991) and the competition for light is equivalent to the competition by roots for space in terrestrial plants (Duarte and Kaff, 1987). Bulthuis (1983), on the other hand, writes that seagrass distribution data indicate that submerged angiosperms need a minimum of 5 – 15 % of surface irradiance to survive. High turbidity and low light can therefore reduce the photosynthetic potential of submerged macrophytes thereby negatively affecting their growth and reproduction. During the closed mouth conditions of TOCE, when light is favourable due to low freshwater and sediment input, submerged macrophytes grow and expand (Adams and Riddin, 2007). Results from Duarte and Kaff (1987) showed that plants growing in high irradiated tropical waters were five times heavier than those produced in low irradiated temperate waters. In contrast, plants growing in turbid waters (secchi disk <1.5 m) and very deep marine localities (>30 m) had ten times less biomass than the weight predicted for the general limit line of submerged plants. Flood events can cause an in increase in siltation and turbidity and a decrease in available light (Taylor, 1983). This can occur in POEs and TOCEs, during violent flood events as a result of high rainfall which in turn can negatively influence the cover abundance of submerged macrophytes. For example, turbidity was considered to be one of the important variables in the Wilson Inlet, a seasonally closed lagoonal estuary in Australia, which resulted in 40 % of the change in seasonal cover abundance of Ruppia megacarpa (Carruthers and Walker, 1999).

The distribution of submerged macrophytes is determined by turbidity, temperature, light, water velocity, salinity, substratum, nutrient availability and water depth. Light availability and substratum are believed to be the most important environmental factors controlling the occurrence, abundance and distribution of these plants (Howard-Williams, 1979; Howard-Williams and Allanson, 1981; Spence, 1982; Weisser et al., 1992). Submerged macrophytes are not present in estuaries where the sediment is constantly being modified by 47 dynamic processes, such as tidal currents, sedimentation or flooding events. Ideal conditions favouring their establishment in estuaries are high water clarity, limited sedimentation, low water velocity and a salinity range within the tolerance range of the species (Adams and Riddin, 2007). In the Gamtoos Estuary, a POE in South Africa, submerged macrophytes could not be located along the greatest length of the estuary, but were found growing towards the mouth and head of the estuary. Physical factors, such as depth and high turbidity caused by a strong freshwater input, are elevated in the former areas. High turbidity levels are created when there is consistent freshwater inflow and high velocity which transports suspended sediments and debris, thereby reducing irradiance. As a consequence of the reduced light and water clarity, the distribution of the submerged macrophytes is limited to the head and mouth areas of the estuary (Adams et al., 1992). In TOCEs, such as the East Kleinemonde Estuary, during the closed mouth condition and when river inflow is low, turbidity levels are lower (12 NTU) than the open mouth condition (36 – 100 NTU) (van Niekerk et al., 2008). The lower turbidity will therefor favour the growth and expansion of submerged macrophytes.

Light availability can be regarded as one of the chief environmental factors determining macrophyte zonation in lakes (Spence, 1982). An increase in irradiance into the water column would be especially significant for charophytes as they are usually restricted to a thin layer close to the substrate (Coops, 2002), although they do have lower light requirements than the angiosperm submerged macrophytes (Nygaard and Sand-Jensen, 1981; Chambers and Kalff, 1985; Blindow, 1992; Schwarz et al., 1996; Middleboe and Markager, 1997; Steinman et al., 1997, 2001; Schwarz et al., 2002). Steinman et al. (2001) indicated a positive relationship of irradiance with Chara abundance, while abundance was negatively related to water depth in Lake Okeechobee, a large subtropical lake in south-central Florida, USA. The authors hypothesized that the decline in the lake water level due to a managed drawdown increased bottom irradiance thereby enhancing the conditions for Chara to grow. In a positive feedback loop, expansion of Chara beds reduced sediment re-suspension and further improved the underwater light availability allowing for further recruitment. This would be particularly relevant to TOCEs during the closed mouth condition where low turbidity levels would improve irradiance to the bottom sediments and encourage the growth of C. vulgaris and other charophytes since they are often considered as pioneer species (Coops, 2002). This in turn should encourage R. cirrhosa growth and expansion as a result of the improved light conditions.

Submerged macrophytes tend to show a distinct distribution between lakes, determined mainly by water chemistry and, in part, due to transparency related to trophic level (Duarte and Kalff, 1990; Vestergaard and Sand-Jensen, 2000). Vestergaard and Sand-Jensen (2000) indicated that the trophic state and transparency of 82 Danish lakes also influenced the distribution of species. Very eutrophic lakes had only a few robust elodeid species which could compensate for turbid conditions, while small elodeids and slow-growing isoetid species were absent. With an increased nutrient concentration in eutrophic lakes the secchi-depth declines and irradiance and transparency is reduced (Chambers, 1987; Arts et al., 1990; Rørslett, 1991; Srivastava et al., 1995). This modifies the species composition due to the different light requirements of the species (Vestergaard and Sand-Jensen, 2000). This suggests that the higher turbidity levels expected in POEs, such as the Kowie Estuary (Day, 1981), due to the influence of daily tidal exchange (coupled with sediment type) may constrain submerged macrophyte occurrence and distribution. In comparison, TOCEs, that are predominantly closed, such as the East Kleinemonde Estuary, will favour the occurrence and distribution of submerged macrophytes.

48

Steinman et al. (1997) writes that a high annual mean temperature of 24.7°C, including a high alkaline pH 8.2 (James et al., 1995), contributes to the favourable growth of charophytes at Lake Okeechobee, but low light conditions may constrain charophyte productivity and biomass. A distinct seasonal pattern occurred, with maximum biomass accumulation during summer and early autumn. A study undertaken by Steinman et al. (1997), in the same lake, also found that the charophytes displayed a distinctly seasonal phenology, although there was no significant correlation between Chara biomass and temperature. Both distribution and abundance showed a strong seasonal phenology with growth most pronounced from July to November, during summer and autumn. The research suggested that the Chara populations may be a seasonal phenomenon, as Wood (1950) noted that they are transitory plants that suddenly appear for a year or two and rapidly disappear again. Irradiance was also positively correlated to biomass accumulation. This would suggest that TOCEs require stable high water levels during the warmer seasons or when temperatures are milder (which would include autumn in temperature regions) for the chraophyte species to produce an adequate amount of seeds.

In the Swartvlei Estuary, an estuarine-lake complex in South Africa, Howard-Williams and Liptrot (1980) found that the average depth to which photosynthetically active radiation (PAR) is reduced to 5 % of its surface intensity is 3.1 m, whilst the average 1 % level is 5.5 m. They concluded that Potagometon pectinatus is restricted to a water depth where the bottom light level is about 5 % of surface intensity. Chara and Lamprothamnium, in contrast, were both restricted to a lower depth of 1 m due to high turbulence caused by wave action while low light intensity at 3 m depth caused by shading from P. pectinatus also restricted their expansion. Martin (1960), on the other hand found that Potamogeton and Chara grew to within 1.9 to 4 m respectively because of the light limitation beyond 4 m in Groenvlei, just east of Swartvlei. Another example of the importance of light is in Langvlei, which is part of the Wilderness Lakes and is connected to the Touw River Estuary, a TOCE in South Africa that now requires artificial breaching to open the mouth. Chara globularis and Lamprothamnium papulosium were abundant due to clear water conditions in the system. A dino-flagellate bloom was considered to have caused shading which resulted in the almost complete disappearance of the characeae beds. Once phyto-plankton levels decreased, transparency increased and the charophytes returned (Allanson and Whitfield, 1983). This demonstrates that in TOCEs, when turbidity is lower, which is usually during the closed mouth phase, conditions should be more favourable for charophyte growth.

Life-history parameters, such as germination and establishment requirements, including the timing of vegetative production and sexual reproduction, can be important in determining the distribution patterns in charophytes (Casaonva and Brock, 1994). Temperature plays a critical role in the life-cycle stages of submerged macrophytes, from germination to the reproductive periods (Setchell, 1924; Phillips, 1960; Verhoeven, 1979; Casanova and Brock, 1994). Verhoeven (1979) demonstrated that both the growth and flowering of Ruppia cirrhosa were related to water temperature with plant decay occurring four months after exponential growth. According to Verhoevern (1979), R. cirrhosa hibernates during the period of winter quiescence in permanent habitats and that this period is related to temperature. For example in Camargue (France), quiescence occurred at mean air temperatures lower than the min/max interval of 5 - 15°C over a period of 1 - 2.5 months and on Texel (Netherlands) at air temperatures lower than the interval 5 - 10°C for 5 to 6 months. Although these are European examples, warm temperate estuaries of the study region generally experience minimum

49 temperatures of 11 - 16 °C, which shows that temperature should influence macrophyte phenology. In contrast, growth can be negatively influenced by temperatures >30°C in small systems. R. cirrhosa hibernates during winter mainly in a vegetative state but also as seeds. Flowering is also dependant on temperature occurring at min/max water temperature intervals above 15 - 19°C. The decay period is related to high turbulence, water levels >80 cm, high salinity and grazing. Ruppia maritima var. maritima seeds are produced from mid-summer to mid-autumn (June to September) and are dormant till the next spring when temperatures rise again to above the minimum maximum threshold i.e. water temperature >10 – 15°C permitting germination. Similarly, Casanova and Brock (1994) report that the reproduction in dioecious perennials is stimulated by increasing water temperatures while vegetative organs are produced in autumn.

Numerous other studies have shown that the growth and flowering phenology of submerged macrophytes follow a seasonal pattern (Congdon and McComb, 1979; Kiorboe, 1980; Costa and Seeliger, 1989; Fernández- Aláez et al., 2002; Menendez, 2002; Cho and Porrier, 2005; Gesti et al., 2005). Congdon and McComb (1979) investigated the in situ seasonal dynamics, turbidity affects and light requirements of Ruppia maritime L (which could have been R. cirrhosa due to the uncertainty of the authors at the time of the study) in Blackwood River Estuary, Australia. Reduced light and increased turbidity negatively affected biomass accummulation, confirming that Ruppia is most productive in sheltered areas with relatively low water currents and wave action (Wood, 1959). Biomass was generally lowest during late autumn and winter, with rapid increases in spring and summer when flowering generally occurred. Similarly, a study by Obrador and Petrus (2010) in a Mediterranean coastal lagoon also showed a marked seasonal cycle in R. cirrhosa with disappearance in winter, while its spatial distribution was determined by light availability and low turbidity. Inter-annual variability in biomass production (327 – 919 g DW m-2) was ascribed to turbidity during spring and salinity during summer. The macrophytes suddenly disappeared after six years of apparently stable conditions.

One of the most important issues in the survival and propagation strategy of charophytes is the type of reproduction, sexual or asexual. Asexual reproduction is dependent upon water depth and the resultant irradiance intensity, such as the photosynthesis-light relationship (Schwarz et al., 1996; Menendez and Sanchez, 1998) and qualities related to photoinhibition (de Bakker et al., 2001; Vieira and Necchi, 2003), which in turn affect the adaptations of reproductive characteristics. For example, in a study by Asaeda et al. (2007), the gamete and sexual propagule production rates of Chara fibrosa var. fibrosa (A. Br.) and Nitella hyaline (DC.) Ag. declined with water depth, and as a result irradiance, which has been reported in other studies (Bonis and Lepart, 1994; van den Berg et al., 1999). Further, that the shallow water plants did not produce high vegetative biomass, which indicated a potentially higher sexual reproduction rate compared to plants in the deeper water. Consequently, the lower abundance of oospores in deep water and higher abundance in shallow water appeared to be synchronized with the charophyte reproduction strategy, which was influenced by water depth and the resultant irradiance intensity. Germination from seeds or oospores is particularly important in submerged species subjected to fluctuating water levels (Casanova and Brock, 1996; Riddin and Adams, 2009). Seed germination is influenced by light and temperature, as demonstrated by Casanova and Brock (1996). The germination percentage (76 %) of Chara muelleri oospores increased after subjection to fluctuations in light intensity and temperature. In comparison, the germination percentage (41 %) of C. australis oospores was highest after being kept wet for four months.

50

In this study the submerged macrophytes were investigated in the TOCE only because the selected species were not found during the sampling period in the POE. It is anticipated that submerged macrophyte phenology in TOCEs will be positively correlated with seasonal cycles relating to temperature (Setchell, 1924; Phillips, 1960; Verhoeven, 1979; Casanova and Brock, 1994), but strongly influenced by water depth, low turbidity and favourable light conditions as well (Martin, 1960; Howard-Williams and Liptrot, 1980; Spence, 1982; Schwarz et al., 1996; Menendez and Sanchez, 1998; de Bakker et al., 2001; Vieira and Necchi, 2003). Further, that these conditions will be met during the closed mouth phase as opposed to the open mouth phase when turbidity and water depth is lower (Riddin and Adams, 2008a; van Niekerk et al., 2008).

2.4 SEED VIABILITY

Research has shown that as habitat disturbance increases, seed production and growth from seed reserves are more important (Verhoeven, 1979; Kautsky, 1990; Casanova and Brock, 1996; Combroux and Bornette, 2004). Knowledge of seed viability in the context of stressful and fluctuating environmental conditions in TOCEs where species may be locally extirpated is therefore important. The viability of the seeds is a measure of how many seeds are living and able to develop into new plants that have the potential to reproduce in the future. Interactions between salinity and temperature control germination viability (Khan and Ungar, 1984; Khan et al., 2000), although osmotic effects and ion toxicity are considered to be the two main factors causing germination inhibition (Greenwood, 2008). High salinity may inhibit the germination of macrophytes (Chapman, 1974; Ungar, 1978, Silva et al., 2006) but it does not necessarily affect seed viability (Greenwood, 2008). In contrast, glycophyte seeds are frequently unable to survive high salinity concentrations (Khan et al., 2000a/b; Qu and Huang, 2005; Greenwood, 2008). Shumway and Bertness (1992) demonstrated that reducing sediment salinity by watering bare areas in the salt marsh in situ increased germination by >50 % in Distichlis spicata, Juncus gerardii, Atriplex patula, Iva frutescens, Limonium nashii, and Solidago sempervirens. This is in contrast to seeds of glycophytes which do not retain their viability under salt stress (Ungar, 1995; Ungar, 2001). Salinity tolerance, like other stressors such as moisture and temperature, is highest in dormant seeds (Baskin et al., 1998). Although salinity may not exclude long-term survival (i.e. seed viability) it may still be compromised (Katembe et al., 1998; Pujol et al., 2001; Redondo et al., 2004; Goodman et al., 2011).

Several macrophyte seeds have developed some form of plasticity in their germination responses, for example Salicornia, Atriplex, Chenopodium, Plantago and Salsola (Casanova and Brock, 1991; Khan and Gul, 1998). One of the general adaptation mechanisms that have evolved in annual species, such as Salicornia europaea and Atriplex triangularis Willd growing in saline wetland or estuary habitats, is the production of a persistent seed bank of viable but dormant seeds. This allows for multiple germination opportunities, beyond the growing season in which the seeds were produced. Seed dimorphism (or polymorphism), such as differences in seed size or physiological behaviour, is of selective advantage because it helps plants to respond directly to changing environmental conditions (Ungar, 1978; Khan and Ungar, 1984a/b). Polymorphic seeds vary in their dormancy level, which provides an extended period of germination in salt marsh and desert environments (Khan and Gul, 1998) given the stressful nature of these environments e.g. high salinity or lack of moisture. Atriplex triangularis, for example, has small and large seeds where the larger seeds are able to germinate at

51 higher salinity levels (Khan and Ungar, 1984a/b; Khan and Gul, 1998). Polymorphic seeds have different physiological requirements therefore they provide alternative conditions for germination to occur when conditions vary in the environment, remaining dormant yet viable during unfavourable periods. Consequently, polymorphism is of selective advantage because it helps seeds respond directly to the environment (Khan and Ungar, 1984a; Khan and Gul, 1998). For example, Redondo et al. (2004) recorded a reduction of seed viability due to increasing salinity in Sarcocornia perennis (low marsh), Sarcocornia fruticosa (high marsh) and S. perennis spp. fruticosa (middle marsh). Dormancy of viable seeds therefore guarantees that germination can occur in the future under favourable conditions or an appropriate season, whereas germination pattern may be an adaptation to unpredictable environments (Casanova and Brock, 1996). This strategy will be particularly important in TOCEs due to the unpredictable nature of these environments, as a large persistent seed bank will allow germination in the future and the persistence of species.

Seed dimorphism occurs in Salicornia ramosissima J. Woods with the small seeds developing from the lateral flowers and the big seeds developing from the large central flower. Delayed germination in the small seeds may be the result of polymorphism, observed by Silva et al. (2006), because they are more dormant and less tolerant of high salinity than the big non-dormant seeds (Philipupillai and Ungar, 1984; Silva, 2000). Delayed germination may be a genetic response to hypersaline conditions in the upper salt marsh (Jefferies and Gottilieb, 1982), while it has also been reported that larger seeds have an enhanced chemical constitution (hormones) and larger nutrient reserves which improves germination success (Khan and Ungar, 1985; Austenfeld, 1988; Vleeshouwers et al., 1995; Lombardi et al., 1997; Silva et al., 2006). This plasticity will be particularly important in TOCEs during either the open or closed mouth phases because hypersalinity and inundation may occur, which may inhibit germination. Once the water recedes to expose bare areas, germination of the more viable and stronger seeds can take place. Similarly, in POEs where the supratidal areas are too saline due to high summer temperatures and limited rainfall or drought, these seeds will remain dormant and then germinate when conditions become more favourable and/or when germination requirements are met.

In predictable environments, such as POEs, early germination provides a competitive advantage while in less predictable habitats, such as in TOCEs, it may constitute a risk (Silvertown, 1988). Consequently, dormancy of viable seeds in TOCEs is advantageous. Several factors may induce dormancy, such as reduced light, burial or the requirement of an after ripening period (Ungar, 1984). For example, Casanova and Brock (1991; 1996) found that the prerequisite for breaking the dormancy of viable seeds of charophytes in temporary waters is a period of desiccation including alternating periods of moisture. Similarly, Sederias and Colman (2007) observed that primary dormancy at initial release from Chara vulgaris was broken by exposure to low temperatures (4C°), followed by a secondary dormancy period that required stratification and light exposure. Ruppia seeds can remain dormant in the sediment for three years (Kantrud, 1991) and when the water level is high seeds germinate from a large seed bank (Riddin and Adams, 2007). The seeds of several submerged species are known to be viable for at least 15 years (Hutchinson, 1975; van Vierssen, 1993), while de Winton et al. (2000) found viable seeds after 17 and 23 years. Zannichellia pedunculata Reichen. and Zannichellia obtusifolia Talavera have been depicted as having reduced long-term viability (Bonis et al., 1995; Capers, 2003). Seed viability for the emergent macrophyte, Bolboschoenus maritimus, has been recorded at 20 years (USDA, 2010).

52 Seed viability and dormancy are therefore particularly important in unpredictable and dynamic environments, such as TOCEs. The ability to remain dormant yet viable for long periods (up to decades) guarantees that seed dormany can be broken under favourable conditions and provides numerous opportunities for germination to occur to ensure the persistence of macrophytes in dynamic environments.

The annual eelgrass, Zostera marina L. (eelgrass) of Bahia Concepción, the Gulf of California, is a unique annual subtidal meadow that grows only in the coldest season of the year, while surviving the adversely warm summer season as seeds. The plant is adapted to complete its life-cycle in approximately six months with a brief period of vegetative growth and very high seed production. High seed output is vital for survival of the population during the unfavourable season of the year, notwithstanding the requirement for viable seeds to germinate during succeeding growing seasons (Santamaria-Gallegos et al., 2000). Seed dormancy frequently occurs when the parent plant from which the seeds were produced will not survive the existing conditions (Vleeshouwers et al., 1995). This strategy enhances the persistence of the population at a given locale, even if seeds are not produced annually. This is particarly important in TOCEs because unfavourable conditions may extend beyond an annual life-cycle which means species will not be able to produce seeds in a particular year. However, the presence of a persistent seed bank will guarantee that viable yet dormant seeds are available for germination in the proceeding years when favourable conditions return, such as high water level and low turbidity during the closed mouth phase in the case of submerged macrophytes compared to lower water level in the case of salt marsh plants. A residual seed bank provides the potential for habitat re-establishment after a germination event that does not produce a reproductive population, which is characteristic of ephemeral wetlands in variable climates (Brock and Rogers, 1998; Brock et al., 2003), as can be said for TOCEs.

The importance of persistent seed banks compared to transient seed banks was demonstrated in a study by Greenwood (2008). The mature seeds of Juncus acutus are able to germinate rapidly in freshwater and light. However, when exposed to high salinity or buried under sediment, seeds remain viable and may persist for several years until favourable conditions return. In contrast, the seed bank of Phragmites australis is transient and only replenished during spring seed production (Fenner and Thompson, 2005). Seeds gradually decrease once dispersed due to several factors, such as ephemeral persistence and predation, resulting in a low seed bank by late autumn or winter. This gives J .acutus a competitive advantage over P. australis and is especially important where P.australis has become a problem in estuaries (Greenwood, 2008). The ability to accumulate persistent seed banks that respond to declining water levels is a key factor in the vegetation dynamics of wetlands with fluctuating water levels (van der Valk and Davis, 1978 cited in Fenner and Thompson, 2005), such as TOCEs. For example, Rumex species from frequently flooded areas have persistent seed banks compared to those higher up the elevation gradient that have transient seed banks (Voesenek and Blom, 1992 cited in van der Sman, 1993). Flooding during the closed mouth phase of TOCEs will therefore prevent germination, but when the water level drops germination can occur due to a persistent seed bank of dormant yet viable seeds.

Some mature macrophyte seeds require a period of over-wintering or after ripening and therefore will not germinate directly on collection from the parent plant. The germination of the emergent species Scirpus is dependent on the permeability of the seed coat or pericarp, despite high viability (Harris and Marshall, 1960;

53 O'Neill, 1972; Lacroix and Mosher, 1995). Some of the seeds are therefore dormant at the time of seed set, as has been found for many wetland plants (Haslam, 1972; Dietert and Shontz, 1978; Baskin et al., 1989). Scirpus acutus, for example, requires an over-wintering period (Lacroix and Mosher, 1995) and therefore germination can occur the following spring or when temperatures are warm or milder. The low-temperature limit for seed germination is unknown, but germination in many species may be prevented only by freezing (Fenner and Thompson, 2005). Similarly, Scirpus robustus Pursh requires an after ripening period and a pre-chilling treatment increases germination (Dietert and Shontz, 1978). Baskin et al. (1989) found that the seeds of Gratiola viscidula Pennell and Scirpus lineatus Michx were dormant at maturity and became non-dormant by early to mid-winter. Freshly exposed seeds of Cyperus odoratus L. and Penthorum sedoides L. were conditionally dormant at maturity but germinated to near 100 % only at high temperatures in light. They could therefore germinate to high percentages throughout the growing season, whereas G. viscidula and S. lineatus had higher germination percentages in spring compared to summer (Baskin et al., 1989). Bolboschoenus maritimus achenes are known to overwinter in submerged sediments and in the spikelets on the plant, germinate poorly at maturity and achieve 97 % germination two months after maturation (Kantrud, 1996). Clevering (1995) also states that fresh achenes require stratification or other treatments to improve germination whereas Isely (1944 cited in Kantrud, 1996) found that an after-ripening period is necessary. It has been recorded that the oospores of the submerged species Chara vulgaris required a 60 day after ripening period (Sederias and Colman, 2007) and the germination of Ruppia seeds are stimulated by a period of desiccation (Kantrud, 1991). In contrast, fresh Chara zeylanica klein ex Willd and Chara contraria A. Br. oospores have shown high germination rates ranging from 75 – 95 % without a period of after ripening (Proctor, 1960).

In contrast to the above, other macrophyte seeds can germinate rapidly when matured on the plant. Various researchers that have collected seeds from salt marsh plants when fruit are ripe have shown this. For example, ripe fruits of Salicornia ramosissima were collected from shrubs growing in Odiel salt marshes, Spain. They showed a very high germination percentage, close to 95 %, while 71 % of the seeds that did not germinate were in dormancy and 29 % were unviable. Redondo et al. (2004) collected seeds from ripe plants and germination results ranged from 85 – 88 % in Sarcocornia perennis, S. fruticosa and S. perennis spp fruticosa. Germination trials by Ungar (1977) on the seeds of the annual Salicornia europaea collected from plants was ± 62 %, whereas trials by Khan et al. (2000a) had 95 % germination success in the annual S. rubra. Seeds of S. europaea on the same inflorescence have shown different levels of dormancy (Philipupillai and Ungar, (1984), which probably explains the lower germination results by Ungar (1976). Seeds of Halocnemum strobilaceum (Pallas) M.B., a rare halophyte that grows in Spain, were collected and 100 % germination success was achieved (Pujol et al., 2001). Germination trials by Greenwood and MacFarlane (2009) on seeds harvested from Juncus acutus and J. kraussi showed 95 % germination success, as did Vicente et al. (2007) for J. acutus. Mateos-Naranjo et al. (2008) conducted in situ germination trials with ripe Spartina alternifolia seeds in the Odiel marshes of Spain. The highest germination result was 37 % and the lack of germination was considered to be due to anoxic conditions and low moisture content.

Germination trials on seed banks and seed age on emergent macrophytes, on the other hand, have also shown good germination results, despite the fact that seed age influences germination responses (Baskin and Baskin, 1998; Fenner and Thompson, 2005). Sarcocornia tegetaria seeds collected from seed banks in the East

54 Kleinemonde Estuary had 82 % germination success (Riddin and Adams, 2009). Geissler and Gzik (2008) found that Juncus atratus seed germination remained high after various flood and drought treatments over a three year period i.e. 71 - 82 %. Noe and Zedler (2002) demonstrated that out of 11 annual salt marsh species, the seed age of eight species did not affect percentage germination. Zedler et al. (1992) also state that annual salt marsh plants probably develop long-lived seed banks that can germinate years later. However, viable seed banks of ephemeral wetlands in Australia have been shown to decline due to long periods of drought and high salinity (Goodman et al., 2011). Davy (2001) also reports that seeds of Salicornia in Britain that have remained in the seed bank after one year lose their viability, but that persistent S. europaea seed banks have been reported under more severe and less predictable conditions. Phragmites australis seed banks are transient (Fenner and Thompson, 2005) but plants produce dormant and non-dormant seeds which either germinate immediately or the following spring/autumn (Ekstam and Forseby, 1999; Martinez-Ghersa et al., 2000; Kettenring and Whigham, 2009). Literature for seed germination trials in seeds collected from seed banks demonstrating percentage germination success, other than emergent trials, was limited in the emergent macrophytes. Consequently comparions could not be adequately made. More importantly however, the literature suggests that the seeds of most emergent species remain viable over the long term which allows the persistence of species in dynamic habitats, such as TOCEs.

Germination trials in submerged species showed varied results. Experiments by Spencer and Ksander (2002) resulted in 76 % germination success in the seeds of the perennial Zannichellia palustris L. which were collected from sediment cores. In comparison, both Chara vulgaris and Ruppia cirrhosa had low germination percentages, between 11 and 15 %, from the East Kleinemonde Estuary sediments (Riddin and Adams, 2009). Germination trials by de Winton et al. (2000) on seed bank oospores from 15 New Zealand lakes recorded a mean seedling emergence of 0.2 – 58 % in Chara and Nitella species. Chara globularis was the predominant species with the highest germination rate (57 %) whereas Nitella species had very low germination rates (5 %). Seed bank results from Boedeltjie et al. (2002) showed poor germination in unscarified seeds of four Potamogeton species (1 – 23 %), while germination in the scarified seeds of P. natans was high (85 – 90 %). Bonis and Grillas (2002) provide the germination rates recorded by various researchers for eight Chara species and two Nitella species. Germination from oospores collected from the seed banks ranged from 0 – 48 % compared to freshly harvested oospores (0 – 96 %) under varied temperatures (0 - 28°C and 6 – 25 % respectivly). The authors suggest that the low germination results from the seed banks may be because seed banks are perennial. This means that oospores are able to buffer against the risk of reproductive failure over the long term due to the perennial strategy and consequent germination pattern. Fresh seeds collected from the submerged freshwater plant, Vallisneria americana, achieved 75 – 80 % germination success (Campbell, 2005), whereas Greenwood and du Bowy (2005) showed that Zannichellia palustris seeds from the seed bank had a higher germination success than fresh seeds (49 versus 36 %). Results from Sederias and Colman (2007) also showed higher germination in sedimentary Chara vulgaris oospores (28 – 45 %) compared to fresh oospores (4 - 5 %). The literature demonstrates that the seeds of submerged species remain viable over the long term which is advantageous in disturbed habitats, such as TOCEs, as this allows for the persistence of the submerged species despite the occurrence of unfavourable periods e.g. low water level and dessication.

55 In unpredictable habitats such as wetlands, or TOCEs, it might be expected that older seeds in the seed bank are non-dormant and germinate directly in response to favourable environmental conditions (Baskin et al., 1989) allowing for ongoing regeneration of the species (Baskin and Baskin, 1979). Seed viability over the long term therefore allows the species to regenerate after local extirpation due to stressful conditions, such as fluctuating water levels in TOCEs. The fact that only freezing conditions may prevent germination (Fenner and Thompson, 2005) suggests that the potential for species to germinate outside of their usual seasonal periods e.g. spring for salt marsh plants (Packham and Willis, 1997; Davy, 2001), is possible in TOCEs where the inundation of the intertidal and supratidal zones during closed mouth conditions may occur when seasonal germination usually takes place.

Flowering time is an important consideration for the production of viable seeds, particularly in unstable habitats like TOCEs. This is because the annual flowering period may be interrupted due to inundation in TOCEs during the closed mouth phase as a result of rising water level, storm surges and freshwater inflow that is too low to breach the estuary mouth. As a result, flowering does not occur and seeds are not produced. Further, the production of viable seeds can vary between plants and populations of the same species. Plants that flower during the peak flowering period, for example, may produce the highest number of filled seeds with improved germination and higher kernel weight (Fang et al., 2004). This was demonstrated in Spartina alterniflora, which generally produces seeds of low viability that remain viable for approximately one year (Mooring et al., 1971). Self-pollinated plants generally have lower seed set compared to open pollinated plants, with the added disadvantage of producing no viable seed (Daehler and Strong, 1994). Charpentier et al. (2000) also found that fecundity may be reduced due to clonal growth and self pollination in small populations of Bolboschoenus maritimus. The study by Fang et al. (2004) found that the total number of florets per panicle of S. alterniflora steadily decreased, and the percentage of filled florets and germination increased as the flowering season progressed (Mullins and Marks, 1987). Results indicated that heavier kernels were associated with good seed set, and late flowering within the peak flowering period could result in a better more viable seed set. Similarly, Boedeltjie et al. (2004) writes about hydrochorous plants that have dispersal peaks in spring or summer, which release mainly non-dormant diaspores versus those with peaks in autumn and winter that release mainly dormant diasporas. This suggests that autumn, winter and early spring floods disperse mainly dormant diasporas, which allows species to bridge unfavourable seasons, or unfavourable conditions in TOCEs.

Knowing the seasonal or annual reproduction period is therefore important in dynamic environments, such as TOCEs, because this will assist with mouth management policy. For example, managers will need to ensure that open mouth conditions occur during periods when flowering and seed production is at its peak to allow macrophytes to produce viable seeds in order to replenish seed stocks. A persistent seed bank of viable seeds and the variable germination of these seeds in response to fluctuating environmental conditions are critical in TOCEs as favourable conditions are not constant but spatio-temporally dynamic. In contrast, although POEs may experience fluctuations in environmental conditions that may be unfavourable for germination to occur, such as high salinity during summer, these fluctuations tend to be less dynamic and may not occur for extended periods like in TOCEs (e.g. two years). Viability is therefore important in both systems but especially important in TOCEs when favourable periods may not exist for long periods or may fluctuate rapidly over short periods and which may result in die-back of parent plants prior to peak flowering. A persistent seed bank

56 means that a population could persist without immigration, even if seed production does not take place over a year or more due to flooding during the growing season (Baskin and Baskin, 2004). All the selected macrophytes in this study, apart from Phragmites australis and most probably Sporobolus virginicus, have persistent seed banks which assist with their long term survival during periods of unfavourable conditions, such as flooding. Understanding whether seeds require an over-wintering or after ripening period is also critical to determine if an extended period is required before germination can occur as this allows managers to determine when germination is possible and lower water level is necessary. This in turn will influence management decisions in TOCEs. Scirpus acutus, for example, requires an overwintering period (Lacroix and Mosher, 1995) while Bolboschoenus maritimus overwinters in the sediment and/or spikelets (Kantrud, 1996). As a result, seeds will only germinate after a period of chilling (e.g. low winter temperatures break dormancy). Consequently, water level should be low at the time when seeds are able to germinate and preferably during ideal conditions, particularly when considering transient seed banks, such as P. australis and S. virgnicus.

2.5 SYNOPSIS

It is clear from the literature review that macrophyte phenology is influenced by multiple environmental factors. Further, macrophytes are affected differently at different stages of their life-cycle. Estuaries, particularly TOCEs, are highly unpredictable environments that experience rapid changes in environmental conditions in response to mouth condition. Life-cycle and reproductive strategies are closely coupled with water level fluctuations and the associated fluctuation in environmental factors. These fluctuations can have a significant effect on species presence and distribution.

Seasonal change, with the associated temperature, rainfall and irradiance fluctuations, triggers changes in both the growth and reproductive phenology of estuarine macrophytes (Thompson and Grime, 1983; Packham and Willis, 1997; Zedler et al., 2000; Laegdsgaard, 2006). Numerous research findings indicate that high salinity reduces vegetative production, while it is decisive during the reproductive period reducing both reproductive output and germination (Ungar, 1962; Chapman, 1974; Ungar, 1977; Ungar, 1978; Ungar; 1998; Riddin and Adams, 2008a). Research on the East Kleinemonde Estuary showed that water level fluctuations are a key driving force affecting the spatial and temporal distribution of macrophytes (Riddin and Adams, 2008; 2010). Flooding events may result in complete die-back of submerged species in response to a drop in water level (Riddin and Adams, 2008a). Tides are a significant factor influencing the health and functioning of salt marshes (Congdon and McComb, 1980; Adam, 1990; Clarke and Jacoby, 1994; Zedler et al., 2001; Adam, 2002; Greenwood, 2008). Regular tidal exchange, such as in POEs, generates well developed salt marshes (Adams, 1991; Davy, 2000; Rogel et al., 2000; Rogel et al., 2001; Bockelmann et al., 2002; Costa et al., 2003; Ursino et al., 2004). Due to changing mouth condition in TOCEs, tidal exchange can vary. During the closed mouth phase, the occurrence and distribution of macrophytes may be reduced due to the lack of tidal exchange. During the open mouth condition, tidal exchange should encourage the growth of supratidal and intertidal salt marsh. Although salt marsh soils are often anoxic, increased submergence, particularly during closed mouth conditions, reduce sediment redox potential. Macrophyte phenology will be negatively influenced and die-back is likely to occur because many of the species are intolerant of protracted inundation (> two months) and highly reduced sediments. Sediment moisture content is critical for seed germination, seedling establishment and 57 macrophyte growth, particularly so with reference to its influence on salinity and redox potential. Sediment pH fluctuations will occur in response to both water level and salinity (Baldwin and Mendelssohn, 1998; Al-busaidi and Cookson, 2003). These fluctuations or changes in pH have the potential to influence macrophyte phenology in estuaries if the threshold pH range is not met.

The distribution of submerged macrophytes is determined by turbidity, temperature, light, water velocity, salinity, substratum, nutrient availability and water depth. Light availability can be regarded as one of the chief environmental factors determining macrophyte zonation in lakes (Spence, 1982). Numerous studies have shown that the growth and flowering phenology of submerged macrophytes follow a seasonal pattern (Congdon and McComb, 1979; Kiorboe, 1980; Costa and Seeliger, 1989; Fernández-Aláez et al., 2002; Menendez, 2002; Cho and Porrier, 2005; Gesti et al., 2005). Temperature plays a critical role in the life-cycle stages of submerged macrophytes, from germination to the reproductive periods (Setchell, 1924; Phillips, 1960; Verhoeven, 1979; Casanova and Brock, 1994).

Research has shown that as habitat disturbance increases, seed production and growth from seed reserves are more important (Verhoeven, 1979; Kautsky, 1990; Casanova and Brock, 1996; Combroux and Bornette, 2004). One of the general adaptation mechanisms that have evolved in species growing in estuary habitats is the production of a persistent seed bank of viable but dormant seeds. Seed viability over the long term allows the species to regenerate after local extirpation due to stressful and fluctuating environmental conditions, such as fluctuating water levels in TOCEs. A persistent seed bank of viable seeds and the variable germination of these seeds in response to fluctuating environmental conditions are critical, especially in TOCEs. This is because favourable conditions are not constant in TOCEs but spatio-temporally dynamic. Transient seed banks, on the other hand, do not provide viable seeds over the long term. Consequently, water level should be low at the time when seeds are viable and able to germinate, preferably during ideal conditions.

Physico-chemical fluctuations are comparatively short in duration in POEs, such as freshwater flooding, and are relatively predictable because these events are linked to seasonal rainfall. In contrast, these fluctuations occur unpredictably in TOCEs causing sudden changes that may last for several months or even years. Consequently, it is anticipated that due to the permanent open mouth condition in POEs, physico-chemical fluctuations are not expected to be as spatio-temporally dynamic as they are considered to be in TOCEs. The elevated spatio-temporal dynamism in TOCEs can have significant implications for macrophyte phenology, which distinguishes these systems from POEs where the disturbance to macrophyte phenology is considered less extreme.

58 3. CHAPTER 3: MATERIALS AND METHODS

3.1 LOCATION OF STUDY SITES The East Kleinemonde (33°32′S, 27°03′E) and Kowie (33°36'S; 26°54'E) estuaries are situated along the south- eastern Cape coastline at the small coastal towns of Kleinemonde and Port Alfred in South Africa. Both estuaries are classified as warm-temperate systems, which occur from the Silwermyn Estuary in False Bay, near Cape Town, to the Mendu Estuary in the Eastern Cape, north of East London (Figure 3.1).

Figure 3.1: Map of southern Africa showing the three major biogeographic regions (Whitfield, 2000) and the East Kleinemonde and Kowie estuaries along the south-eastern Cape coastline of South Africa (Cowley et al., 2001).

3.2 CLIMATE OF STUDY SITES Both estuaries fall within the warm temperate geographical region where rainfall is approximately 500 mm per year occurring almost equally in all seasons, although during autumn (March) and spring (October/November) rainfall is slightly higher (Heydorn and Tinley,1980; Schulze, 1984) creating a bimodal rainfall pattern (Heydorn and Grindley, 1982; Jury and Levy, 1993). Rainfall is considered highly variable with minimum rainfall occurring in June (Lubke, 1983; Kopke, 1988). All months have at least 60 mm of rain. However, over the study period rainfall was less than 60 mm during several of the months and the study area experienced drought from early 2009. According to Jury and Levy (1993) drought cycles occur every 3.45 to 18.2 years. Average rainfall in Port

59 Alfred was generally higher in most of the months compared to East Kleinemonde (Figure 3.2). Temperatures of this region are regarded as mild in both winter and summer ranging between 10-22°C, with wind reducing the heat and humidity in summer (Lubke, 1988a). The average annual temperature is 17ºC with a maximum mean daily temperature of 22°C in December and January; and a minimum mean daily temperature of 14°C in July.

Figure 3.2: Monthly rainfall (mm) over the study period from February 2009 to June 2010 in the East Kleinemonde and Kowie estuaries.

3.3 THE EAST KLEINEMONDE ESTUARY 3.3.1 Physical characteristics The East Kleinemonde Estuary is one of 175 temporarily open/closed estuaries (TOCE) in South Africa. It is classified as a medium sized TOCE that is ranked number 54 out of 250 in terms of its conservation importance estuaries in South Africa (Turpie et al., 2002). The surface area of the estuary is 35.7 ha and its catchment is approximately 43.5 km² (Badenhorst, 1988; Riddin and Adams, 2009). Mean annual run-off has been estimated at 2 million m3 (Badenhorst, 1988). The estuary is navigable for 3 km and is approximately 100 m wide in the lower and middle reaches and 25 m wide in the upper reaches. The main channel has an average depth of 2.5 m, but the majority of the estuary has a littoral zone of less than 1 m deep and a large shallow mudflat occurs upstream of the bridge.

Mouth dynamics of the system have been studied showing that mouth breaching events and the number of days the mouth is open are directly related to river inflow (van Niekerk et al., 2008). Freshwater discharge after heavy rainfall events exceeding 100 mm is typically the cause of the estuary mouth breaching (Cowley and Whitfield, 2001; van Niekerk et al., 2008). The mouth usually breaches subsequent to high river inflow or high water levels (>2 m amsl) caused by overwash, or high rainfall events. Mouth closure usually occurs at water levels varying between 0.5 and 1 m amsl (Riddin and Adams, 2008a). Water levels rise to 2 - 2.5 m amsl during extended periods of mouth closure due to a sand bar developing at the mouth when extensive back flooding occurs (van Niekerk et al., 2008). From March 1993 to August 1997 the estuary was predominantly closed, while open mouth conditions were unseasonal and of only short duration after high rainfall (Cowley and

60 Whitfield, 2001). The estuary can remain closed for extensive periods (e.g. two years) due to the sand bar at the mouth (Riddin and Adams, 2008a) and low rainfall or freshwater input. The estuary is predominantly closed and on average opens 2.6 times per year (van Niekerk et al., 2008; Riddin and Adams, 2008a) and overwash events occur 16.4 % of the time (Cowley, 1998). The estuary is known to experience cyclical 1:30 year floods (Whitfield and Bate, 2007).

Average monthly water temperatures range from 15°C (June) to 27°C (January). Average monthly salinity ranges from 0 to 25 ppt depending on the amount of rainfall and mouth condition (van Niekerk et al., 2008). During the intermittently open/closed (high flow) condition, the East Kleinemonde Estuary experiences salinity levels ranging from 37 ppt, in the lower to middle reaches, to 1 ppt in the upper reaches. During peak flood conditions the flood plain is inundated for a short period but otherwise flooding of the floodplain does not occur. Oxygen levels are high and turbidity levels have been known to range from 36 to 100 NTU. In comparison, during the closed mouth condition the estuary is usually saline (>25 ppt) and inundation of saltmarsh may occur -1 at water levels above 1.8 m amsl. Oxygen levels may drop to below 2 mg L , while turbidity levels are variable but average around 12 NTU (van Niekerk et al., 2008).

Land usage in the East Kleinemonde River catchment is characterized by extensive agricultural activities, principally cattle farming. Relatively pristine valley thicket covers the stream and river valleys. The lower reach of the estuary is built up with residential homes along the banks, which increases the amount of freshwater run- off into the estuary.

3.3.2 Macrophytes Salt marsh grows along the west bank above the road bridge and along the east bank in the middle to upper reaches of the estuary. The dominant species are Juncus kraussii, Sarcocornia decumbens and Sporobolus viriginicus in the supratidal zone (>1.8 m amsl) and Sarcocornia tegetaria and Salicornia meyeriana in the intertidal zone (<1.3 m amsl). Stands of Phragmites australis are usually associated with small patches of Bolboschoenus maritimus and grow intermittently along the banks in the lower and middle reaches of the estuary. Submerged macrophytes grow in a continuous band along both banks above the road bridge. The dominant submerged macrophyte is Ruppia cirrhosa while Chara vulgaris has a fragmented and scattered distribution. The seagrasses Halophila ovalis, Potamogetun pectinatus and Lamprothamnium papulosum can also occur in the estuary (Adams and Riddin, 2007; Riddin and Adams, 2008b). During the open mouth phase, submerged macrophytes are not present due to substrate instability, high water velocity and high turbidity. In contrast, during the closed mouth phase when water levels are >1.5 m amsl, Ruppia can achieve a maximum potential biomass of 2 883 g DW m-2, Chara 599 g DW m-2 and Halophila 101 g DW m-2, while intertidal salt marsh is inundated. The supratidal salt marsh is only affected when water level is >1.8 m amsl (Riddin and Adams, 2008b).

61 3.4 THE KOWIE ESTUARY 3.4.1 Physical characteristics The Kowie Estuary is a permanently open estuary (POE) that is ranked number 31 out of 250 estuaries in South Africa in terms of its conservation importance (Turpie et al., 2002). The Kowie River is 70 km long with a catchment area of 769 km² (Heydorn and Grindley, 1982). Average annual rainfall is 638 mm (Day, 1981). The estuary is tidal for 21 km and has a surface area of approximately 120 ha with mean annual run off of the river estimated at 17x106 m³ to 50x106 m³ (Forbes, 1998; Heydorn and Grindley, 1982). Although a perennial river, under drought conditions flow can stop for three to four months (Stewart et al., 1962 cited in Heydorn and Grindley, 1982). It is subject to large variations in flow patterns (Day, 1981) and is considered erratic due to the frequent droughts and floods in the catchment (Cowley et al., 2003). Short periods of violent flooding have been known to occur in the past (Heydorn and Grindley, 1982).

The upper reaches of the estuary are frequently vegetated to the water‘s edge and meanders extensively through steep banks. It has a narrow channel and intertidal zone, less than 10 m wide, and average water depths range from 2-6 m. The middle reaches are sandy with intertidal mudflats of 100 m or more. It broadens out to 100-150 m wide and has an average water depth of 3 m. The lower reaches have an 80 m wide artificial channel, with a marina linked to the estuary. The mouth of the estuary is extended by two 75 m piers (Heydorn and Grindley, 1982; Forbes, 1998). The upper estuary has fine sand to silt sediments, in contrast to the middle reaches which have a greater proportion of fine sand and mud on the intertidal banks and sandy sediments in the main channel.

River flow is strongly influenced by the tide and bottom sediments are subject to strong tidal flows. In South Africa, spring tide has a tidal height of more than 1 m amsl and during neap tide is only 0.25 m amsl (Schumann et al., 1999), while high tide levels in large open estuaries, such as the Kowie Estuary, are similar to the adjacent coastline (Whitfield and Bate, 2007). In the Kowie Estuary, the tidal range is 1.62 m amsl and the spring tidal ranges are 1.7 in the lower reaches, 1.5 in the middle reaches and 1.1 in the upper reaches (; Heydorn and Grindley, 1982). The river usually carries low silt loads (mean secchi disc transparency 71–103 mm) due to the sediment type and well established vegetation in the majority of the catchment (Heydorn and Grindley, 1982; Cowley et al., 2003). Currents in the vicinity of the mouth are driven mainly by waves, while the natural littoral sediment drift has been modified by the construction of piers either side of the mouth. Due to the presence of the piers, a build up of sediment on the north-west side of the west pier and a depletion of sediment on the north-east side of the east pier has resulted (Heydorn and Grindley, 1982).

Average monthly water temperatures range from 20 - 28 °C in summer and 11-16 °C in winter. Salinity is generally above 30 ppt but during dry years it has increased to 40 ppt, while water entering the river is brackish (Day, 1981; Heydorn and Grindley, 1982). Bok (1983) found water temperatures to differ in the upper reaches and mouth region ranging from 11 – 27°C and 14 - 22°C. Surface waters are almost fresh during flood periods which can last up to two to four weeks, but the estuary mouth remains 30 – 35 ppt (Day, 1981; Whitfield et al., 1994). Turbidity is usually high due to the sediments of the Bokkeveld rocks which are red and clayey (Day, 1981; Giffen, 1970 cited in Heydorn and Grindley, 1982). In previous studies, the average salinity in the Kowie

62 Estuary was 30.4 ppt, while the dissolved oxygen and turbidity measured 7 mg L-1 and 6.9 NTU respectively (Harrison, 2003). Surface waters are usually well oxygenated and influxing water may be supersaturated with oxygen, ranging from 11.2 – 13.5 mg L-1. The river has a high pH (mean 8.2) and a high alkalinity (139 – 185 ppm CaCO3) due to the underlying shale and sandstone (Cowley et al., 2003).

Land usage in the catchment of the Kowie River is predominantly intensive agriculture, although indigenous thicket vegetation grows adjacent to the estuary banks along the upper and middle reaches (Heydorn and Grindley, 1982). The lower reaches of the estuary are surrounded by residential and business development for approximately 10 km upstream of the mouth (Forbes, 1998). The lower part of the estuary is flanked by loose stone packed berms along the banks, which according to Day (1981), severely modify the natural configuration of the estuary resulting in a general impoverishment of the system. The marina development situated on the east bank, approximately 100 m from the mouth, destroyed the Blue Lagoon, which was a clean body of water with depths of approximately 2 m and extensive salt marshes (Heydorn and Grindley, 1982). Considerable ecological degradation has occurred in the estuary, particularly due to eelgrass beds and reed habitat destruction in the lower reaches, as well as in its catchment (Coetzee et al., 1997).

3.4.2 Macrophytes The dominant macrophyte species, which grow particularly in the middle reaches of the estuary, include Phragmites australis, Bolboschoenus maritimus, Juncus acutus, Disphyma crassifolia, Chenolea diffusa, Sarcocornia tegetaria, Sarcocornia pilanssii and Sporobolus viriginicus. Spartina maritima grows along the intertidal zone and is probably the most dominant macrophyte. Patches of Sarcocornia decumbens and Salicornia meyeriana also occur intermittently along the length of the estuary. Plantago carnosa and Cynodon dactylon are also abundant along the estuary banks (Heydorn and Grindley, 1982). Zostera capensis and Halophila ovalis occur submerged along the estuary banks at the lower intertidal zone and in shallow waters (Cowley and Daniel, 2001). Ruppia cirrhosa and R. maritima have been found in the lagoons adjacent to the estuary in the past (Day, 1981; Heydorn and Grindley, 1982; Cowley and Daniel, 2001). According to Day (1981), the canalization of the lower estuary probably reduced the presence of Z. capensis and salt marsh vegetation.

3.5 MACROPHYTE SPECIES SELECTED Species were selected based on the various estuarine habitats and the most common species found within each habitat. In addition, these species are influenced by water level due to their position along the elevation gradient. Macrophytes vary in their adaptability and tolerance to a particular water regime and interspecific differences in plant responses to fluctuating water level are fundamental to habitat differentiation in estuaries (Noe, 2002; Olff, et al., 1988; Mitsch and Gosselink, 1993).

The following macrophyte species were selected and are grouped per habitat:  Supratidal salt marsh: Juncus kraussii Hochst, Juncus acutus L., Sporobolus virginicus (L.) Kunth.

 Intertidal salt marsh: Upper intertidal - Sarcocornia decumbens (Tölken) A.J. Scott.

63 Lower intertidal - Sarcocornia tegetaria, Salicornia meyeriana Moss.

 Reeds and sedges: Phragmites australis (Cav.) Trin ex Steud, Bolboschoenus maritimus (L.) Palla.

 Submerged macrophytes: Ruppia cirrhosa (Petagna) Grande and Chara vulgaris L.

3.6 MACROPHYTE GROWTH AND REPRODUCTIVE OUTPUT IN THE SELECTED SPECIES OF THE EAST KLEINEMONDE AND KOWIE ESTUARIES

The macrophytes were surveyed along three permanent transects within the lower and middle reaches of the East Kleinemonde Estuary (Riddin and Adams, 2008a/b) and at one sample site in the middle reaches of the Kowie Estuary (Figures 3.3 and 3.4). Only one sample site was selected in the Kowie Estuary due to logistical and safety reasons. A once off sampling trip was undertaken along the length of the estuary in May 2010 to determine whether macrophyte phenology, in general, within the Kowie Estuary was similar to the study sample site. It was confirmed that the phenology of the selected species was similar at other sites along the length of the estuary (Semenya, 2010).

Sampling was undertaken over a 17 month period, from February 2009 to June 2010, at monthly intervals. All species were monitored in a minimum of 10 permanent 1 m² quadrats (n = 10). The quadrats were positioned within the submerged, supratidal, intertidal and reed and sedge habitats.

3

2

1

Figure 3.3: Sampling points (white circles) and transects (numbered white lines) in the East Kleinemonde Estuary (Source: Google Earth, 2008).

64

Figure 3.4: Sampling site in the Kowie Estuary (Source: Google Earth, 2010).

3.6.1 Emergent macrophytes Quadrats in the supratidal and intertidal habitat of the East Kleinemonde Estuary were inundated with high water level from July 2009 to June 2010 and from July 2009 to January 2010 respectively. Consequently, the majority of seeds of Sarcocornia decumbens and Sarcocornia tegetaria; and all the Salicornia meyeriana seeds germinated in areas where the water receded further up the elevation gradient in the supratidal habitat. Supplementary quadrat replicates were therefore established from January 2010 and as the seedlings expanded into new bare areas. S. meyeriana germinated in new quadrats in January 2010 (n = 8) and expanded to June 2010 (n = 12), therefore the total number of quadrats, including the 10 original quadrats which were inundated till June 2010, was 22. S. tegetaria germinated in January 2010 in two of the original quadrats plus seven new quadrats (n = 9) and expanded to June 2010 (n = 23), therefore the total number, including the eight original quadrats which were inundated till June 2010, was 31. S. decumbens germinated in April 2010 (n = 8) and expanded to June 2010 (n = 15) therefore the total number, including the two original quadrats which were inundated till June 2010, was 25.

In the Kowie Estuary, a Sarcocornia hybrid was misidentified as S. decumbens at commencement of the study period because of its position in the supratidal zone. Consequently, quadrats for S. decumbens were only established in May 2009. Juncus kraussii was not found in the Kowie Estuary and therefore Juncus acutus was used as a comparison as these species are very similar. Although J. acutus grows in the East Kleinemonde Estuary, their abundance is low, particularly when compared to J. kraussii.

The following biotic features were monitored in 1 m² quadrats to determine the rate of growth and reproductive output: percentage vegetation cover, plant height, number of inflorescences, percentage inflorescence phase, percentage inflorescence cover and seed numbers per m² (average value per m²). The average percentage plant cover was determined visually within each 1 m² quadrat using a metal 1 m² square grid (n = 10 per species) i.e. the percentage of surface occupied by the species. Plant height of 10 individual plant stems was measured using a wooden calibrated pole (cm) within each 1 m² quadrat per species. The number of

65 inflorescences for Phragmites australis, J. kraussii, J. acutus and Bolboschoenus maritimus were counted within each 1 m² quadrat. Due to the high density of inflorescences for S. decumbens, S. virginicus, S. tegetaria and S. meyeriana, these were counted within a 15 cm X 15 cm quadrat (225 cm²) within the 1 m² quadrat. In the Kowie Estuary, S. meyeriana plants were more compact and highly branched compared to those in the East Kleinemonde Estuary. Therefore plants were collected from ten replicate 225 cm² quadrats and the number of inflorescences counted to calculate the average number of inflorescences per 225 cm² (i.e. 406), which was used in subsequent sampling trips. The following formula was used to extrapolate the average number of inflorescences per m² from the number of inflorescences counted in 225 cm² for all these species:

Average no. of inflorescences (Inf.) per m² = (no. of inf. per 225 cm² X species cover per m²) X % inf. cover per m²

The percentage inflorescence phase referred to that percentage of the inflorescences that were flowering, fruiting and seeding and/or had released the seeds (Figure 3.5). The fruiting and seeding phase was combined as the fruit and seed developed together. In some instances it could be determined that the seed was immature. For example, Bolboschoenus maritimus fruit are brown when mature and therefore green fruit indicated that seed were immature. This influenced the identification of the peak seeding month(s).

a b

Figure 3.5: (a) S. decumbens inflorescences that were 100 % seeding indicated by the corky perianths; and (b) S. meyeriana inflorescences that were 100 % flowering.

It is possible that seed production of S. virginicus in the Kowie Estuary was over-estimated because it was difficult to determine the phase of the inflorescences without the use of a microscope. The inflorescences that were collected outside of the quadrats were used to extrapolate to the number of seeds contained within the inflorescences of those in the 1 m² quadrats. Ten supplementary quadrats were established in October 2009

66 due to considerable flowering in the surrounding habitat compared to the quadrats established at commencement of the sampling period.

The percentage inflorescence cover per species was determined visually within each permanent 1 m² quadrat i.e. the percentage of species cover occupied by its inflorescences. In order to extrapolate to the average number of seeds per m² per species, 20 inflorescences per species were collected in the East Kleinemonde Estuary and the average number of seeds per inflorescence determined (Table 4.2.5). The following formula was then applied:

Average no. of seed per m² = (no. of inf. per m² X no. of seed per inf.) X % fruiting & seeding phase inflorescences /100

The average flower, fruit or seed per inflorescence for the species were compared with available literature.

3.6.2 Submerged macrophytes The following were monitored for submerged species to determine the rate of growth and reproductive output: average monthly above-ground biomass (g DW), plant height, number of flowering and fruiting inflorescences, number of flowers, number of fruits and seeds and number of antheridia and oogonia.

Average monthly biomass of the submerged species was determined by collecting ten biomass samples (n = 10) within the area where the plants were growing using a 10.5 cm diameter (86.9 cm²) corer. Ruppia cirrhosa was collected at transect one and transect three adjacent to the west bank, while Chara vulgaris was collected at transect one adjacent to the west bank (Figure 3.5 above). The material was transferred to a re-sealable bag and transported to the laboratory in a cooler box. Biomass was cleaned of debris and sediment using a 1 mm (R. cirrhosa) and 250 µm (C. vulgaris) mesh filter, towel dried and allowed to air dry. Air drying of material took approximately 1 – 7 days depending on the weather. Material was considered air dried if the material felt dry. Biomass was then oven dried for two hours at 60°C. The oven-dried material was weighed to the third decimal place (0.000 g). Plant height of the submerged species was determined by measuring 20 individual plant stems and leaves with a wooden calibrated pole (cm). Flowers and fruit/seeds, including flowering and fruiting inflorescences were counted to determine the average number of flowers, fruit/seeds and inflorescences or oogonia and antheridia per m². Due to the massive number of sexual organs produced by C. vulgaris, three replicate sub-samples (n = 3) of the biomass collected from the 86.9 cm² corer was sampled. The number of oogonia and antheridia were counted under a microscope. To extrapolate to the average number of inflorescences and seeds (or fruit, oogonia, and antheridia) per m² from the number of inflorescences and seeds counted from the three replicate sub-samples, the following formula was used:

Average no. of inflorescences (Inf.) or seeds per m² = no. of inf. or seeds counted X 115.5*

(* 115.5 = no. of corers per m²)

67 3.6.3 Terms used in this study Phenology is the study of the timing of periodic life-cycle or biological events of plants, the cause of their timing with regard to biotic and abiotic forces or how these are influenced by the environment, particularly temperature changes as driven by seasonal and interannual variations in climate (Leith, 1974 cited in Pierce, 1983; Scwartz, 2003). In this study, the research investigates the effects of environmental conditions on the growth and sexual reproduction (flowering and seeding) of macrophytes. The reproductive period or cycle refers to that period when the macrophytes are actively flowering and producing fruit and seed. Peak flowering was considered that period when the majority of inflorescences were flowering i.e. the highest number of flowering inflorescences. Flowering was defined as the period when flowers could be identified on the macrophyte peduncle via the emergence of flowering components e.g. anthers, stamen or early buds. R. cirrhosa was considered flowering when the inflorescence peduncle had emerged and extended by at least 0.5 cm, although penduncle lenghth of R. cirrhosa is determined by water depth (Verhoeven, 1979). The first flowering period refers to flowering that occurred first, which for some species started in 2008 before commencement of the sampling period (2009) while for others it commenced in the first half of 2009. The second flowering period occurred after the first flowering period, either during the second half of 2009 or in 2010. Seed set is refered to as the number of ovules that successfully develop into seeds i.e. seed production per plant. A viable seed is a mature or fully developed seed which has the capacity to germinate and produce another plant, also known as seed viability. A dormant seed is a mature seed that is viable but does not have the capacity to germinate in a specified time period under normal physical environmental factors that are usually favourable for germination (Baskin and Baskin, 2004), also known as seed dormancy. After-ripening is a period when specific changes must occur in some dormant but fully developed or viable seed before germination can take place. It is a time and environmentally regulated process in the dry seed that determines the potential for seed germination (Baskin and Baskin, 1998). Over-wintering is the normal exposure of viable seed to a winter season. Some species may germinate, whereas others only germinate during the warmer spring temperatures after dormancy over the winter period. Peak seeding refers to that phase of the reproductive cycle when seeds are at their maximum on the plant. It represents the end of new fruit and seed production during the cycle, including the end of the reproductive period and seed release or loss from the inflorescences predominates. The seasons of this study are defined in Table 4.2.3.

3.7 THE TIME FROM SEED GERMINATION TO SEED FORMATION IN THREE EMERGENT MACROPHYTES AND TWO SUBMERGED MACROPHYTES Germination was recorded when Sarcocornia tegetaria, Sarcocornia decumbens and Salicornia meyeriana seedlings emerged in situ. These three emergent species were selected because they readily germinate from seed compared to the other emergent species which reproduce vegetatively via rhizomes. This makes it difficult to determine new seedling growth from seeds without destroying the seedlings. In the East Kleinemonde Estuary, supplementary 1 m² quadrat replicates were established because seed germinated where the water receded further up the elevation gradient and the majority of the original quadrats were still inundated when the study was concluded in June 2010. During successive sampling trips, additional 1 m² quadrats were established as seedlings expanded. S. meyeriana germinated in January 2010 (n = 8 quadrats) and expanded its distribution until June 2010 (n = 11). S. tegetaria germinated in January 2010 (n = 2) and expanded its distribution until June 2010 (n = 23). S. decumbens germinated in April 2010 (n = 8) and expanded to June 2010 (n = 15). In the Kowie Estuary, only S. meyeriana and Sarcocornia hybrid seeds germinated in situ. S.

68 meyeriana germinated in the original quadrats (n = 11) while the Sarcocornia hybrid germinated in July 2009 (n = 5) and expanded to June 2010 (n = 11).

3.8 THE VIABILITY OF THE SEEDS ATTACHED TO THE MACROPHYTES AND IF VIABILITY CHANGED AFTER MATURATION The importance of testing for seed viability was threefold, firstly to determine if the seeds required an extended period of time to reach full maturity in a TOCE, i.e. an after ripening or over-wintering period, after seeds have matured on the inflorescence as all seeds will require a certain period to develop to maturity in the fruit. Seeds were therefore only collected from the parent plants once considered matured. Secondly, to test how long the seeds remain viable, as seed viability will influence macrophyte reproductive plasticity in relation to unpredictable and changing environmental conditions in a TOCE. Germination trials were therefore only conducted for the macrophytes from the TOCE due to the varying mouth conditions as opposed to the POE where a permanent open mouth condition does not necessitate seed dormancy or long term viability. Thirdly, in order to provide mouth management recommendations for TOCEs, seed viability data were required to establish when the mouth needs to be open or closed and for how long i.e. timing and duration for germination to take place.

Studies by Fang et al. (2004) suggest that flowering is an important determinant in influencing seed set, and that selection should be made for plants that flower within the peak flowering period as this will result in the highest number of filled seeds with improved germination and higher kernel weight. Selection was therefore assured during the peak seeding period (which follows on from the peak flowering period) and as long as there were seeds attached to the plant, before and after peak flowering.

Seeds were collected on a monthly basis when mature and for as long as they were attached to the inflorescences. No seeds were collected for germination trials if they were immature e.g. green Bolboschoenus maritimus achenes or small and white undeveloped Juncus kraussii seeds. Seeds of Sarcocornia tegetaria, Sarcocornia decumbens and Salicornia meyeriana were collected when the perianth was corky (Figure 3.5). At the end of the sampling period, seeds from several succulent inflorescences of S. tegetaria and S. meyeriana were collected and the percentage germination was high. As a result, in hindsight, seeds should have been collected for viability testing prior to the perianth turning corky. At commencement of the sampling period in February 2009, seeds of B. maritimus, J. kraussii and J. acutus had already developed and therefore seeds from the second flowering period during late 2009 and 2010 were required to determine viability more accurately from the time of maturation.

Although the objective was to test the viability of seeds from the East Kleinemonde Estuary or TOCE, inundation prevented seed collection for three of the selected species and seeds from the Kowie Estuary were therefore tested. Due to the prolonged period of inundation in the TOCE, S. decumbens and S. virginicus did not flower and produce seeds during the second flowering period. Consequently, seeds from the Kowie Estuary were harvested. B. maritimus seeds from the Kowie Estuary were also collected early in 2010 during the

69 second flowering period, although seeds developed aseasonally in the East Kleinemonde Estuary during mid 2010 and seed viability could be tested.

Seeds were placed on two Whatman No. 1 filter papers (diameter 900 mm) inside a 900 mm tight-fitting plastic petri dish and placed in a nursery shade-house. The filter paper was moistened with freshwater daily or as moisture was lost. Each species had five replicates, with 20 seeds each. Germination was defined by the emergence of the radicle (Naidoo and Kift, 2006). Germination was monitored daily until the onset of germination and then for every one to three days (Geissler, 2008). Experiments occurred until germination ceased, which usually took up to 28 (Naidoo and Kift, 2006) to 30 days. However, final germination percentages were determined after six months for B. maritimus because achenes germinate poorly at maturity, but 97 % germination is achievable after two months (Kantrud, 1996). The viability of seeds was plotted on a graph indicating percentage germination.

3.9 ENVIRONMENTAL VARIABLES AND MACROPHYTE PHENOLOGY 3.9.1 Submerged macrophytes Temperature data were retrieved from the South African Weather Bureau, while rainfall data were acquired from a local East Kleinemonde resident. The following ambient physico-chemical conditions of the water column were measured at the three transects within the submerged habitat: temperature, salinity, electrical conductivity, pH, redox potential, water level and turbidity. Water temperature was recorded using an YSI 650 MDS Multiprobe and a secchi disc to measure turbidity. A wooden calibrated pole (cm) was used to measure water depth. A water sample was taken for measuring redox potential and pH in the laboratory. Although submerged species were not sampled in the Kowie Estuary, a water sample was taken to measure salinity in the laboratory using a refractometer as change in salinity in response to daily tidal inundation of the intertidal zone and monthly spring tides were expected.

3.9.2 Emergent macrophytes Rainfall data for the East Kleinemonde Estuary was acquired through a local resident. The following ambient physico-chemical conditions of the sediment were measured for the emergent habitats: water level, salinity, electrical conductivity, pH and redox potential. Sediment organic matter content and moisture content data were only available for the East Kleinemonde Estuary as the data were acquired from another study in the estuary and did not form part of the original proposal for this study. These data were available from February 2009 to February 2010. Three replicate sediment samples for the measurement of salinity (5 g), electrical conductivity (250 g), pH (5 g), redox potential (5 g), organic matter content (5 g) and moisture content (5 g) were taken within each species habitat for measurement in the laboratory. Sediment samples were transferred to re-sealable plastic bags, placed in a cooler box and transported to the laboratory.

3.9.2.1 Water depth Average daily water level for the estuary was obtained from a continuous water level recorder (P4H002) mounted beneath the R72 Bridge which crosses the lower reaches of the estuary. In the Kowie Estuary, six of

70 the ten S. tegetaria quadrats and all the B. maritimus quadrats were inundated daily due to tidal exchange. Average water depths in both habitats were 2 cm and 29.5 cm during high tide respectively. All the other species were positioned within the supratidal habitat and daily inundation did not take place. Monthly water level in the Kowie Estuary was therefore not recorded.

3.9.2.2 Sediment salinity An air-dried sample of sediment (5 g) was shaken together with 25 ml of distilled water in a 50 ml beaker on an electrical horizontal shaker at 90 rpm for 20 minutes. The suspension was filtered through Whatman No. 1 filter paper. This was done in replicates of three. The salinity was measured using a SDT salinity meter calibrated at 20C. The salinity was determined in parts per thousand (Barnard, 1990).

3.9.2.3 Sediment electrical conductivity An air-dried soil sample weighing 250 g was placed in a 600 ml glass beaker and moistened with de-ionised water. The sediment was then tested for the properties of a saturated paste by tapping the beaker with a spatula on the bench to see if any excess water remained on the top of the paste. If the paste was dry, more distilled water was added. The amount of de-ionised water added was recorded. This gave an indication of the amount of water lost through evaporation. The saturation paste was allowed to stand for at least an hour (Barnard, 1990). The saturation paste was then filtered through a Buchman filter (porous ceramic filter) using Whatman No.40 filter paper. The conductivity of the filtrate was measured using a Beckman  310 series conductivity meter (handheld). This was done in replicates of three.

3.9.2.4 Sediment pH Distilled water (25 ml) was added to 10 g of an air-dried sediment sample and thoroughly mixed on an electrical horizontal shaker at 90 rpm for 30 minutes. The sample was allowed to settle for 30 minutes and thereafter stirred (Black, 1965). The pH readings were taken after 10 seconds using a Beckman  310 series EDTA pH meter (Black, 1965). The pH meter was calibrated at 20C for pH 7 and 10. Three replicates of each sample were taken (Black, 1965).

3.9.2.5 Sediment redox potential Redox potential was measured within 24 hours of collection using a Metrohm AG 9101 electrode (redox probe) which was calibrated with a Metrohm Standard redox solution (6. 2306. 020) at a potential of U = + 250.5 mV and at a temperature of 20C. Three replicates of each sample were taken.

3.9.2.6 Sediment moisture content The methodology as set out by Black (1965) was followed. All visible organic matter and debris were removed from the samples. The 10 g samples were weighed to 0.01 g accuracy, placed in a crucible and a drying oven

71 for 48 hours at 100C. The samples were re-weighed and the percentage moisture content was determined using the following equation: Wet mass – Dry Mass/Wet mass * 100.

3.9.2.7 Sediment organic matter content The methodology as set out by Briggs (1977) was followed. The dried sediment samples from the sediment moisture content experiment were placed in a muffle furnace (ashing oven) for 8 hours at 550 °C. The crucibles were removed from the ashing oven and placed in a desiccator containing anhydrous silica crystals until cool. The percentage organic matter was calculated as a loss of mass during ashing as a percentage of the initial mass using the following equation:

% sediment organic matter = (W1 – W2)/ W1 * 100. W1 = Dried organic matter (g); W2 = Ashed (g).

3.10 STATISTICAL ANALYSIS Values presented are always means (±SE) of the recorded data (n = 10). Data were tested for normality using Statistica (Statistical software developed by Statsoft, Inc.) and proved to be normal. Significant t-tests were run using Statistica (Statistical software developed by Statsoft, Inc.) to obtain significant monthly and seasonal differences in plant cover and plant height per estuary and between estuaries. Significant monthly differences in inflorescence number and seed number during peak inflorescence and seed production periods between the two estuaries were compared. Significant monthly differences in percentage germination and between estuaries were also compared, where data was available.

The seasonal species and environmental data for the East Kleinemonde and Kowie estuaries were analysed using CANOCO for Windows (Version 4.5, Ter Braak and Smilauer, 2002). Principle Components Analysis (PCA) was used to obtain an ordination of the macrophyte data constrained by environmental variables. Multivariate analysis (or PCA) was used because the research investigated the affect of multiple environmental variables on macrophyte phenology i.e. the statistical methodology allows the researcher to investigate more than one response variable at the same time. Further, PCA is a linear method of indirect gradient analysis which looks for one or more mutually independent gradients that represent "optimal" predictors for fitting the regression models of linear species response (Leps and Smilauer, 1999). The result of the PCA was plotted as a two-dimensional graph using CANODRAW for Windows (Version 4.5, Ter Braak and Smilauer, 2002). The environmental variables were plotted as arrows originating from the center of the graph. The origin represents the mean value of each separate variable and the direction of the arrow line represents an increase in the value of that particular variable. Each arrow points in the expected direction of the steepest increase in the value of the environmental variable. Species arrows or the arrows relating to the biodiversity features e.g. plant cover, seed number also point in the direction of the steepest increase in the value of the macrophyte data. The angles between the arrows indicate correlations between individual environmental variables and between individual macrophyte data. The length of the environmental arrow indicates the importance of the variable and is equal to the multiple correlation of the variable with the displayed ordination axes. If the species arrow lies far from the coordinate origin, in the direction indicated by the environmental arrow, the two variables are predicted to have a positive correlation (covariance). They have a negative correlation if the arrows point in the opposite direction. An arrow near the coordinate origin suggests that the two variables have a low correlation. A similar

72 interpretation can be based on the angle between the two arrows. Statistical results were displayed in a table below each PCA ordination diagram. This data is required for an accurate interpretation of the ordination diagram (CANODRAW for Windows, Version 4.5, Ter Braak and Smilauer, 2002).

73 4. CHAPTER 4: RESULTS

4.1 ABIOTIC CONDITIONS IN A TEMPORARILY OPEN/CLOSED ESTUARY COMPARED WITH A PERMANENTLY OPEN ESTUARY

Tables 4.1.1 to 4.1.3 provide a summary of the physico-chemical characteristics associated with the habitat for the individual species per estuary. Statistical data are provided in Appendix 9.1 per habitat.

4.1.1 Mouth condition of the East Kleinemonde Estuary During the sampling period the mouth was closed and the water levels ranged from 1.5 to 2.4 m amsl (Figure 4.1.1). Several overwash events took place, with a significant overwash occurring on the 24 June 2009 due to a storm event. The overwash event increased the water level of the estuary from 2 to 2.4 m amsl. The supratidal zone was therefore inundated for a long period i.e. from July 2009 to January 2010 and intertidal zone to June 2010. Salinity was high for most of the study period ranging from 30-42.3 ppt, excluding February 2009 (23.9 ppt), which was due to high rainfall in January 2009 (average 81 mm).

Figure 4.1.1: Average water level and salinity in the East Kleinemonde Estuary (EK), including water salinity in the Kowie Estuary (KW). Arrows indicate the overwash events in the East Kleinemonde Estuary.

4.1.2 Supratidal habitat The species inhabiting the supratidal zone included Juncus kraussii, Juncus acutus, Sporobolus virginicus and Sarcocornia decumbens. In the East Kleinemonde Estuary, the sediment salinity ranged from 7.3 to 38.7 ppt but was mostly between 16 and 30 ppt (Table 4.1.1). The electrical conductivity ranged from 12.6 to 77.2 mS and pH ranged from 3.8 to 7.9, but was mostly alkaline throughout the study period. Conditions were very acidic for three months i.e. February 2009 to March 2009. The redox potential ranged from -313.9 to +455.9 mV, but was mostly negative throughout the study period, due to inundation. The monthly average water level measured in the quadrats ranged from 0 to 74.9 cm. The S. virginicus habitat experienced the highest water levels for the longest period due to its position closer to the water along the elevation gradient. The percentage

74 sediment moisture content and organic matter content ranged from 6.4 to 22.4 % and 0.7 to 4.9 % respectively (Table 4.1.2).

In the Kowie Estuary, the sediment salinity ranged from 5 to 30 ppt but was mostly between 10 and 20 ppt (Table 4.1.1). The electrical conductivity ranged from 10 to 53.9 mS. The pH ranged from 6.4 to 8.2 but was mostly alkaline throughout the study period. The redox potential ranged from -192.8 to +385.1 mV but was mostly positive throughout the study period. The supratidal habitat was not affected by tidal inundation during the sampling period. Sampling did not occur during spring tide and, as a result, it was not determined whether this habitat was affected during this period. However, the wrack line was never observed above the intertidal zone, which indicated that the supratidal zone was most likely not inundated during spring tide given the spring and neap high tide range of ~1 to 0.25 m amsl respectively.

Sediment salinity and EC in the TOCE were significantly higher compared to the POE (t = 4.9; 5.9; p < 0.05), while sediment pH was more extreme or significantly lower in the TOCE versus the POE (t = -6.5; p < 0.05;). Sediments were highly reduced for a prolonged period in the TOCE, whilst sediments were mostly well aerated in the POE. Sediment redox potential was therefore significantly different between the two estuaries (t = -5.6; p < 0.05). The supratidal habitat in the TOCE was inundated for most of the sampling period, compared to the POE where there was no tidal inundation. Statistical data are provided in Appendix 9.1.

4.1.3 Intertidal habitat The species inhabiting the intertidal zone included Salicornia meyeriana and Sarcocornia tegetaria. In the East Kleinemonde Estuary, the sediment salinity ranged from 14.9 to 53.8 ppt but the majority of readings were between 20 and 30 ppt (Table 4.1.1). The electrical conductivity ranged from 29 to 97.8 mS but the majority of readings were between 29 and 56 mS. The pH ranged from 6.2 to 8.3 but was mostly alkaline throughout the study period. The redox potential ranged from -269.7 to +276.8 mV but was mostly negative throughout the study period due to inundation. The monthly average water level measured in the quadrats ranged from 0 to 65.6 cm. February 2009 was the only month that the quadrats were not inundated. The percentage sediment moisture content and organic matter content ranged from 7.9 to 31.4 % and 1.6 to 5.2 % respectively (Table 4.1.2).

In the Kowie Estuary, the sediment salinity ranged from 4.5 to 30.7 ppt but the majority of readings were between 16 and 28 ppt (Table 4.1.1). The electrical conductivity ranged from 8.9 to 55.4 mS. The pH ranged from 6.8 to 8.4 but was mostly alkaline throughout the study period. The redox potential ranged from -159.5 to +266.7 mV but was mostly positive throughout the study period.

Sediment was highly reduced for a prolonged period in the TOCE, whilst sediments were mostly well aerated in the POE. Sediment redox potential was therefore significantly different between the two estuaries (t = -5.2; p < 0.05). The intertidal habitat in the TOCE was inundated for almost the entire sampling period, compared to the POE where six of the ten quadrats were inundated daily. Mean sediment pH was significantly lower (t = -4.6; p

75 < 0.05; and mean sediment electrical conductivity and salinity were significantly higher in the TOCE compared to the POE (t = 3.5; 2.5; p < 0.05). Statistical data are provided in Appendix 9.1.

4.1.4 Reed and sedge habitat The species inhabiting the reed and sedge habitat included Phragmites australis and Bolboschoenus maritimus. In the East Kleinemonde Estuary, the sediment salinity ranged from 9.2 to 27.3 ppt but the majority of readings were between 15 and 26.6 ppt (Table 4.1.1). The water column salinity ranged from 23 to 42.3 ppt but the majority of readings were between 30 and 35 ppt. The electrical conductivity ranged from 17.8 to 54.4 mS. The pH ranged from 6.5 to 8.1 but was mostly alkaline throughout the study period. The redox potential ranged from -324.5 to +120.3 mV but was mostly negative throughout the study period due to inundation. The monthly average water level measured in the quadrats ranged from 3.6 to 106.1 cm but was high for most of the study period. The percentage sediment moisture content and organic matter content ranged from 5.7 to 30.8 % and 1.6 to 5.2 % respectively (Table 4.1.2).

In the Kowie Estuary, the sediment salinity ranged from 7.8 to 37.2 ppt but the majority of readings were between 15 and 29 ppt (Table 4.1.1). The water column salinity ranged from 21 to 34.2 ppt (Figure 4.1.1 above), but the majority of readings were between 30 and 35 ppt. The electrical conductivity ranged from 15.5 to 57.8 mS. The pH ranged from 7 to 8.2 and the redox potential ranged from -403.9 to +297.1 mV. Redox potenital was mostly negative in the B. maritimus habitat due to inundation during high tide and mostly positive in the Phragmites australis habitat due to its location higher up in the supratidal zone.

Sediments were highly reduced and inundated for a prolonged period in the TOCE, whilst sediments were mostly well aerated and never inundated in the P. australis habitat in the POE. Sediment redox potential was therefore significantly different between the two estuaries (t = -2.3; p < 0.05). Mean sediment pH was significantly lower in the TOCE compared to the POE (t = -4.9; p < 0.05).

76

(a)

(b)

Figures 4.1.2: (a) Transect one (west bank) in the East Kleinemonde Estuary with J. kraussii in the background, S. meyeriana, S. tegetaria and S. virginicus in the foreground. (b) Transect two (east bank) with P. australis and B. maritimus habitat in the East Kleinemonde Estuary. Both photographs were taken at the start of the sampling period (February 2009).

77

(a) (b)

Figures 4.1.3: (a) Transect one in the Kowie Estuary with S. tegetaria and patches of S. meyeriana and S. decumbens. (b) Transect two with P. australis and J. acutus habitat in the Kowie Estuary. Both photographs were taken at the start of the sampling period (February 2009).

4.1.5 Submerged habitat The species inhabiting the submerged habitat included Ruppia cirrhosa and Chara vulgaris in the East Kleinemonde Estuary only. Salinity ranged from 23 to 42.3 ppt but the majority of readings were between 30 and 35 ppt (Table 4.1.3). The electrical conductivity ranged from 37.6 to 63.3 mS. The pH ranged from 7.2 to 8.5. The redox potential ranged from -295.5 to +246.9 mV but was mostly positive throughout the study period. The monthly average water level in the quadrats where R. cirrhosa occurred ranged from 6 to 83.3 cm and from 1.3 to 85.7 cm for C. vulgaris. The turbidity ranged from 0.8 to 1.7 m while the water temperature ranged from 14.6 to 28.7 °C. The percentage sediment moisture content and organic matter content ranged from 8.1 to 22.5 % and 1.1 to 3.8 % respectively.

78

Table 4.1.1: Characteristics of the habitats of the emergent species in the East Kleinemonde Estuary compared with the Kowie Estuary.

Sediment Sediment Redox Potential Sediment Salinity Sediment pH Electrical (mV) (ppt) Species Conductivity (mS)

EK KW EK KW EK KW EK KW Supratidal habitat Juncus spp. Mean±SE 12.2±69.3 143.9±40.3 7.0±0.2 7.7±0.1 44.2±3.1 27.9±2.8 23.0±1.4 15.0±1.9 Minimum -264.3 -112.5 5.0 7.1 22.3 11.0 12.0 5.6 Maximum 456.0 311.9 8.0 8.1 60.6 40.7 30.3 26.3 S. virginicus Mean±SE -43.2±48.3 160.6±50.1 6.9±0.3 7.7±0.1 38.4±3.0 29.6±3.4 20.2±1.7 16.1±2.0 Minimum -271.2 -192.8 4.7 6.8 25.8 10.0 13.2 5.0 Maximum 276.7 385.1 7.9 8.1 63.7 53.9 35.1 29.7 S. decumbens Mean±SE -4.6±68.0 147.3±49.4 6.3±0.4 7.5±0.1 40.3±5.1 30.4±3.7 20.5±2.5 16.3±2.3 Minimum -314.0 -133.9 3.9 6.8 12.7 10.7 7.3 5.4 Maximum 427.3 384.0 7.9 8.1 77.2 50.0 38.7 30.0 Intertidal habitat Sarcocornia Mean±SE 141.4±37.4 7.6±0.2 28.4±2.9 15.9±1.8 hybrid Minimum -131.9 6.5 10.9 5.5 Maximum 253.5 8.2 50.5 27.3

S. meyeriana Mean±SE -83.3±37.6 125.8±43.4 7.6±0.1 7.9±0.1 50.7±4.8 34.9±3.9 26.3±2.6 18.5±2.1 Minimum -269.7 -138.0 6.8 7.3 30.1 9.0 16.0 4.5 Maximum 162.4 265.5 8.3 8.4 97.8 55.4 53.8 28.0 S. tegetaria Mean±SE -74.9±51.9 90.1±45.9 7.1±0.2 7.6±0.1 42.3±3.6 37.0±2.8 22.4±2.0 22.1±1.9 Minimum -264.4 -159.5 6.2 6.8 29.0 14.7 14.9 7.4 Maximum 276.8 266.7 8.0 7.8 77.7 47.4 42.7 30.7

Reed and sedge habitat P. australis Mean±SE -85.7±33.0 120.4±46.5 7.3±0.2 7.7±0.1 43.7±3.6 37.6±3.5 22.3±1.6 22.3±2.4 Minimum -279.1 -191.4 6.5 7.1 21.5 15.5 12.3 7.8 Maximum 107.7 297.1 8.0 8.2 54.4 57.8 27.3 37.2 B. maritimus Mean±SE -167.9±33.7 -205.3±47.3 7.4±0.1 7.8±0.1 33.3±3.4 31.7±2.6 17.5±1.7 16.8±1.3 Minimum -324.5 -403.9 6.6 7.3 17.8 21.1 9.2 10.3 Maximum 120.3 227.4 8.1 8.2 53.0 46.9 26.6 24.3

79

Table 4.1.2: Characteristics of the habitats of the emergent species in the East Kleinemonde Estuary. N/a = Not applicable.

Sediment Water level Water level Sediment organic (cm) (cm) Species moisture matter content (%) (first (second content (%) generation) generation) Supratidal habitat Juncus spp. Mean±SE 17.3±1.3 2.6±0.3 10.4±3.9 Minimum 8.1 1.1 0.0 N/a Maximum 22.4 3.9 31.1 S. virginicus Mean±SE 16.6±1.6 2.3±0.3 35.4±8.3 Minimum 6.4 0.7 0.0 N/a Maximum 22.4 3.9 74.9 S. decumbens Mean±SE 18.1±1.2 2.7±0.3 14.4±5.4 10.6±4.0 Minimum 11.3 1.5 0.0 0.0 Maximum 22.4 4.8 40.8 29.6 Intertidal habitat S. meyeriana Mean±SE 18.2±1.9 3.0±0.3 31.0±7.5 0.5±0.2 Minimum 7.9 1.8 0.0 0.0 Maximum 31.4 5.2 64.0 1.8 S. tegetaria Mean±SE 18.7±1.7 2.6±0.3 27.7±7.1 9.7±4.3 Minimum 9.0 1.6 0.0 0.0 Maximum 31.4 5.2 65.6 36.2 Reed and sedge habitat P. australis Mean±SE 22.2±2.3 4.3±0.4 61.4±9.9 Minimum 5.7 1.8 13.4 N/a Maximum 30.2 6.6 106.1 B. maritimus Mean±SE 23.8±1.4 4.0±0.4 43.2±8.3 Minimum 13.2 1.0 3.6 N/a Maximum 30.2 6.5 83.9

80

Table 4.1.3: Characteristics of the habitat of the submerged species in the East Kleinemonde Estuary. Submerged habitat: Submerged species are plants that are anchored in soft subtidal or low intertidal substrata and adapted to be completely submersed for most states of the tide (Adams, 1994).

Water column characteristics Sediment characteristics Sediment Sediment Redox Electrical Temp- organic Species Salinity Turbidity Water moisture potential pH concutivity erature matter (ppt) (m) level (cm) content (mV) (mS) (°C) content (%) (%) R. cirrhosa Mean±SE 15.9±50.2 7.8±0.1 51.8±2.2 33.0±1.2 21.4±1.4 1.2±0.1 18.0±1.1 2.1±0.2 44.6±8.6 Minimum -295.5 7.2 37.6 23.8 14.6 0.8 8.1 1.1 6.0 Maximum 246.9 8.5 63.3 42.3 28.7 1.7 22.5 3.8 83.9 C. vulgaris Mean±SE 1.8±54.6 7.8±0.1 53.5±2.0 34.0±1.0 20.7±1.5 1.3±0.1 17.6±1.2 2.2±0.3 48.5±8.1 Minimum -295.5 7.2 41.6 30.0 14.6 1.0 8.1 1.1 1.3 Maximum 246.9 8.5 63.3 42.3 28.7 1.7 22.5 3.8 85.7

4.2 MACROPHYTE PHENOLOGY IN A TEMPORARILY OPEN/CLOSED ESTUARY COMPARED WITH A PERMANENTLY OPEN ESTUARY

Figures 4.2.39 to 4.2.40 and Table 4.2.2 provide a summary of the macrophyte growth rate data in both estuaries. The flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical data are provided in Appendix 9.2 per species.

4.2.1 Juncus kraussii and Juncus acutus Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 46.4 ± 7.9% to 15.9 ± 6.5 % with a mean monthly increase of 13 % (Figure 4.2.2a). Change in cover over the entire sampling period was -9.1 %, while the average monthly rate of die-back was -23 % per month during the period of high water level. A significant decrease in live cover occurred from March to April 2009 (t = 3.31; p < 0.05) and from November to December 2009 (t = 3.68; p < 0.05) due to high water level and salinity (Refer Section 4.3.1 and Figure 4.3.1a). A significant increase occurred from January 2010 to February 2010 (t = -2.44; p < 0.05) due to lower water level and summer growth (Refer Section 4.3.1 and Figure 4.3.1a). An increase in plant height was evident from February 2009 to November 2009 and a decline from December 2009 to June 2010, which was due to culm die-back and young, short culms emerging (Figure 4.2.3a). The average monthly increase in height was 2.1 % or 3.1 cm per month. In the Kowie Estuary, the live plant cover ranged from 55.6 ± 8.7% to 41 ± 5.9 % with a mean monthly increase of 6% (Figure 4.2.2a). Live cover was significantly higher in summer compared to winter (t = 4.3; p < 0.05). There was a significant decline in plant height from November to December 2009 (t = 4.37; p < 0.05) (Figure 4.2.3b). The average monthly increase in height was 2.8 % or 3.2 cm per month. At the end of the sampling period, plant cover was significantly higher in the POE compared to the TOCE (t = -3.48; p < 0.05). Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

81

Reproductive phenology

All the J. kraussii inflorescences developed prior to February 2009 with a maximum average of three old inflorescences per m² (Figure 4.2.4a). The second flowering period commenced in November 2009, but inflorescences developed outside of the quadrats and were sparse due to the high water level. Field observations indicated that the peak flowering period was from November to December 2009. The average number of flowers (or fruit) and seeds per inflorescence was 1 344 and 28 224 respectively (Table 4.2.1 and Figure 4.2.1). Seeds developed in January 2010, taking approximately two months to develop. Peak seeding occurred from January 2010 to February 2010 in the surrounding habitat. This conclusion was based on collection of inflorescences and seeds for the germination trials (i.e. no inflorescences developed in the quadrats and therefore no graph was plotted). Seed production was low due to high water level, with capsules containing only one to three fully developed seeds i.e. 1 411 - 3 951 seeds per inflorescence. There were five inflorescences in the surrounding habitat at Transect 1, which amounted to 7 055 – 19 755 seeds in total; or 0.03 inflorescences and 77 seeds per m².

Table 4.2.1: Extrapolation results for the number of flowers, fruit and seed per inflorescence for a J. kraussii plant in the East Kleinemonde Estuary.

Mean no. of rachis st nd Mean no. of 1 Mean no. of 2 Mean no. of flowers on the Mean no. of seed order branches on order branches on (or fruit) on a 2nd inflorescence st per capsule (fruit) the rachis 1 order branch order branch peduncle 12 7 2 8 21

The mean number of flowers or fruit per inflorescence = (12 * 7 * 2) * 8

The mean number of seed per inflorescence = (12 * 7 * 2) * (8 * 21)

Peduncle

Rachis X 12

Flowers/Fruit X 8

2nd order 1st order branch X 2 branch X 12

Figure 4.2.1: Structural components of the J. kraussii paniculate inflorescence. The peduncle holds the entire inflorescence and the rachis is the main stem holding the flowers or branches of the inflorescence.

In the Kowie Estuary, fruiting and seeding phase inflorescences were already present from the first flowering period at commencement of the sampling period. Consequently, peak flowering occurred prior to

82 commencement of the study. The fruiting and seeding phase inflorescences remained on the plants to April 2010 from the first flowering period (Figures 4.2.4b and 4.2.5a). At the onset of the second flowering period in September 2009, a significant increase in J. acutus inflorescences occurred from September 2009 to October 2009 (t = -3.08; p < 0.05), (Figure 4.2.4b). The maximum mean of 41 inflorescences per m² was produced at a rate of 78 %. Peak flowering occurred from September 2009 to October 2009 (Figure 4.2.5b). Seeds developed in November 2009, taking approximately two months to develop (Figures 4.2.5b and 4.2.6a). Maximum fruiting and seeding occurred in March 2010 with a mean of 318 779 seeds ± 76 903 per m² (Figure 4.2.10a). The mean number of seeds per inflorescence was 19 328, which was based on a mean of 177 capsules per inflorescence and 109 seed per capsule. The mean monthly rate of new seed produced was 13.2 %. The mean number of inflorescences and seeds in the POE was significantly higher compared to the TOCE due to prolonged inundation in the TOCE (t = -3.38; -2.39; p < 0.05). Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

Seed viability

No significant differences in monthly percentage germination were evident for J. kraussii from January 2010 (60 %) to February 2010 (72 %). A test for change in viability could not be determined from seeds collected during the first flowering period, suffice to say that germination was high (95 %). Seeds took an average of 9 days to germinate.

J. acutus seeds, which developed in November 2009, showed significant monthly differences in percentage germination from November 2009 to January 2010 (t = 3.29; -3.31; p < 0.05), and from March 2010 to May 2010 (t = 3.06; -7.89; p < 0.05) (Figure 4.2.10b). The significant differences were due to significant declines in germination percentage from November (77 %) to December 2009 (52 %) and from March (88 %) to April 2010 (69 %), the percentage germination increased from November 2009 (77 %) to June 2010 (92 %). However, due to the significant monthly declines in percentage germination, viability did not increase with time. Seeds took an average of 12 days to germinate. Statistical data are provided in Appendix 9.2.

83 (a) Jk

(b) Ja

Figure 4.2.2: Mean monthly percentage cover per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) of J. acutus (Ja) in the Kowie Estuary February 2009 to June 2010 (± SE).

(a) Jk

(b) Ja

Figure 4.2.3: Mean monthly height (cm) per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE).

84

(a) Jk

(b) Ja

Figure 4.2.4: Mean monthly number of inflorescences per m² of (a) J. kraussii (Jk) in the East Kleinemonde Estuary; and (b) J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE). (a) Ja

(b) Ja

Figure 4.2.5: Mean monthly percentage for the phase of the inflorescences per m² during (a) the first flowering period; and (b) the second flowering period for J. acutus (Ja) in the Kowie Estuary from February 2009 to June 2010 (± SE).

85

(a) Ja

(b) Ja

(c) Jk (EK)

Figure 4.2.6: Mean monthly (a) number of seed per m²; and percentage germination and number of days to germination for (b) J. acutus (Ja) seeds harvested from the Kowie Estuary from February 2009 to June 2010 and (c) J. kraussii seeds harvested from the East Kleinemonde Estuary (± SE).

86

4.2.2 Sporobolus virginicus Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 40.6 ±5 % to 0.3 ±0.3 % with a monthly increase in cover of 36 % and decrease of -23 % (Figure 4.2.8a). The lower cover in March 2009 and April 2009 was due to visual error because of inundation (Refer Section 4.3.2 and Figure 4.3.2a). Live plant cover decreased significantly from August 2009 to September 2009 (t = 3.52; p < 0.05), due to inundation from March 2009. A significant decrease in cover occurred from February 2009 to January 2010 (t = 8.97; p < 0.05) due to prolonged inundation. Re-growth occurred in February 2010, increasing with lower water level and decreasing with higher water level. Plant height remained relatively stable to June 2009 before inundation occurred and height could no longer be monitored after inundation occurred (Figure 4.2.9a). The mean monthly increase in height was 16.4 % or 2 cm per month. In the Kowie Estuary, the live plant cover ranged from 10.3 ± 4.6 % to 22.8 ± 8.4 % with a monthly cover increase of 10.5 % (Figure 4.2.8b). Cover declined during autumn and early winter and increased during late winter and summer due to changes in redox potential (Refer to Section 4.3.2 and Figure 4.3.2a). Stem height was significantly lower in March 2009 compared to August 2009 (t = -3.5; p < 0.05) (Figure 4.2.9b). A significant increase occurred from October to November 2009 (t = -5.4; p < 0.05) and from March 2010 to June 2010 (t = -3.99; p < 0.05). Significant increases in stem height were due to higher rainfall events. The mean monthly increase in height was 8.8 % or 1.5 cm per month. Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, the maximum number of inflorescences occurred in February 2009 with a mean of 35 inflorescences per m² (Figure 4.2.10a). Peak flowering had occurred prior to February 2009 and peak seeding was in March 2009 with a mean of 716 ± 354 seed per m² (Figure 4.2.10b and c). The average number of flowers (or fruit) and seed per inflorescence was 52 (Figure 4.2.7). The rate of new seed produced from February to March 2009 was 81 %. Due to inundation in July 2009, no inflorescences could be observed. Since sampling commenced after peak flowering, seed development probably took 1-2 months. During the second flowering period, only a few flowering inflorescences developed in April 2010 and May 2010, but no seeds developed due to prolonged inundation. The rate of new inflorescences produced was 97 %. In the Kowie Estuary, maximum inflorescences occurred in October 2009 with a mean of 101 inflorescences per m² (Figure 4.2.13a). The rate of new inflorescences produced from September 2009 to October 2009 was 100 %. Peak flowering occurred during October 2009 and November 2009 (Figure 4.2.10b). Peak seeding occurred in December 2009 with a mean of 2 428 ± 2192 seeds per m² (Figure 4.2.10c). Seeds took approximately one month to develop. The rate of new seed produced from January to February 2010 was 19.4 %. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

87

a b

Figure 4.2.7: A paniculate inflorescence of (a) S. virginicus changing into an infructescence from which a mean of 52 caryopsis (fruit) will develop and produce one seed each (b).

Seed viability

Seeds were originally harvested from East Kleinemonde Estuary (April 2009 to June 2009) (Figure 4.2.11). However, due to inundation and the lack of seed production in the TOCE, seeds from the Kowie Estuary were harvested (January 2010 to March 2010). Germination ranged from 37 – 73 % from seeds harvested in the TOCE compared to 78.3 – 90 % from seeds harvested in the POE. Mean germination in the TOCE, over the three month period, was 55 % compared to 83.3 % in the POE, but was not significant (t = -2.4; p > 0.05). Significant differences in germination percentage were evident between April 2009 (54 %) compared to both January 2010 (90 %) and February 2010 (78.3 %) (t = -4.1; -5.1; p < 0.05). Germination was significantly higher in January 2010 (90 %) compared to June 2009 (37 %) (t = -7.4; p < 0.05). Statistical data are provided in Appendix 9.2.

Percentage germination ranged from 37 - 90 % for both estuaries. No significant differences in monthly percentage germination were evident from seeds harvested in the TOCE or the POE. Therefore viability did not change over the fruiting period in either estuary once seeds matured (i.e. brown fruit). The mean percentage germination was significantly higher in seeds harvested in 2010 from the POE (85 % ranging from 78 - 90 %) compared to the seeds harvested in 2009 from the TOCE (55 % ranging from 37 - 73 %) (t = -3.5; p < 0.05). Seeds took an average of 11.5 days to germinate.

88 (a)

(b)

Figure 4.2.8: Mean monthly percentage cover per m² of S. virginicus in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.9: Mean monthly height (cm) per m² of S. virginicus in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

89 (a)

(b)

(c)

Figure 4.2.10: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seeds per m² of S. virginicus in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

Figure 4.2.11: Mean monthly percentage germination and number of days to germination of S. virginicus seeds harvested from the East Kleinemonde Estuary, for the period April 2009 to June 2009, and the Kowie Estuary, for the period January 2010 to March 2010 (± SE).

90 4.2.3 Sarcocornia decumbens Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 0.5 ± 0.5 % to 36.3 ± 7.9 % with a monthly increase of 31.9 % (Figure 4.2.13a). This decreased significantly from July 2009 to August 2009 (t = 2.34; p < 0.05) and over the period from August 2009 to June 2010 (t = 2.99; p < 0.05). Plant height increased significantly over the period from February 2009 to July 2009 (t = -6.19; p < 0.05) due to prolonged inundation (Figure 4.2.14a). Refer to Section 4.3.2 and Figure 4.3.2. Second generation seedlings emerged in April 2010 where the water receded. The mean monthly increase in plant height was 12.3 % or 3.2 cm. In the Kowie Estuary, the live plant cover ranged from 55 ± 5.9 % to 72 ± 5.8 % with a monthly increase of 4.3% (Figure 4.2.13b). Significant increases occurred over the period from May to July 2009 (t = -2.66; p < 0.05) and from November 2009 to April 2010 (t = -3.58; p < 0.05) due to higher rainfall and lower pH (Refer Section 4.3.3 and Figure 4.3.3b). Plant height was significantly higher in summer compared to autumn and winter (t = 3.52 and 2.79; p < 0.05) (Figure 4.2.14b). The mean monthly increase in plant height was 9.5 % or 2.5 cm per month. By the end of the sampling period, plant cover in the TOCE was significantly lower compared to the POE (t = - 13.45; p < 0.05). Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, maximum inflorescences occurred in June 2009 with a mean of 720 ± 273 inflorescences per m² increasing at a mean monthly rate of 47.6 % (Figure 4.2.15a). Inundation occurred and inflorescences could no longer be monitored. Peak flowering occurred from February 2009 to March 2009 and peak seeding was in June 2009 with a mean number of 102 847 ± 39 443 seeds per m² (Figure 4.2.15b and c). The mean monthly rate of new seed produced was 37.4 %. Fruit and seeds took approximately one month to develop. The mean number of seed (flowers or fruit) per inflorescence was 163, which was based on a mean of 18 fertile segments per inflorescence and nine seeds (flower or fruit) per fertile segment (Figure 4.2.12). No new inflorescences developed in 2010 because seedlings only germinated in April 2010 after the water receded in February 2010. As a result, the time from seed germination to seed formation could not be determined. In the Kowie Estuary, maximum inflorescences occurred in June 2009 with 347 ± 66 inflorescences per m² and in May 2010 with 203 ± 88 inflorescences per m² (Figure 4.2.14a). Peak flowering occurred before February 2009 and from January to March 2010 during the second flowering period (Figure 4.2.15b). Inflorescences increased at a mean monthly rate of 50.5 %. Peak seeding occurred in June 2009 with a mean of 48 576 seed per m² and in May 2010 with a mean number of 20 661 seed per m² (Figure 4.2.15c). Seeds took approximately three months to develop while the monthly increase in seed was 24.9 %. No seedlings germinated in the Kowie Estuary and the time from seed germination to seed formation could not be determined. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

91

(a) (b)

Fertile segment

Inflorescence

Figure 4.2.12: S. decumbens plants had an average of nine seeds (a) per fertile segment (b) and an average of 18 fertile segments per inflorescence.

Seed viability

Percentage germination for seeds harvested from the East Kleinemonde Estuary was high and ranged from 89 – 99 % (Figure 4.2.16). There were no significant differences in monthly percentage germination from April 2009 to October 2009 (t = 0.9; -0.7; 2.1; -0.9; 1.1; -0.6; p > 0.05). Seeds harvested from the Kowie Estuary showed significant differences in monthly percentage germination from April 2010 to June 2010 (t = -11.2; 4.19; p < 0.05), increasing from April (9 %) to May 2010 (47 %) and decreasing to June 2010 (21 %). Seeds took an average of 6 days to germinate. Mean monthly germination was significantly higher in seeds harvested from the East Kleinemonde Estuary compared to the Kowie Estuary (t = 9.2; p < 0.05). Statistical data are provided in Appendix 9.2.

92 (a)

(b)

Figure 4.2.13: Mean monthly percentage cover per m² of S. decumbens in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.14: Mean monthly percentage cover per m² of S. decumbens in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

93 (a)

(b)

(c)

Figure 4.2.15: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of S. decumbens in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February/May 2009 to June 2010 (± SE).

Figure 4.2.16: Mean monthly percentage germination and number of days to germination of S. decumbens seeds harvested from the East Kleinemonde (EK) Estuary (for the period April 2009 to October 2009) and the Kowie (KW) Estuary (for the period April 2010 to June 2010) (± SE).

94

4.2.4 Sarcocornia hybrid

Growth phenology

In the Kowie Estuary, the live plant cover ranged from 37.8 ± 8.8 % to 9.6 ± 2.6 % with a monthly increase of 12.5 % (Figure 4.2.17a). Cover declined significantly from February to August 2009 (t = 3.47; p < 0.05), increased significantly from August 2009 to September 2009 (t= -2.27; p < 0.05). Cover declined significantly over the period from September 2009 to June 2010 (t = -2.27; p < 0.05). Plant cover decreased in response to a reduction in redox potential and low temperatures (Refer Section 4.3.3 and Figure 4.3.3). Plant height decreased significantly from February 2010 to April 2010 (t = 3.07; 7.68 p < 0.05) due to lower temperatures (Figure 4.2.17b and refer Section 4.3.3 and Figure 4.3.3). Seedlings (second generation plants) germinated in July 2009 and expanded to June 2010 while seedling height increased to February 2010, with a significant increase from January to February 2010 due to higher temperatures (t = -4.87; p < 0.05). The mean monthly increase in plant height was 18.4 % or 1.3 cm per month. At the end of the sampling period, the plant cover of S. decumbens was significantly higher compared to the Sarcocornia hybrid in the Kowie Estuary (t = -4.13; p < 0.05), despite both growing in the supratidal habitat. Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

Seedlings germinated in situ and took approximately seven months to develop flowering inflorescences and nine months to develop seed, while seed took two months to develop after flowering. Maximum inflorescences were produced in May 2009 with 2 188 ± 693 inflorescences per m² and in April 2010 with 398 ± 270 inflorescences per m² (Figure 4.2.18a). The mean monthly rate of new inflorescences produced was 50.5 %. Peak flowering occurred in February 2009 and in February to March 2010 (Figure 4.2.18b). Peak seeding was in April 2009 with 129 953 ± 48 613 seeds per m² and April 2010 with 41 526 ± 29 882 seeds per m² (Figure 4.2.18c). The rate of new seed produced was 82 % per month. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

95 (a)

(b)

Figure 4.2.17: Mean monthly (a) percentage cover; and (b) height (cm) per m² of the Sarcocornia hybrid in the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a) (b)

(c) Figure 4.2.18: Average monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of the Sarcocornia hybrid in the Kowie Estuary (KW) from February

2009 to June 2010 (± SE).

96 4.2.5 Salicornia meyeriana Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 0 to 12.5 ± 2.7 % and expanded at a mean monthly rate of 43.6 % (Figure 4.2.19a). In July 2009 complete die-back had occurred and all the quadrats were inundated to June 2010. Seedlings (second generation plants) emerged in January 2010 where the water receded and cover expanded to May 2010. The mean monthly increase in plant height was 43 % or 4 cm per month (Figure 4.2.20a). In the Kowie Estuary, the live plant cover ranged from 20.5 ± 4.1 % to 0.2 ± 0.2 % expanding at a mean monthly rate of 30.4 % (Figure 4.2.19b). Complete die-back of first generation plants occurred in May 2009 at which time seedlings (second generation plants) germinated. Cover declined after February 2010 due to die-back. Third generation seedlings germinated in May 2010. The mean monthly increase in plant height was 26 % or 1.9 cm per month (Figure 4.2.20b). Cover in the TOCE was significantly lower compared to the POE, i.e. 5.7 ± 2.7 to 7 ± 2.2 % % (n = 22) in 2009 versus 16.7 ± 4.2 % to 22.3 ± 5 % (n = 11) in 2010 (t = -3.2; t = -3.9; p < 0.05). Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, maximum inflorescences developed in April 2009 with 439 ± 479 inflorescences per m² and in June 2010 with 553 ± 284 inflorescences per m² (Figure 4.2.21a). The mean monthly rate of inflorescence increase was 42.4 %. Peak flowering occurred in March 2009 and April 2010 (Figure 4.2.21b). Peak seeding occurred in April 2009 with a mean of 24 050 ± 25 614 seeds per m² and May 2010 with a mean of 27 643 ± 1 604 seeds per m² (Figure 4.2.21c). The mean number of seed (flowers or fruit) per inflorescence was calculated at 107, which was based on a mean of 18 fertile segments per inflorescence and 6 seeds, flower or fruit per fertile segment (Figure 4.2.12). Seed production increased by 100 % per month. Inflorescences took approximately 2.5 months and seed approximately three months to develop from the time of germination and seed approximately 0.5 months after flowering. Seed that developed in April 2009 prior to inundation germinated after nine months where the water had receded. In the Kowie Estuary, the maximum number of inflorescences were recorded in May 2009 with 4 008 ± 2 423 inflorescences per m² and in March 2010 with 6 275 ± 1816 inflorescences per m² (Figure 4.2.21a). The mean monthly rate of inflorescence increase was 41.5 %. The number of inflorescences were significantly higher in the Kowie Estuary compared to the East Kleinemonde Estuary in 2009 (t = 3.34p < 0.05). Peak flowering occurred in February 2009 and in February 2010 (Figure 4.2.21b). Peak seeding occurred in April 2009 with a mean of 264 224 ± 103 929 seeds per m² and May 2010 with a mean of 640 292 ± 194 571 seeds per m² (Figure 4.2.21c). Seed production increased by 100 % over one month. Inflorescences took nine months and seed approximately ten months to develop from the time of germination. Seeds took approximately one month to develop after flowering. The number of inflorescences and seeds were significantly higher in the POE compared to the TOCE (t = 3.34; 3.44; p < 0.05). Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

Seed viability

There was a significant decline (t = 2.95; 11.85; p < 0.05) in the percentage germination of seeds from May 2009 (70 %) to June 2009 (26 %) after two months from maturation in April 2009; and from May 2010 (80 %) to June 2010 (29 %) after one month from maturation in May 2010 (Figure 4.2.22). Viability therefore declined after maturation. Seeds took an average of 5 days to germinate. Statistical data are provided in Appendix 9.2.

97 (a)

(b)

Figure 4.2.19: Mean monthly percentage cover per m² of S. meyeriana in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.20: Mean monthly height (cm) per m² of S. meyeriana in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

98 (a)

(b)

(c)

Figure 4.2.21: Mean monthly (a) number of inflorescences, (b) phase of inflorescences; and (c) number of seed per m² of S. meyeriana in the East Kleinemonde Estuary Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

Figure 4.2.22: Mean monthly percentage germination and number of days to germination of S. meyeriana seeds harvested from the East Kleinemonde (EK) Estuary (± SE).

99 4.2.6 Sarcocornia tegetaria Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 0.09 ± 0.1 % to 21.5 ± 5.9 % expanding at a mean monthly rate of 46 % (Figure 4.2.23a). Plant cover decreased significantly from February 2009 to December 2009 (t = 3.75; p < 0.05) due to prolonged inundation (Refer to Section 4.3.5 and Figure 4.3.5a). Plant height increased significantly from February to March 2009 (t = -2.49; p < 0.05) and decreased significantly from March 2009 to April 2009 (t = 5.86; p < 0.05) due to inundation (Refer to Section 4.3.5 and Figure 4.3.5a) (Figure 4.2.24a). Re-growth occurred from April 2010. Seedlings (second generation plants) germinated in January 2010 where the water receded, while third generation seedlings emerged in June 2010. A new generation therefore emerged after five months. The mean monthly increase in plant height was 31.7 % or 4.8 cm per month. In the Kowie Estuary, the live plant cover ranged from 43.7 ± 10.8 % to 58.4 ± 8.4 % expanding at a mean monthly rate of 5.2 % (Figure 4.2.23b). Cover increased significantly from April 2009 to May 2009 (t = -3.09; p < 0.05) due to an increase in water salinity (Refer to Section 4.3.5 and Figure 4.3.5b). Plant height increased significantly from February 2009 to July 2009 (t = - 2.87; p < 0.05) and from June 2009 to July 2009 (t = -2.76; p < 0.05) due to an increase in rainfall (Refer to Section 4.3.5 and Figure 4.3.5b) (Figure 4.2.24b). The mean monthly increase in plant height was 5.7 % or 0.8 cm per month. Plant cover was significantly higher in the POE compared to the TOCE in June 2010 (t = -4.13). Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, the maximum number of inflorescences occurred in May 2009 with 409 ± 341 inflorescences per m² and in May 2010 with 229 ± 166 inflorescences per m² (Figure 4.2.28a). The mean monthly rate of inflorescence increase was 99.8 %. Peak flowering occurred in February and March 2009 and March and April 2010 (Figure 4.2.25b). Peak seeding occurred in May 2009 with a mean of 45 542 ± 39 731 seeds per m² and in June 2010 with a mean of 7 001 ± 4 828 seeds per m² (Figure 4.2.25c). The average number of seed (flowers or fruit) per inflorescence was calculated at 116, which was based on a mean of 19 fertile segments per inflorescence and 6 seeds (flower or fruit) per fertile segment (Figure 4.2.12). Inflorescences took two months and seeds four months, to develop from the time of germination and seed approximately two months after flowering. Seeds that developed in April 2009 prior to inundation germinated after nine months where the water had receded. Seeds germinated one month after developing in May 2010. In the Kowie Estuary, maximum inflorescences occurred in June 2009 with 200 ±122 inflorescences per m² and in February 2010 with 69 ± 50 inflorescences per m² (Figure 4.2.25a). The mean monthly rate of inflorescence increase was 53.8 %. Peak flowering occurred in February and March 2009 and from January to March April 2010 (Figure 4.2.25b). Peak seeding occurred in June 2009 with a mean of 16 958 ± 10 772 seeds per m² and in May 2010 with a mean of 1 587 ± 1 275 seeds per m² (Figure 4.2.25c). Maximum seed numbers were significantly higher in the East Kleinemonde Estuary compared to the Kowie Estuary in 2010 (t = 2.2; p < 0.05). No seeds germinated in situ in the Kowie Estuary. Seeds took approximately two months to develop after flowering. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

Seed viability

Percentage germination ranged from 24 to 66.5 % (Figure 4.2.26). A significant decline (t = 3.41; p < 0.05) in percentage germination occurred from May 2009 (66 %) to June 2009 (38 %), possibly due to fungi development, seed age and/or unviable seeds. Seeds took an average of 4.8 days to germinate. Statistical data are provided in Appendix 9.2.

100 (a)

(b)

Figure 4.2.23: Mean monthly percentage cover per m² of S. tegetaria in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.24: Mean monthly height (cm) per m² of S. tegetaria in the (a) East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

101 (a)

(b )

(c )

Figure 4.2.25: Mean monthly (a) number of inflorescences, (b) phase of inflorescences and (c) number of seed per m² of S. tegetaria in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

Figure 4.2.26: Mean monthly percentage germination and number of days to germination of S. tegetaria seeds harvested from the East Kleinemonde (EK) Estuary (± SE).

102

4.2.7 Phragmites australis

Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 3.6 ± 3.6 % to 17.1 ± 3.4 % expanding at a mean monthly rate of 21.2 % (Figure 4.2.28a). Cover declined significantly over the period from February 2009 to August 2009 (t = 3.18; p< 0.05) due to an extended period of high water level and salinity (Refer to Section 4.3.6 and Figure 4.3.6a). Culm height increased from February 2009 to November 2009, with significant increases from March 2009 to April 2009 (t = 3.32; p< 0.05) and from June 2009 to July 2009 (t = -5.17; p< 0.05) (Figure 4.2. 29a). Plants were grazed by cows in December 2009, reducing plant height. Due to the significant height increase from March 2010 to June 2010 (t = -2.35; p < 0.05), culms would have most likely continued to increase due to the prolonged high water level. The mean monthly rate of increase in height was 14.8 % or 18.2 cm. In the Kowie Estuary, the live plant cover ranged from 21.8 ± 6.5 % to 7.8 ± 3.2 % expanding at a mean monthly rate of 14.3 % (Figure 4.2. 28b). Plant cover increased from spring to autumn, with significant monthly increases from October 2009 to December 2009 (t = -2.77; -3.35; p < 0.05) due to higher temperature (Refer to Section 4.3.6 and Figure 4.3.6b). Plant cover declined significantly from April 2010 to May 2010 (t =3.11; p < 0.05). Summer and autumn cover were significantly higher compared to winter cover (t = 3.26; 3.94; p < 0.05). Young culms emerged from April 2009 through to September 2009 (autumn to spring). A significant increase in plant height occurred from October 2009 to December 2009 (t = -4.02; t = - 2.68 p < 0.05), from January 2010 to February 2010 (t = -3.86; p < 0.05) and from March 2010 to April 2010 (t = -1.99; p < 0.05) due to temperature and pH increases (Refer to Section 4.3.6 and Figure 4.3.6b) (Figure 4.2.29b). The mean monthly rate of increase in height was 9.6 % or 9.1 cm. Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, inflorescences developed outside of the quadrats from May 2009 to July 2009, and from April 2010 to June 2010. Abundance was extremely low during the second flowering period compared to the first, with only 5 inflorescences or 520 seeds produced. Peak flowering occurred during May 2009 and April 2010, while peak seeding occurred in June 2009 and May 2010. Seeds took approximately one month to develop. In the Kowie Estuary, inflorescences developed outside of the quadrats from April 2009 to July 2009 and in the quadrats from March 2010 to June 2010. Maximum inflorescences occurred in May 2010 with 1 ± 0.6 inflorescences m-2 (Figure 4.2.30a). The mean monthly rate of inflorescence production was 13.3 %. Peak flowering occurred in April 2009 and April 2010 (Figure 4.2.30b). Peak seeding occurred in May 2009 and May 2010 with a mean of 12 ± 6 seeds m-2 (Figure 4.2.30c). The P. australis stand was approximately 900 m² in extent and therefore 10 800 seeds were produced. Seeds took approximately one month to develop. Seeds increased from April 2010 to May 2010 by 100 %. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

The potential average number of seeds per inflorescence was 3 187, which was based on the average number of 27 rachis per inflorescence, 59 spikelets per rachis and 2 flowers (fruit or seeds) per spikelet (Figure 4.2.27).

103 Seed recovered from inflorescences in the East Kleinemonde and Kowie estuaries was negligible. Based on the number of seeds recovered from inflorescences in both estuaries over the entire study period, the average number of seeds was 104 seeds per inflorescence or 3.3 % of the potential average number of seed per inflorescence (3 187). Flowering data are summarized in Tables 4.2.4 to 4.2.6.

Peduncle

Rachis (x 27)

Spikelets (x 59) (b) (a)

Figure 4.2.27: P. australis panicle (inflorescence) with (a) an average of 27 rachis per peduncle, 59 spikelets per rachis and 2 flowers/fruits per spikelet; and (b) a caryopsis (fruit).

Seed viability

Very little seed was recovered from P. australis and therefore seed viability was not adequately tested. However, of the 22 seeds recovered from the East Kleinemonde Estuary in June 2009 and June 2010, and from the Kowie Estuary in June 2010, a mean of 49.1 % seeds germinated. Fungi developed rapidly on the seeds, which probably influenced germination. Seeds took an average of 6 days to germinate.

104

Figure 4.2.28: Mean monthly percentage cover per m² of P. australis in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

Figure 4.2.29: Mean monthly plant height (cm) per m² of P. australis in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

105

(a)

(b)

(c)

Figure 4.2.30: Mean monthly (a) number of inflorescences, (b) phase of inflorescences, and (c) number of seeds per m² of P. australis in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

106

4.2.8 Bolboschoenus maritimus

Growth phenology

In the East Kleinemonde Estuary, the live plant cover ranged from 0 % to 29.6 ± 8.4 %, declining significantly from February 2009 to August 2009 (t = 3.57; p < 0.05) at a mean monthly rate of -11.3 % due to extended saline and high water levels (Refer to Section 4.3.7 and Figure 4.3.7a) (Figure 4.2.32a). Re-growth occurred in June 2010 in 29 cm of water, but plants established in the supratidal zone outside of the quadrats from February 2010 and expanded to June 2010 where the water had receded. Plant height increased from February 2009 to July 2009, increasing significantly from June 2009 to July 2009 (t = -3.46; p < 0.05) due to prolonged high water levels (Refer to Section 4.3.7 and Figure 4.3.7a) (Figure 4.2.33a). The mean monthly increase in height was 27.5 % or 9.7 cm. In the Kowie Estuary, the live plant cover ranged from 3.5 ±1.1 % to 38.5 ±2 % expanding at a mean monthly rate of 36.3 % (Figure 4.2.32b). Significant monthly decreases occurred from February 2009 to April 2009 (t = 4.52; t = 3.83; p < 0.05). New shoots developed from April 2009 to October 2009 and again in March 2010. Significant monthly increases occurred from October 2009 to December 2009 (t = -5.03; -3.27; p < 0.05) in spring and summer, whilst a significant decline occurred from March 2010 to April 2010 (t = 3.01; p < 0.05). Live culm cover was significantly higher in summer than in autumn and winter (t = 3.52; 6.54; p < 0.05). Plant height generally decreased from February 2009 to August 2009 in autumn and winter, with a significant decline from March 2009 to April 2009 as new shorter shoots emerged (Figure 4.2.33b). A significant increase occurred in spring from August 2009 to September 2009 (t = - 2.84; p < 0.05). Height decreased from February 2010 to June 2010 during autumn and early winter due to die- back and new shorter shoots emerged. Culm height was significantly higher in summer than in autumn and winter (t = 5.87; 7.58; p < 0.05). The mean monthly increase in height was 7.4 % or 5.5 cm. Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

In the East Kleinemonde Estuary, maximum inflorescence occurred in February 2009 with 54 ± 33 inflorescences m-2 (Figure 4.2.34a). The lower value in April 2009 is most likely due to error. Inflorescences developed outside of the quadrats from April 2010 to June 2010 in the supratidal areas, but abundance was low. Peak flowering occurred prior to February 2009 and peak seeding was in May 2009 with a mean of 11 221 ± 5 276 seeds m-2. The average number of seeds (flowers or fruit) per inflorescence was calculated at 232, based on an average of 8 spikelets per inflorescence and 30 seeds (flowers or achenes) per inflorescence (Figure 4.2.31). The mean monthly increase in seed production was 22.8 %. Flowering inflorescences developed outside of the permanent quadrats from March 2010 to June 2010 and seed developed in May 2010. Seeds took two months to develop. In the Kowie Estuary maximum inflorescence occurred in March 2009 with 14 ± 4 inflorescences m-2 and in February 2010 with 12 ± 5 inflorescences per m² (Figure 4.2.34a). The mean monthly increase in inflorescences was 65.5 %. Peak flowering had occurred prior to February 2009 and occurred in February 2010 during the second flowering period (Figure 4.2.34b). Peak seeding occurred in March 2009 with a mean of 3 295 ± 976 seeds per m² and in January 2010 with a mean of 200 ± 82 seeds per m² (Figure 4.2.34c). Seed was significantly lower in January 2010 (t = 3.4; p < 0.05). Peak seeding occurred 107 prior to the peak flowering period because the majority of the flowers that developed in February 2010 did not produce fruit or seeds. As a result, seed developed from flowering inflorescences that developed in October 2009 and November 2009. The mean monthly increase in seed production was 64.9 % and seeds took approximately 2-3 months to mature. Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

(a) (b)

Figure 4.2.31 (a) The flowering spikelets of B. maritimus and (b) seedlings germinating from an achene (fruit), which contains one seed.

Seed viability

Percentage germination ranged from 0 to 36 % (Figure 4.2.35a). Seeds that were harvested from old inflorescences (which developed prior to February 2009) and from seeds that developed in May 2010 and June 2010 did not indicate significant monthly differences in percentage germination. Mature brown seeds were harvested from the Kowie Estuary in January 2010 and May 2010 due to inundation in the East Kleinemonde Estuary (Figure 4.2.35b). Plants in the Kowie Estuary did not produce adequate numbers of seeds from February 2010 to April 2010 hence the lack of germination trial results between January 2010 and May 2010. Seed germination was significantly higher in January 2010 (46 %) and May 2010 (26.4 %) for the Kowie Estuary seeds compared to seeds harvested from the East Kleinmonde Estuary in May 2010 (5 %) (t = 4.3; p < 0.05) and June 2010 (11 %) (t = -3.7; p < 0.05) respectively. Seed germination was also significantly higher in May 2010 (26.4 %) for the Kowie Estuary seeds compared to seeds harvested from the East Kleinmonde Estuary in May 2010 (5 %) (t = -5.6; p < 0.05). Seed germination was substantially higher in September 2009 (36 %) compared to the previous months (range 0 – 11 %) from seed harvested in the TOCE. Seeds took an average of 37.8 days to germinate in the TOCE compared to 15.5 days in the POE. Statistical data are provided in Appendix 9.2.

108

(a)

(b)

Figure 4.2.32: Mean monthly percentage cover per m² of B. maritimus in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.33: Mean monthly height (cm) per m² of B. maritimus in (a) the East Kleinemonde Estuary (EK); and (b) the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

109 (a)

( b)

(c)

Figure 4.2.34: Mean monthly (a) number of inflorescences, (b) phase of inflorescences and (c) number of seed per m² of B. maritimus in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW) from February 2009 to June 2010 (± SE).

(a) (b)

60 50

40

30 20

10

Figure 4.2.35: Mean monthly percentage germination and number of days to germination of B. maritimus seeds harvested from the (a) East Kleinemonde Estuary Estuary (EK); and (b) the Kowie Estuary (KW) during 2010 (± SE).

110

4.2.9 Ruppia cirrhosa and Chara vulgaris

Growth phenology

R. cirrhosa biomass ranged from 2.3 ± 0 to 2 135 ± 1.1 g DW m-2 and expanded at a mean monthly rate of 43.7 % (Figure 4.2.36a). Significant monthly increases occurred from April 2009 to May 2009 and from August 2009 to November 2009 (t = -3.84; -5.2; -3.39; -3.4; p < 0.05) while a significant decline from January 2010 to February 2010 (t = 4.72; p < 0.05), in response to water level and pH (Refer to Section 4.3.8 and Figure 4.3.8a). Although R. cirrhosa died back completely within the areas sampled, plants grew near the estuary banks up until June 2010. Significantly monthly increases in height occurred from August 2009 to October 2009 (t = -7.03; -4.26; p < 0.05), while significant decreases occurred from December 2009 to February 2010 (t = 2.93; 5.27; p < 0.05) (Figure 4.2.37a). The mean monthly increase in height was 33.4 % or 8.8 cm. C. vulgaris biomass ranged from 16.7 ± 0 to 50.3 ± 0.2 g DW m-2, expanding at a mean monthly rate of 35.2 % (Figure 4.2.36b). No data were collected in July 2009 as cover was insignificant. A significant increase occurred from September 2009 to October 2009 (t = -4.21; p < 0.05) and a significant decline occurred from January 2010 to February 2010 (t = 0.28; p < 0.05) in response to water level and water temperature (Refer to Section 4.3.8 and Figure 4.3.8b). The mean monthly increase in height was 30.4 % or 2.9 cm (Figure 4.2.37b). Growth data are summarized in Figure 4.2.39 to 4.2.40 and Tables 4.2.1 to 4.2.3. Statistical growth data are provided in Appendix 9.2.

Reproductive phenology

R. cirrhosa flowering inflorescences and fruiting/seeding inflorescences took four and five months respectively to develop after germination in situ. Flowering commenced in September 2009 and peak flowering occurred in November 2009 with a mean of 16 193 flowers and 8 097 ±1 109 flowering inflorescences m-2 and increasing at a mean monthly rate of 74 % (Figure 4.2.38a). A significant increase in flowers and seeds occurred from September 2009 to October 2009 (t = -3.9, -5.2; p < 0.05). Peak fruiting and seeding also occurred in November 2009 with a mean of 6 064 ± 793 fruiting inflorescences and 26 242 ± 10 755 fruit/seeds m-2 and increasing at a mean monthly rate of 88.4 % (Figure 4.2.38b). Seeds took approximately one month to develop after flowering. Flowering and fruiting inflorescences were present till June 2010 in the stands along the estuary banks. C. vulgaris oogonia and antheridia took three months to develop after germination in situ (Figure 4.2.38a). Peak oogonia and antheridia production was in November 2009 with a mean of 196 998 ± 48 003 oogonia and 46 016 ± 11060 antheridia m-2 increasing at a mean monthly rate of 80.3 % and 60.8 % respectively (Figure 4.2.38b). Flowering data are summarized in Figure 4.2.41 and Tables 4.2.3 to 4.2.4. The statistical flowering data are provided in Appendix 9.2.

Seed viability

R. cirrhosa seeds did not germinate. Viability tests for C.vulgaris were not conducted.

111 (a)

(b)

Figure 4.2.36: Mean monthly biomass (g DW) per m² of R. cirrhosa; and (b) C. vulgaris in the East Kleinemonde Estuary (EK) from February 2009 to June 2010 (± SE).

(a)

(b)

Figure 4.2.37: Mean monthly height (cm) per m² of R. cirrhosa; and (b) C. vulgaris in the East Kleinemonde Estuary (EK) from February 2009 to June 2010 (± SE).

112

(a) (b)

(c) (d)

(e ) Figure 4.2.38: Average monthly (a) number of R. cirrhosa flowers and flowering inflorescences; (b) C. vulgaris oogonia; (c) number of R. cirrhosa

fruiting and seeding inflorescences; (d) C. vulgaris antheridia; and (e) number of R. cirrhosa seeds per m² in the East Kleinemonde Estuary from February 2009 to June 2010 (± SE).

113 The figures 4.2.39 and 4.2.40 below are a summary of the significant growth periods (p < 0.05) and periods of cover decline (or die-back), as represented in the preceding species results and graphs on macrophyte growth phenology. These periods are based on the mean monthly percentage cover m-², which was used to determine the maximum growing periods/season for open mouth recommendations.

Figure 4.2.39: Summary of growth periods and periods of decline for the selected species in the East Kleinemonde Estuary.

= No significant growth or decline = Growth = Significant growth (p < 0.05) within the growth period = Decline

Su = Summer. Sp = spring. Au = Autumn. Wi = Winter.

Jul

Apr Oct Apr

Jun - Jan Jun

Feb Feb

Mar Mar

Aug Sep Nov Dec

- May - - May

- - -

- - - -

- - - - -

Month -

09

09 09 10

09 10 10

09 09 10 10

09 09 09 09

09 10

Season Su Au Au Au Wi Wi Wi Sp Sp Sp Su Su Su Au Au Au Wi J. kraussii S. virginicus S. decumbens S. meyeriana S. tegetaria P. australis B. maritimus R. cirrhosa

C. vulgaris

Figure 4.2.40: Summary of growth periods and periods of decline for the selected species in the Kowie Estuary.

= No significant growth or decline = Growth = Significant growth (p < 0.05) within the growth period = Decline

Su = Summer. Sp = spring. Au = Autumn. Wi = Winter.

Jul

Apr Oct Apr

Jun - Jan Jun

Feb Feb

Mar Mar

Aug Sep Nov Dec

- May - - May

- - -

- - - -

- - - - -

Month -

09

09 09 10

09 10 10

09 09 10 10

09 09 09 09

09 10

Season Su Au Au Au Wi Wi Wi Sp Sp Sp Su Su Su Au Au Au Wi J. acutus S. virginicus S. decumbens Sarcocornia hybrid S. meyeriana S. tegetaria P. australis B. maritimus

114 Table 4.2.2 below provides a summary of the results discussed in the preceding sections. Growth rate was determined by measuring percentage plant cover and plant height (cm) per 1 m² (n = 10). As a result, the rate of change in plant height has been recorded as a mean percentage, for comparative purposes, and in mean centimeters as it is more meaningful for plant height. The mean rate of expansion or growth was calculated for those periods in which the macrophytes demonstrated growth during the 17 month sampling period. The mean rate of decline in plant cover, not height, was calculated for those periods in which the macrophytes demonstrated die-back during the 17 month sampling period.

Table 4.2.2: Mean monthly change in percentage cover and height of the selected macrophytes within 1 m2 quadrats in the East Kleinemonde Estuary (EK) and the Kowie Estuary (KW). ND = No data. n = 10.

Growth rate Rate of decline Growth rate Growth rate (cover) (cover) (height) (height) Species % change % change % change cm change

EK KW EK KW EK KW EK KW

Supratidal habitat

Juncus kraussii (EK) 13 6 - 7 -12.8 2.1 2.8 3.1 3.2 Juncus acutus (KW)

Sporobolus virginicus 36 10.5 -130.7 – 17.4 16.4 8.8 2 2

Intertidal habitat

Sarcocornia decumbens 31.9 4.3 –137.8 -2.6 12.3 9.5 3.2 2.5

Sarcocornia Hybrid ND 12.5 ND -23.5 ND 18.4 ND 1.3

Salicornia meyeriana 43.6 30.4 -35.5 -15.76 43 26 4 1.9

Sarcocornia tegetaria 45.4 5.2 - 1 807 – 8.4 31.7 5.7 4.8 0.8

Reed and sedge habitat

Phragmites australis 21.2 14.3 -42.3 – 20.5 14.8 9.6 18.2 9.1

Bolboschoenus maritimus ND 36.3 -11.3 -95.8 27.5 7.4 9.7 5.5

Submerged habitat

Ruppia cirrhosa 43.7 ND - 88.1 ND 33.4 ND 8.8 ND

Chara vulgaris 35.2 ND -88.1 ND 30.4 ND 2.9 ND

115 Figure 4.2.41: Phenogram of the reproductive cycles (including seed release periods) of the selected species in the East Kleinemonde Estuary (EK) and Kowie Estuary (KW).

= Peak flowering period = Fruiting and seeding period = Peak seeding period (part of the fruiting/seeding period) = Seed release period

Su = Summer. Sp = spring. Au = Autumn. Wi = Winter.

09 10

09 09 09 09

09 09 10 10

09 09 10 10

09 10

- 09 -

- -

- -

- - - -

- - - -

- -

General flowering period -

Jul

Apr Oct Apr

Jun Jan Jun

Feb Mar Aug Sep Nov Dec Feb Mar

May May Month (Literatures sources)

Season Su Au Au Au Wi Wi Wi Sp Sp Sp Su Su Su Au Au Au Wi

Supratidal habitat

J. kraussii EK Early spring while fruit ripen by mid- summer (Jones and Richards, J. acutus KW 1954). Nov-Feb

S. virginicus EK Throughout year. Peak: Su-Au

S. virginicus KW (Eleuterius & Caldwell, 1984)

S. decumbens EK November – July

S. decumbens KW (Steffen et al., 2008)

Sarcocornia hybrid KW

Intertidal habitat

S. meyeriana EK Su – Au (Davy, 2001;

S. meyeriana KW Allanson, 2000)

S. tegetaria EK January – July

S. tegetaria KW (Steffen et al., 2008)

Reeds and sedges

P. australis EK Su–Au (Haslam, 1972; Auld &

P. australis KW Medd, 1987; Boedeltjie et al., 2004)

B. maritimus EK Su (October-March) (Wilman, 2006

B. maritimus KW Diggory & Parker, 2010)

Submerged species

R. cirrhosa Summer – Autumn

C. vulgaris (Setchell, 1924; Verhoeven, 1979)

116 Table 4.2.3: Mean monthly rate of new reproductive output (percentage), including the mean number of inflorescences and seed produced m-2 (±SE) per flowering period in the selected species of the East Kleinemonde and Kowie estuaries. ND = No data. O = Oogonia. 100 % = Seed increase occurred over a period of one month (i.e. the first month of developing).

Rate of new Rate of new Mean no. of inflorescences m-2 Mean no. of reproductive reproductive (maximum inflorescence seeds m-2 output (mean output (mean production per flowering period) Species no. of no. of seed, (peak seeding per flowering period) inflorescences, %) %)

EK KW EK KW EK KW EK KW

J. kraussii (EK) ND 78 ND 13.2 0.03 * 41 ±9.9 77 * 318 779 ± 76 903 J. acutus (KW)

S. virginicus 97 12.8 81 19.4 135 ±11 101 ±78 716 ± 354 2 428 ± 2192

S. decumbens 347 ±66 to 102 847 ±39 443 48 576 ± 9 143 to 47.6 38.8 37.4 24.9 723 ±273 203 ±88 20 661 ±9 584

Sarcocornia Hybrid 2 188 ±693 to 129 953 ±48 613 to ND 50.6 ND 82 ND ND 398 ±270 41 526 ±29 882

S. meyeriana 439 ±479 to 4 008 ± 2 423 to 24 050 ±25 614 to 264 224 ± 103 929 to 42.4 41.5 100 100 553 ±284 6 275 ± 1816 27 643 ±16 042 640 292 ± 194 571

S. tegetaria 409 ± 341 to 200 ±122 to 45 542 ±39 731 to 16 958 ± 10 772 to 99.8 53.8 76.2 81.4 229 ± 166 69 ±50 7 001 ±4 828 1 587 ± 1 275

P. australis ND 13.3 ND 100 ND 1 ±0.6 ND 12 ±6

B. maritimus 14 ±4 to 3 295 ± 976 to ND 41.5 22.8 65.5 54 ±33 11 221 ±5 276 12 ±5 200 ±82

R. cirrhosa 74 ND 88.4 ND 8 097 ±3 509 ND 26 242 ±10 755 ND

C. vulgaris 196 998 ±48 003 ND 80.3 ND 60.8 ND NA ND (O)

* No Standard Error provided as the mean value was based on five inflorescences surrounding transect one (± 300 m2). .

117 Table 4.2.4: Flowering characteristics of the selected species in the East Kleinemonde and Kowie Estuaries. N/a = Not applicable. ND = No data. TOCE = Temporarily open/closed estuary. POE = Permanently open estuary.

Species Average rachis, branches Average no. of Average Average no. of Average no. Average no. and fertile segments per flowers/fruit no. of seeds per of months to of months to inflorescence etc per seeds per inflorescence develop develop inflorescence fruit viable seeds viable seeds after after germination flowering

J. kraussii 12 rachis per inflorescence; 1 613 21 33 787 ND 2 7 X 1st order branches per rachis; 2 X 2nd order branches per 1st order branch and a cluster 8 flowers or fruit per 2nd order branch

J. acutus N/a 177 109 19 328 ND 2

S. virginicus N/a 52 1 52 ND 2

S. decumbens 18 fertile segments per 152 1 152 ND POE: 2-3 inflorescence

S. meyeriana 18 fertile segments per 107 1 107 TOCE: 3 1 inflorescence POE: 9

S. tegetaria 19 fertile segments per 116 1 116 TOCE: 4 2 inflorescence POE: 9 (S. hybrid)

P. australis 27 rachis per inflorescence, 3 187 1 3 187 ND 1 59 spikelets per rachis and 2

flowers or fruit per spikelet

B. maritimus 8 spikes per inflorescence 230 1 230 ND TOCE: 2

POE: 2-3

R. cirrhosa N/a 2 flowers 1 8 5 1

8 fruit

C. vulgaris N/a ND ND ND 3 ND

118 4.3 MULTIVARIATE ANALYSIS: ENVIRONMENTAL VARIABLES AND MACROPHYTE PHENOLOGY

The following section provides the results of the multivariate analysis (Principle Component Analysis). Multivariate analysis determined whether there was a correlation between the species data (response variables) and environmental data (predictor variables). The analyses also identified the significant variables that influenced macrophyte phenology.

4.3.1 Juncus kraussii and Juncus acutus

In the East Kleinemonde Estuary, the first canonical axis (horizontal) described 49 % and the second canonical axis (vertical) described 81 % of the variation of the species – environment relation (Figure 4.3.1 and Table 4.3.1). The most significant environmental factors influencing J. kraussii were pH (-0.393) and moisture content (-0.215). High pH, sediment electrical conductivity and water level caused an increase in dead plant cover. Sediment moisture and organic content were positively related with an increase in live plant material while salinity was negatively related with an increase in plant cover. All inflorescences in the quadrats were old and high water level was related to the decay of inflorescences. For J. acutus in the Kowie Estuary, the first canonical axis (horizontal) described 98 % of the variation of the species – environment relation (Figure 4.3.1 and Table 4.3.2). The most important environmental factor influencing live plant cover and height was temperature (0.471). The number of inflorescences and seeds increased with low pH whereas dead plant cover was correlated with an increase in pH. An increase in the number of inflorescences and seeds during the second flowering period was related to an increase in sediment electrical conductivity and salinity.

4.3.2 Sporobolus virginicus

In the East Kleinemonde Estuary, the first canonical axis (horizontal) described 85 % and the second canonical axis (vertical) described 94 % of the variation of the species – environment relation (Figure 4.3.2 and Table 4.3.3). The most significant environmental factor was redox potential (0.309). High redox potential was correlated with an increase in live plant cover, plant height and the number of inflorescences and seeds. A high water level was negatively correlated with the number of seeds and plant height. Sediment moisture content, organic matter content, electrical conductivity, salinity, temperature and pH were all positively related with an increase in live plant material. In the Kowie Estuary, the first canonical axis (horizontal) described 99.8 % of the variation of the species – environment relation (Figure 4.3.2 and Table 4.3.4). The most significant environmental factors were electrical conductivity (0.226) and salinity (0.174). High electrical conductivity and salinity were correlated with an increase in the number of seeds.

119

(a)

(b)

Figure 4.3.1: Ordination diagram based on a PCA of species and environmental data for (a) J. kraussii and (b) J. acutus. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Dead cover = dead plant cover; Height = plant height; Inflorescence = inflorescence number; Seed = seed number; Inflorescence 2nd gen. = inflorescence number of second generation; Seed 2nd gen. = seed number of second generation.

120

Table 4.3.1: Summary of PCA of species- and environmental data for J. kraussii in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.588 0.259 0.147 0.006 1.000 Species-environment correlations 0.381 0.466 0.468 0.339 Cumulative percentage variance Of species data 58.8 84.7 99.4 100.0 Of species-environment relation 48.9 81.1 99.6 100.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.175

Table 4.3.2. Summary of PCA of species and environmental data for J. acutus in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.978 0.022 0.000 0.000 1.000 Species-environment correlations 0.504 0.297 0.000 0.000 Cumulative percentage variance of species data : 97.8 100.0 0.0 0.0 of species-environment relation: 99.2 100.0 0.0 0.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.251

121

(a)

(b)

Figure 4.3.2: Ordination diagram based on a PCA of species- and environmental data for S. virginicus in in the (a) East Kleinemonde Estuary; and (b) the Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Live cover = live plant cover; Dead cover = dead plant cover;

Height = plant height; Inflorescence = inflorescence number; Seed = seed number.

122 Table 4.3.3: Summary of PCA of species and environmental data for S. virginicus in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance Eigenvalues 0.960 0.020 0.010 0.006 1.000

Species-environment correlations 0.274 0.596 0.327 0.529

Cumulative percentage variance

Of species data 96.0 97.9 99.0 99.6

Of species-environment relation 85.2 93.5 94.8 96.8

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.085

Table 4.3.4: Summary of PCA of species and environmental data for S. viriginicus in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance Eigenvalues 0.998 0.002 0.000 0.000 1.000

Species-environment correlations 0.371 0.126 0.199 0.143

Cumulative percentage variance

Of species data 99.8 100.0 100.0 100.0

Of species-environment relation 100.0 100.0 100.0 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.137

4.3.3 Sarcocornia decumbens

In the East Kleinemonde Estuary, the both the first (horizontal) and second (vertical) canonical axis described 100 % of the variation of the species – environment relation (Figure 4.3.3 and Table 4.3.5). The most significant environmental factor was redox potential (0.364). High redox potential was correlated with an increase in live plant cover, plant height and the number of inflorescences and seeds. Sediment organic matter content was also positively related with an increase in these biotic features, whilst an increase in water level was negatively related with them. An increase in dead plant cover, including height and cover of second generation plants was correlated with an increase in water level and pH. In the Kowie Estuary, both the first (horizontal) and second (vertical) canonical axis described 100 % of the variation of the species – environment relation (Figure 4.3.3 and Table 4.3.6). The most significant environmental factor was redox potential (0.432). High redox potential was correlated with an increase in the number of seeds. Low pH, electrical conductivity and salinity was correlated with an increase in live plant cover and the number of inflorescences.

123

(a)

(b)

Figure 4.3.3: Ordination diagram based on a PCA of species - and environmental data for S. decumbens in the (a) East Kleinemonde Estuary; and (b) the Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Live 2nd gen = Second generation live plant cover; Dead cover = dead plant cover; Height = plant height; Height 2nd gen. = Second generation plant height; Inflorescence = inflorescence number; Seed = seed number.

124

Table 4.3.5: Summary of PCA of species and environmental data for S. decumbens in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000 Species-environment correlations 0.552 0.241 0.000 0.000 Cumulative percentage variance Of species data 100.0 100.0 0.0 0.0 Of species-environment relation 100.0 100.0 0.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.305

Table 4.3.6: Summary of PCA of species and environmental data for S. decumbens plants in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000 Species-environment correlations 0.522 0.239 0.000 0.000 Cumulative percentage variance Of species data 100.0 100.0 0.0 0.0 Of species-environment relation 100.0 100.0 0.0 0.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.264

4.3.4 Sarcocornia hybrid

In the Kowie Estuary, both the first (horizontal) and second (vertical) canonical axis described 86 % and 92 % of the variation of the species – environment relation respectively (Figure 4.3.4 and Table 4.3.7). The most significant environmental factors on the horizontal axis were rainfall (-0.1528), temperature (-0.112) and redox potential (0.111). The most significant environmental factors on the vertical axis were salinity (-0.173), electrical conductivity (-0.172) and redox potential (0.113). High rainfall was correlated with an increase in second generation plant height and cover and a decrease in first generation plant cover. A high redox potential was correlated with an increase in inflorescences and seeds of both generations. High temperature was related with a decrease in first generation inflorescences, seeds and live plant cover. Plant height of first generation plants increased in response to an increase in salinity.

125

Figure 4.3.4: Ordination diagram based on a PCA of species and environmental data for the Sarcocornia hybrid in the the Kowie Estuary. Abbreviations: pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Live cover 2nd gen = Second generation live plant cover; Dead cover = dead plant cover; Height = plant height; Height 2nd gen. = Second generation plant height; Inflorescence = inflorescence number; Inflorescence 2nd gen = Second generation inflorescence number; Seed 2nd gen = Second generation seed number.

Table 4.3.7: Summary of PCA of species- and environmental data for Sarcocornia hybrid plants in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance Eigenvalues 0.858 0.142 0.000 0.000 1.000

Species-environment correlations 0.328 0.248 0.285 0.281

Cumulative percentage variance

Of species data 85.8 100.0 100.0 100.0

Of species-environment relation 91.4 100.0 100.0 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.101

126

4.3.5 Salicornia meyeriana

In the East Kleinemonde Estuary, the first canonical axis (horizontal) described 66 % and the second canonical axis (vertical) described 55 % of the variation of the species – environment relation (Figure 4.3.5 and Table 4.3.8). The most significant environmental factor influencing the first axis was water level (-0.132). High water level was correlated with a decrease in all the biotic features of the first generation plants. The most significant environmental factors on the second axis were sediment moisture content (-0.214) and organic matter content (-0.203). Low sediment organic matter and moisture content were related to increases in all the biotic features of the second generation plants. In the Kowie Estuary, both the first canonical axis (horizontal) and the second canonical axis (vertical) described 100% of the variation of the species – environment relation (Figure 4.3.5 and Table 4.3.9). The most significant environmental factors were temperature (0.296) and sediment redox potential (0.392). A high temperature was related with increases in live plant cover and decreases in the plant height of first and third generation plants. A high redox potential was related to an increase in inflorescences. An increase in salinity, electrical conductivity and pH was related to an increase in live plant cover.

4.3.6 Sarcocornia tegetaria

In the East Kleinemonde Estuary, the first and second canonical axes described 96 % and 94 % of the variation of the species – environment relation respectively (Figure 4.3.6a and Table 4.3.10). The most significant environmental factor on the horizontal axis was water level (-1.05). An increase in water level was related to a decrease in live plant cover and the number of inflorescences and seeds and an increase in dead plant cover. The most significant environmental factor on the vertical axis was rainfall (0.35). An increase in rainfall was related to an increase all the biotic features of the second generation plants. In the Kowie Estuary, the first canonical axis (horizontal) described 100 % of the variation of the species – environment relation (Figure 4.3.6b and Table 4.3.11). The most significant environmental factor was the water column salinity (0.141). High water column salinity was related with an increase in the number of inflorescences and live plant cover. Rainfall increased dead plant cover and plant height.

127

(a)

(b)

Figure 4.3.5: Ordination diagram based on a PCA of species- and environmental data for (a) S. meyeriana in the East Kleinemonde Estuary; and (b) S. meyeriana in the Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Live cover 2nd = Second generation live plant cover; Dead cover = dead plant cover; Dead cover 2nd = = Second generation dead plant cover; Height = plant height; Height 2nd gen. = Second generation plant height; Height 3rd gen. = Third generation plant height; Inflorescence = inflorescence number; Inflorescence 2nd gen = Second generation inflorescence number; Seed 2nd gen = Second generation seed number.

128

Table 4.3.8: Summary of PCA of species and environmental data for S. meyeriana in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.663 0.337 0.000 0.000 1.000

Species-environment correlations 0.232 0.297 0.591 0.225

Cumulative percentage variance

Of species data 66.3 100.0 100.0 100.0

Of species-environment relation 54.5 99.9 100.0 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.066

Table 4.3.9: Summary of PCA of species and environmental data for S. meyeriana in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000

Species-environment correlations 0.553 0.441 0.000 0.000

Cumulative percentage variance

Of species data 100.0 100.0 0.0 0.0

Of species-environment relation 100.0 100.0 0.0 0.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.305

129

(a)

(b)

Figure 4.3.6: Ordination diagram based on a PCA of species and environmental data for S. tegetaria (a) first generation plants; and (b) second generation plants in the East Kleinemonde Estuary; and (c) in the Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; Water salinity = water column salinity; EC = sediment electrical conductivity; Water = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Dead cover = dead plant cover; Height = plant height; Height 2nd gen. = Second generation plant height; Inflorescence = inflorescence number; Inflorescence 2nd gen = Second generation inflorescence number; Seed 2nd gen = Second generation seed number.

130

Table 4.3.10: Summary of PCA of species and environmental data for first generation S. tegetaria plants in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.961 0.034 0.003 0.002 1.000

Species-environment correlations 0.289 0.395 0.329 0.123

Cumulative percentage variance

Of species data 96.1 99.6 99.8 100.0

Of species-environment relation 93.4 99.6 100.0 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.086

Table 4.3.11: Summary of PCA of species and environmental data for S. tegetaria in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000 Species-environment correlations 0.333 0.217 0.156 0.000 Cumulative percentage variance Of species data 100.0 100.0 100.0 0.0 Of species-environment relation 100.0 100.0 100.0 0.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.111

131

4.3.7 Phragmites australis

In the East Kleinemonde Estuary, the first canonical axis (horizontal) described 98 % of the variation of the species – environment relation (Figure 4.3.7 and Table 4.3.12). The most significant environmental factor was water level (0.592). High water level was related with an increase in plant height and dead plant cover and a decrease in live plant cover. Redox potential and rainfall influenced live plant cover, while temperature influenced plant height. No inflorescences developed in the quadrats. In the Kowie Estuary, the first (horizontal) and second (vertical) canonical axes described 52 % and 98 % of the variation of the species – environment relation (Figure 4.3.7 and Table 4.3.13). The most significant environmental factors were pH (0.211) and temperature (-0.208). High pH and temperature were positively correlated with an increase in plant height and live plant cover. An increase in salinity and electrical conductivity was related to an increase in the number of inflorescences and seeds.

4.3.8 Bolboschoenus maritimus

In the East Kleinemonde Estuary, both the first (horizontal) and second (vertical) canonical axis described 100 % of the variation of the species – environment relation (Figure 4.3.8 and Table 4.3.14). The most significant environmental factor was water level (-0.486). High water level was related with a decrease in all the biotic features. An increase in salinity and electrical conductivity was related with a decrease in all the biotic features. Plant height was influenced by sediment organic matter content. In the Kowie Estuary, the first canonical axis (horizontal) described 100 % of the variation of the species – environment relation (Figure 4.3.8 and Table 4.3.15). The most significant environmental factor was temperature (0.294). A high temperature was positively correlated with live plant cover, height and the number of inflorescences and negatively correlated with dead plant cover.

4.3.9 Ruppia cirrhosa and Chara vulgaris

In the East Kleinemonde Estuary, the first canonical axis (horizontal) described 90 % of the variation of the species – environment relation for R. cirrhosa (Figure 4.3.9 and Table 4.3.16). The most significant environmental factors were pH (0.404) and water level (0.378). High pH and water level were related with an increase in plant height and biomass. Water temperature and moisture content also positively influenced biomass, the number of flowers and flowering inflorescences. The first canonical axis (horizontal) described 99 % of the variation of the species – environment relation for C. vulgaris (Figure 4.3.9 and Table 4.3.17). The most significant environmental factors were water level (0.244) and water temperature (0.236). High water level was related with an increase in antheridia, oogonia and plant height. An increase in water temperature was related with an increase in biomass, plant height and oogonia.

132

(a)

(b)

Figure 4.3.7: Ordination diagram based on a PCA of species and environmental data for P. australis in the (a) East Kleinemonde Estuary and (b) Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Dead cover = dead plant cover; Height = plant height; Inflorescence = inflorescence number; Seed = seed number.

133

Table 4.3.12: Summary of PCA of species and environmental data for P. australis in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.898 0.086 0.016 0.000 1.000

Species-environment correlations 0.665 0.294 0.343 0.000

Cumulative percentage variance

Of species data 89.8 98.4 100.0 0.0

Of species-environment relation 97.7 99.5 100.0 0.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.407

Table 4.3.13: Summary of PCA of species and environmental data for P. australis in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.585 0.385 0.030 0.001 1.000 Species-environment correlations 0.307 0.353 0.276 0.547 Cumulative percentage variance Of species data 58.5 97.0 99.9 100.0 Of species-environment relation 52.2 97.7 99.8 100.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.106

134

(a)

(b)

Figure 4.3.8: Ordination diagram based on a PCA of species and environmental data for B. maritimus in the (a) East Kleinemonde Estuary and (b) Kowie Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = sediment pH; Redox = sediment redox potential; Salinity = sediment salinity; Water salinity = water column salinity; EC = sediment electrical conductivity; Water Level = quadrat water level; Temp = air temperature; Rain = Rainfall; Live cover = live plant cover; Dead cover = dead plant cover; Height = plant height; Inflorescence = inflorescence number; Seed = seed number.

135

Table 4.3.14: Summary of PCA of species and environmental data for B. maritimus in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000

Species-environment correlations 0.428 0.827 0.419 0.338

Cumulative percentage variance

Of species data 100.0 100.0 100.0 100.0

Of species-environment relation 100.0 100.0 100.0 0.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.183

Table 4.3.15: Summary of PCA of species and environmental data for B. maritimus in the Kowie Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 1.000 0.000 0.000 0.000 1.000

Species-environment correlations 0.506 0.558 0.343 0.493

Cumulative percentage variance

of species data 100.0 100.0 100.0 100.0

of species-environment relation 100.0 100.0 100.0 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.256

136

(a)

(b)

Figure 4.3.9: Ordination diagram based on a PCA of species and environmental data for (a) R. cirrhosa and (b) C. vulgaris in the East Kleinemonde Estuary. Abbreviations: OMC = sediment organic matter content; MC = sediment moisture content; pH = water column pH; Redox = water column redox potential; Salinity = water column salinity; EC = water column electrical conductivity; Water level = quadrat water level; Water Temp = water temperature; Temp = air temperature; Biomass = plant biomass; Height = plant height; Flower inflorescences = flowering inflorescence number; Fruit inflorescences = fruiting inflorescence number; FS = fruit or seed number; A = antheridia number; O = oogonia number.

137

Table 4.3.16: Summary of PCA of species and environmental data for R. cirrhosa in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.863 0.075 0.059 0.002 1.000

Species-environment correlations 0.826 0.744 0.665 0.474

Cumulative percentage variance

Of species data 86.3 93.8 99.7 99.9

Of species-environment relation 89.6 96.0 99.9 100.0

Sum of all eigenvalues 1.000

Sum of all canonical eigenvalues 0.656

Table 4.3.17: Summary of PCA of species and environmental data for C. vulgaris in the East Kleinemonde Estuary.

Axis 1 Axis 2 Axis 3 Axis 4 Total variance

Eigenvalues 0.976 0.024 0.000 0.000 1.000 Species-environment correlations 0.487 0.322 0.000 0.000 Cumulative percentage variance Of species data 97.6 100.0 0.0 0.0 Of species-environment relation 98.9 100.0 0.0 0.0

Sum of all eigenvalues 1.000 Sum of all canonical eigenvalues 0.234

138 5. CHAPTER 5: DISCUSSION

This study reports on the effect of environmental factors on macrophyte phenology in a temporarily open/closed estuary (TOCE) compared to a permanently open estuary (POE). Due to the unpredictable nature of TOCEs, the aim of the study was to determine whether macrophyte phenology in a TOCE was event driven rather than cyclical. A comparison with a POE established the differences between a more stable and predictable environment, as reflected in the POE, versus the more extreme and unpredictable TOCE environment. An understanding of the conditions for macrophytes to complete their life-cycles will assist managers with water allocations and mouth management plans in TOCEs to ensure the persistence of macrophyte habitats and ecological processes in these systems. The specific hypotheses of the research were as follows:

 The life-cycles of macrophytes in TOCEs are event driven i.e. they are driven by the estuary mouth being open or closed due to the associated water level fluctuations and environmental conditions.

 Environmental conditions in a TOCE are more variable than environmental conditions in a POE, which are more stable.

 Macrophytes respond to relatively small water level fluctuations (10 - 20 cm) without the mouth breaching in a TOCE.

 Macrophytes complete their life-cycles more rapidly in a TOCE compared to a POE.

 Macrophytes are highly plastic i.e. adaptable and flexible in response to fluctuating and stressful environmental conditions in a TOCE.

 Intertidal salt marsh requires at least two months for plants to produce viable seeds after flowering and four months after germination in a TOCE.

 Submerged macrophytes require stable water levels for at least three to four months for viable seeds to develop after germination in a TOCE.

The limitation of this phenological study is that it represents a once off measurement of macrophyte characteristics under closed mouth conditions, which are anticipated to be dissimilar or to change under different environmental conditions. However, the sampling period was conducted over a 17 month period in an attempt to represent two growing seasons and reproduction periods.

5.1 ABIOTIC CONDITIONS IN A TEMPORARILY OPEN/CLOSED ESTUARY COMPARED WITH A PERMANENTLY OPEN ESTUARY The East Kleinemonde Estuary remained closed during the entire sampling period (17 months) due to low average rainfall. The estuary opens on average 2.6 times per year, but has been known to close for up to two years (van Niekerk et al., 2008; Riddin and Adams, 2008a). Several overwash events occurred with a major event on 24 June 2009. The last mouth opening event was the 2 / 3 September 2008. During mouth closure, overwash events in the TOCE occur for 26 % of the time in one year (Whitfield et al., 2008). Mouth closure usually occurs at water levels varying between 0.5 and 1 m amsl (Riddin and Adams, 2008a). During this study

139 the average water level was high and ranged from 1.5 - 2.4 m amsl. Subsequent to the overwash event on 24 June 2009, the water depth increased from 1.9 - 2.3 m amsl, which is typical during mouth closure (van Niekerk et al., 2008). Mouth closure caused the intertidal zone to be inundated for 12 months (July 2009 - June 2010). The supratidal zone, on the other hand, was inundated for six to seven months (July 2009 - January 2010). Water depth dropped from 2.3 to 2.1 m amsl in January 2010 (20 cm) and from 2.05 to 1.94 m amsl in February 2010 (11 cm) to expose bare areas that could be colonized due to the die-back of submersed plants. As a result, the estuary was perched with no tidal variation and no intertidal areas for 17 months during the sampling period.

Historically the estuary has a mean salinity of 23 – 25 ppt during a closed period (Riddin and Adams, 2010). During this study salinity was high ranging from 30 - 42 ppt, excluding March 2009 which was 23 ppt. The overwash events and relatively low rainfall over the sampling period contributed to the high salinity (Figure 4.1.1). The study area has experienced drought conditions since early 2009 and the average rainfall in 2009 was the lowest since the 1960‘s (Gumenge, 2010; Petzer, 2010). This is the first record of an extended saline period in the 15 years of monitoring the estuary (Riddin and Adams, 2010). The pH within TOCEs is expected to range between 7 and 8.5 (Snow and Taljaard, 2007) while in the East Kleinemonde Estuary pH tends to range between 7.7 and 8.3 during all conditions. In this study, pH was within the former range i.e. 7.2 – 8.5. Water temperatures showed a distinct seasonal pattern with summer and spring temperatures (20 – 28.7°C) higher than winter and autumn temperatures (14 – 22°C), which is typical of these estuaries (Snow and Taljaard, 2007). The water column redox potential was mostly well aerated although there were several months showing reduced conditions. This variation coincides with previous studies in the estuary during the closed state (Whitfield et al., 2008), where little river inflow reduces the oxygen content and allows for hypoxic or anoxic conditions to develop in the bottom waters (Snow and Taljaard, 2007). Shallow water depths and wind- mixing probably prevented the estuary from becoming hypoxic for an extended period despite the closed mouth condition.

In the Kowie Estuary, flow is considered erratic due to frequent droughts and floods in the catchment (Cowley et al., 2003). Water temperatures have shown seasonal patterns in the past with mean summer temperatures ranging from 19 - 24°C or 20 - 28 °C, and winter temperatures ranging from 13 - 16°C or 11 - 16 °C (Hill and Allanson, 1971; ; Heinecken & Grindley, 1982). During this study, salinity was for the most part high, ranging from 30 - 34.2 ppt, with August 2009 to December 2009 ranging from 21 - 25 ppt. Salinity has generally been recorded above 30 ppt but in dry years it has increased to 40 ppt (Day, 1981; Harrison, 2003), while surface waters have been recorded as almost fresh during flood periods (Whitfield et al., 1994). The pH and turbidity are usually high (mean 8.2), while high tide levels are similar to the adjacent coastline (Whitfield and Bate, 2007).

Sediment salinity and electrical conductivity was generally higher in the TOCE compared to the POE. This was due to the highly saline condition of the water column that inundated all the habitats in the TOCE. This was caused by an extended closed mouth and low freshwater input. In contrast, only the B. maritimus and S. tegetaria habitats in the POE were wholly and partially inundated by the daily tidal exchange respectively. The average salinity in these habitats was very similar in the TOCE compared to the POE, namely 17.5 ± 1.7 ppt

140 versus 16.8 ± 1.3 ppt for B. maritimus respectively and 22.4 ± 2.0 ppt versus 22.1 ± 1.9 ppt for S. tegetaria respectively. All other macrophyte habitats in the Kowie Estuary did not experience tidal inundation and therefore sediment salinity remained lower. In both estuaries sediment pH ranged for the most part from 6.2 - 8.4, but was mostly alkaline. The TOCE did however experience very acidic conditions in the supratidal zone during several months. Sediment redox potential in the TOCE was largely negative due to submerged conditions and poor aeration, wheras in the POE it was mostly positive or well aerated. Only the B. maritimus habitat in the POE was always waterlogged and anoxic. The sediments of salt marshes, particularly in the lower intertidal areas, are frequently anoxic because of high moisture content, organic matter content and microbial activity. This creates sulfides that blacken the sediment and emit a rotten-egg odour (Zedler et al., 2008). Sediment salinity, electrical conductivity, pH and redox potential were significantly different between the two estuaries in the supratidal and intertidal habitats, while sediment pH and redox potential were significantly different in the reed and sedge habitats.

Conditions in the TOCE were therefore not as favourable as those in the POE for macrophyte phenology, and tended to be more extreme due to the estuary mouth remaining closed for an extended period. Water level and salinity were extremely high for most of the study period, causing inundation of the suptratidal and intertidal habitat. The physical environment of these small TOCEs is known to experience large fluctuations, such as during this study, and consequently the macrophytes are mostly influenced by physical rather than biological factors (Riddin and Adams, 2008a). In contrast, daily tidal exchange, including seasonal and cyclical changes, such as reproduction, growth and die-back, was evident in the POE. No major events such as flooding occurred during the sampling period and abiotic conditions remained within the average conditions historically experienced in the estuary, although the region did experience drought.

POEs generally experience more stable environmental conditions relative to TOCEs due to their permanent tidal nature and connection to the sea, coupled with larger freshwater inflows. This enables macrophytes to complete their life-cycles and set seeds unconstrained. The stability of the POE environment is reflected in the well established zonation pattern of the salt marsh. This zonation is particularly well established in estuaries with high tidal ranges, such as POEs (Adams, 1991; Davy, 2000; Rogel et al., 2000; Rogel et al., 2001; Bockelmann et al., 2002; Costa et al., 2003; Ursino et al., 2004). The succession and spatial zonation of salt marsh plants is a balance between tolerance and competition (Gray, 1985), where plant distribution is strongly dictated by inter-specific competition (Snow and Vince, 1984; Bertness and Ellison, 1987; Pennings and Callaway, 1992; Streevers and Genders, 1997, Pennings et al., 2005). Species that produce salt-tolerant ramets and propagate vegetatively are able to colonize bare areas in the lower marsh (Bertness and Ellison, 1987; Bertness, 1991; Shumway and Bertness, 1992; Brewer et al., 1997, Onaindia et al., 2001). Limited space is available for plants to expand due to limited disturbance events in this environment (Streever and Genders, 1997). Small disturbance events, such as wrack stranded by the tide that smothers the underlying vegetation, can create bare space for colonisation in the intertidal or lower zones (Bertness et al., 1992; Minchinton, 2002). In contrast, water level fluctuates in TOCEs depending on mouth condition and storm or freshwater events (van Niekerk et al., 2008). This creates major disturbances which result in significant spatial changes, such as the increase in bare soil for colonisation through sexual or asexual reproduction. In stressful environments, such as TOCEs, abiotic factors are therefore the determinants of species distribution (Castillo et al., 2000, Riddin and

141 Adams, 2008a) while in less stressful and more permanent environments, such as POEs, biotic factors e.g. tolerance and competition, are the determinants (Ungar, 1978; Snow and Vince, 1984; Bertness and Ellison, 1987; Pennings and Callaway, 1992; Streevers and Genders, 1997; Pennings et al., 2005).

5.2 MACROPHYTE PHENOLOGY IN A TEMPORARILY OPEN/CLOSED ESTUARY COMPARED WITH A PERMANENTLY OPEN ESTUARY

5.2.1 Juncus kraussii and Juncus acutus Plant cover of J. kraussii declined due to the supratidal habitat being inundated for 6 - 7 months in the TOCE, showing signs of decline after two months of inundation (Figure 5.1). This coincides with a study in the estuary in which supratidal salt marsh cover was affected after one to two months of inundation and J. kraussii cover occurred declined significantly (Riddin and Adams, 2008a). The sediment pH was high (7 ± 0.2) for most of the sampling period, which also inhibited plant growth as J. kraussii prefers a pH range of 5.4 - 6.6 (Clarke and Jacoby, 1994). Although J. kraussii is very tolerant of high salinity and flooded conditions, biomass has been shown to decrease with increasing salinity (10 – 40 ppt) under flooded conditions (Zedler et al., 1990; Naidoo and Kift, 2006; Riddin and Adams, 2010).). In this study, new growth was limited to late summer when the water had receded, providing high moisture content. However, the combined effect of high pH, water level and salinity reduced plant cover. In the POE, the growth of J. acutus was seasonal because cover responded to temperature. During the cover expansion periods, the mean monthly rate of cover expansion for J. kraussii was 13 % compared to 4 %, which was previously recorded in the estuary over a one year period (Riddin and Adams, 2008a). This study therefore demonstrates the plasticity of the species once favourable conditions resume. At commencement of the study period, mean live plant cover in both estuaries was similar, but by end of the sampling cover was significantly higher in the POE compared to the TOCE, which was due to the more extreme conditions in the TOCE.

Flowering of J. kraussii usually occurs in spring and summer (Muir, 2000). Consequently, plants in the TOCE followed the seasonal pattern by commencing in November 2009, despite four months of high water level (19 - 28 cm) prior to flowering. However, the duration of flowering was shortened by two months and reproductive output was low i.e. 77 seeds m-² due to six months of submergence and high salinity. Research has shown that inflorescence production is affected by season and water depth in J. militaris, with high water levels delaying and reducing inflorescence production (Grace and Wetzel, 1982; Hogeland and Killingbeck, 1985). The negative effect of salinity on the production of J. gerardi inflorescences was found to be amplified when combined with flooding; and fruit maturation was significantly reduced under high water level (Charpentier et al., 2009). Flowering of J. acutus, on the other hand proliferated in the POE, following the seasonal flowering pattern of spring and summer flowering (Department of Nature and Conservation, 2006). Seed output was high with a maximum mean of 318 799 ± 76 903 seeds m-2 during peak seeding. Diggory and Parker (2010) estimated 3 467 seeds m-2 for Juncus arcticus ssp. Littoralis, which is much lower probably due to the smaller inflorescence of the species. Ervin and Wetzel (2001) estimated a maximum seed fall of 2 481 200 seeds m-2 for Juncus effusis. Inflorescences and seeds were significantly higher in the POE compared to the TOCE, which was due to the POE experiencing no flooding. Plants from both estuaries produced viable seeds within two months of flowering commencing.

142

Figure 5.1: Plant cover of J. kraussii in the East Kleinemonde Estuary in April 2009 (left) compared to January 2010 (right), when significant die-back had occurred due to prolonged inundation. Once the water receded in January 2010, re-growth was rapid probably due to the significant die-back that occurred during inundation and the warmer summer temperatures.

5.2.2 Sporobolus virginicus In the TOCE, the S. virginicus habitat was flooded during the entire sampling period i.e. 12 - 75 cm. Both high water level and salinity caused a significant decline in plant cover. Although S. virginicus is tolerant of frequent inundation, long periods of submersion and highly saline conditions (Breen et al., 1977; Marcum and Murdoch, 1992; Adams et al., 1999; Bell and O'Leary, 2003; Walker, 2003), as found in this study, these conditions have been reported to reduce S. virgnicus biomass (Naidoo and Naidoo, 1992; Naidoo and Mundree, 1993; Naidoo and Naidoo, 2000). For example, S. virginicus has been found growing in salinity levels of 28 – 34 ppt (Breen et al., 1977; Marcum and Murdoch, 1992; Naidoo and Naidoo, 1998; 1999; Muir, 2000), while Naidoo and Mundree (1993) found that S. virginicus tolerated 42 days of flooded conditions. In this study, plants survived a period of 7 - 8 months of submergence and vegetative re-growth resumed when the water level declined to approximately 9 cm or when the habitat was exposed. Riddin and Adams (2008a; 2010) recorded re-growth of plant material which was submerged for three months and that supratidal salt marsh decreased significantly with high salinity and water level in the East Kleinemonde Estuary. In the POE, plant cover demonstrated no cyclical or seasonal pattern, which is typical of this species as it grows throughout the year. This finding is also supported by Clarke and Jacoby (1994). In general, cover remained comparatively stable when compared to the TOCE. The mean monthly rate of cover expansion was considerably higher in the TOCE compared to the POE (i.e. 36 % compared to 10.5 %), because S. virginicus re-established rapidly once exposed.

In the TOCE, flowering followed the seasonal cycle with peak flowering in autumn and summer (Eleuterius and Caldwell, 1984). Reproductive output was substantially shortened and was significantly reduced in late autumn due to prolonged high water level. Kercher and Zedler (2004) also found that most of the 17 angiosperm wetland plants tested for flood tolerance had a lower percentage of flowering individuals due to flooding. Warwick and Brock (2003) showed that the reproduction of plants that tolerate fluctuating water depths was

143 slowed or prevented. Many researchers have suggested that flooding can decrease wetland plant production (Blanch et al., 1999; Chen et al., 2002). Although waterlogged conditions can accelerate flowering (Laegdsgaard, 2006), high water level and almost complete die-back in this study prevented reproduction in summer, during the second reproduction period. In contrast, plants in the POE produced inflorescences from spring to autumn over a seven month period with a mean maximum of 101 inflorescences m-², three times more than the TOCE (35 inflorescences m-²). Seeds were also produced in significantly higher quantities than in the TOCE i.e. mean maximum of 2 428 m-² versus 715 seeds m-². No seeds were produced in the second reproductive period after flooding in the TOCE. Seeds took approximately two months to develop after flowering commenced in the POE. A time line could not be accurately established in the TOCE due to flooding, but seeds probably took one to two months to develop. No significant monthly changes in seed viability were evident for both estuaries. The type of seed bank (transient versus persistent) could not be determined from the literature. However, research by Sartor and Marone (2010), including Khan (1993), suggest that Sporobolus spp. represent transient seed banks. Riddin and Adams (2009) found 163 seeds m-2 in the seed bank of the East Kleinemonde Estuary, but seeds could have been shed within a year of collection, which would suggest a transient seed bank.

5.2.3 Sarcocornia and Salicornia species In the TOCE, S. decumbens cover declined significantly after one month of inundation in July 2009 when the mean quadrat water depth was 36 cm. The S. tegetaria habitat was inundated by 12 - 61 cm and S. meyeriana habitat by 16 - 54 cm of water from July 2009 to June 2010. Inundation caused a significant decline in mean live plant cover of both these species, coupled with reduced sediments. Biomass reduction in response to low Eh is a common response in wetland plants (Kludze and de Laune, 1994; Pezeshki et al., 1996). These observations concur with a recent study in the estuary when intertidal salt marsh cover declined due to substantial water level rise (Riddin and Adams, 2010). Sarcocornia natalensis (Steud.) Dur and Schinz has been observed to die-back after two to three months of submergence (Tölken, 1967; CSIR, 1992; Adams et al., 1999). Davy (2006) reports that plants are intolerant of continuous complete submergence although rapid stem elongation does occur. Adams and Bate (1994a) report on extension growth in S. tegetaria in response to submergence, as recorded in this study after one month of submergence. S. decumbens survived complete inundation for six months, while S. tegetaria survived after seven months of submergence. In contrast, studies in the East Kleinemonde Estuary have shown that intertidal salt marsh dies back after three months of submergence (Riddin and Adams, 2008b).

Seeds of all three species survived an 8 - 11 month inundation period showing that seeds of many halophytes can survive long periods of submergence in hypersaline conditions and germinate later once the salinity stress is lowered (Ungar, 1978; 1987; Woodell, 1985). Although the germination percentage in S. meyeriana declined significantly after one and two months in the germination trials, germination in situ was due to the production of viable seeds which lay dormant during inundation. Although S. decumbens seedlings were only identified in mid autumn, after 1.5 months of the water receding from the supratidal, it is probable that seeds germinated earlier i.e. within 6 days (based on the germination trials) to one month of the water retreating in late summer/early autumn. This is supported by Adams and Riddin (2008a) that reported seedling emergence in late winter in the East Kleinemonde Estuary in response to drops in water level and sediment exposure, while Bornman (2002)

144 recorded S. pillansii emerging in winter when rainfall had reduced soil salinity. Seeds of S. tegetaria and S. meyeriana germinated within days of the water receding during mid-summer and also expanded into winter showing aseasonal growth. Sarcocornia seedlings have emerged within three days after the mouth has breached in the East Kleinemonde Estuary (Riddin and Adams, 2008a), while the germination trials of this study showed that S. tegetaria and S. meyeriana seeds germinated within an average of 4.8 - 5 days respectively. S. meyeriana seeds germinated despite the high sediment salinity in mid-summer (28 ppt). It is one of the most salt tolerant species (Chapman, 1974; Davy, 2001) and several Salicornia species have germinated under relatively high salinity (Ungar, 1962; Rivers and Weber, 1971; Chapman, 1974; Ungar, 1978). All species therefore demonstrated aseasonal growth responses due to water level reduction i.e. peak growth outside of the usual spring peak growing season (Adams and Bate, 1994b). Seeds of all three species continued to germinate aseasonally, from summer through to winter i.e. intermittently for five to six months after the water retreated. Similarly, Riddin and Adams (2008b) reported intermittent germination for three months after the estuary breaches. S. meyeriana was unable to complete its annual life-cycle during 2009 because germination was prevented after seed formation in April 2009.

In the POE, S. decumbens, the Sarcocornia hybrid and S. tegetaria plant cover showed seasonal growth patterns. However, S. decumbens also responded positively to higher rainfall, the Sarcocornia hybrid negatively to low Eh and S. tegetaria positively to high water column salinity. Large increases in precipitation may result in significant changes in vegetation structure (Dunton et al., 2001). The seasonal responses were associated with autumn and summer growth peaks, while S. decumbens peaks also occurred in winter. However, plant cover of the Sarcocornia hybrid declined in summer due to low Eh, which was possibly caused by the combined effect of very compact clay sediments moistened by rainfall, including high pH and salinity. Marked seasonal patterns in biomass accumulation which is a characteristic of temperate zone coastal marshes has been observed in other studies (Curcó et al., 2002; Scarton et al., 2002; Palamo, 2009) and according to Onaindia et al. (2001) summer is the period for the optimum vegetative development of halophytic species. Zedler et al. (2000) confer that plants that dominate salt marshes in Mediterranean climates grow best in the warm season (Davy, 2001), although in South Africa peak S. tegetaria growth has been observed in spring (Adams and Bate, 1994a). The annual species, S. meyeriana, germinated in autumn and cover increased through to summer, showing no seasonal pattern but rather an annual cyclical pattern as a third discrete generation emerged the following autumn. The germination periods of Salicornia normally coincide with low sediment salinity (Ungar et al., 1979; Smith, 1985 cited in Davy, 2001; Onuf, 2006), as reflected in this study. Dieguez and Breceda (1992) found that S. bigelovii populations reach their highest stand in spring, while the growing season of S. europea was during summer and autumn in a study by Ungar (1987b). Similarly, in this study the highest stand was reached in summer after nine months of continuous growth over all the seasons.

When compared to the TOCE, the mean monthly plant cover of all three species in the POE remained relatively stable due to the dense matrix of salt marsh cover, limited amount of bare areas and resultant competition (Snow and Vince, 1984; Bertness and Ellison, 1987; Pennings and Callaway, 1992; Streevers and Genders, 1997). Plant cover in the TOCE was significantly lower compared to the POE at the end of the sampling period because of inundtaion. The mean monthly cover expansion of seedlings in all three species was substantially higher in the TOCE. This was due to the availability of bare areas once the water receded and reduced

145 competition which was in contrast to the POE (Snow and Vince, 1984; Bertness and Ellison, 1987; Pennings and Callaway, 1992; Streevers and Genders, 1997).

No S. decumbens or S. tegetaria seedlings were observed germinating in the POE although viable seeds were produced, which may suggest predominance for vegetative reproduction over sexual reproduction. This is typical of established salt marshes where reproduction usually occurs vegetatively (Adams et al., 1999). Kautsky (1988) explains how aquatic macrophytes in stressful environments are able to produce copious seeds and complete their life-cycles rapidly compared to stable environments where plants behave oppositely. Further, seeds under dense vegetation cover in the POE would have reduced germination (Gul and Weber, 1999; Redondo et al., 2004). Davy (2006) writes that S. tegetaria plants can form dense monospecific stands that resist invasion by other species in favourable habitats. High salinity would have reduced seed germination (Ungar, 1978; Redondo et al., 2004; Davy, 2006) and removal of S. tegetaria seeds by high tide may have influenced this as well (Huiskes et al., 1995). The combination of sexual and asexual reproductive strategies also assists with species persistence in unstable habitats, like TOCEs (Desclaux and Roumet, 1996). The Sarcocornia hybrid seeds did however germinate in the POE during winter after higher rainfall. Expansion was considerably slower in the POE as it completed its life-cycle in nine months compared to four months in the TOCE. S. meyeriana also completed its life-cycle very quickly in the TOCE compared to the POE i.e. three months compared to 10 months. This may be a function of the stability of the POE compared to the unpredictability of the TOCE environment, particularly with reference to S. meyeriana as annual plants usually complete their life-cycles rapidly. Riddin and Adams (2008b) observed that after the water level drops and the marsh area is exposed, salt marsh macrophytes re-establish from their seed bank within two months in the East Kleinemonde Estuary. A study by Alexander and Dutton (2002) measured dramatic cover increases in S. bigelovii reducing bare area from 86 to 46 % in eight months in a POE. In comparison to this study, the rate of cover increase is more similar to the POE than the TOCE. Plant decline in the TOCE was more than double than in the POE, where dead bushes remained in situ for 11 months and held seeds for six months in the inflorescences (Figure 5.2).

In the TOCE, flowering of S. decumbens during the second reproduction period did not occur due to inundation, high salinity and low plant cover (6 ± 6%). A significant correlation between low biomass and low reproductive output has been reported for annual terrestrial plants and wetland plants by various researchers (Samson and Werk, 1986; Lovett-Doust, 1989; Titus and Hoover, 1991; Crossle and Brock, 2002; Warwick and Brock, 2003). Reproductive output was also substantially reduced during the second reproductive period of the Sarcocornia hybrid in the POE due to the significant reduction in biomass as a result of low Eh. The second flowering period of S. decumbens in the POE covered a six month period, although seed production during both its reproductive cycles was significantly lower compared to the first reproductive period of the TOCE, namely 102 847 seeds m-2 (TOCE) compared to 20 661 to 31 3598 seeds m-2 (POE). During two reproduction periods, Shaw (2007) estimated 292 984 and 134 277 seeds m-2 produced by Sarcocornia pillansii. An increase in seed production in 10 cm of water occurred, which the author suggested may be a mechanism to enhance the probability of new plants regenerating in waterlogged conditions when stress reduces germination (Mony et al., 2010). Similarly, the fluctuating water level, saturated sediment and the correlation of inflorescences and seeds with water level and redox potential suggest that the stressed conditions may have increased seed output in the TOCE.

146

Figure 5.2: In the Kowie Estuary dead bushes of first generation plants (arrow) remained in situ for 11 months and held seeds for six months in the inflorescences, while second generation plants expanded around and under them.

The second reproduction period in S. tegetaria and S. meyeriana plants was delayed and shortened in the TOCE. A study by Fernández-Illescas et al. (2010) found that S. tegetaria produced a mean of 1 826 inflorescences and 80 016 flowers m-². In contrast, the number of inflorescences and flowers/seeds in the POE and TOCE was considerably lower i.e. maximum mean inflorescences 69 - 400; and flowers/seeds 16 958 - 45 542 m-2 respectively. Seed output was significantly higher in the TOCE compared to the POE despite the fact that plant cover in the TOCE was significantly lower (Samson and Werk, 1986; Lovett-Doust, 1989; Titus and Hoover, 1991), while the reproduction period was one month longer in the TOCE during the second reproduction period. With reference to reproduction in the annual S. meyeriana, Troyo-Dieguez et al. (1994) found that S. bigelovii Tor flowered in summer and inflorescences took seven months to develop. In the TOCE, inflorescences took three months to develop after germination compared to nine months in the POE. Onuf (2006) recorded S. biglevoii flowering in spring and summer; and by autumn the plants had died, four months after flowering. In the POE, flowering occurred in late summer during both reproduction periods. In the TOCE, flowering occurred in autumn during the first reproduction period, but was delayed by one month and extended into winter during the second flowering period. This was due to flooding. According to Beeftink (1985) the annual S. procumbens produced 110 000 seeds m-² while a study by Rubio-Casal et al. (2001) found that the annual S. ramosissima produced 40 450 – 116 120 seeds m-². In this study, maximum seed production occurred in autumn in both estuaries but flowering and seed production in the POE was significantly higher than in the TOCE, with 264 224 – 640 292 seeds m-² compared with 24 050 - 27 643 seeds m-² respectively. This can be attributed to the relative biomass in the estuaries, where plant cover in the POE was higher compared to the TOCE (Samson and Werk, 1986; Lovett-Doust, 1989; Titus and Hoover, 1991). In addition, plants in the POE were more densely branched, producing 275 - 384 inflorescences 100 cm-² compared to 54 - 82 inflorescences 100 cm-² in the TOCE.

147 Based on the duration of the reproduction periods in both the estuaries and seed viability, S. meyeriana required at least one month for flowers to set viable seeds, while S. tegetaria required two months. S. decumbens required two to three months in the POE whereas a time line could not be established in the TOCE. Reproductive dynamics in wetlands is strongly affected by depth, duration and season of inundation (Warwick and Brock, 2003), as reflected in this study. S. meyeriana seeds showed a significant decline in seed viability (germination success) one and two months after maturation in the germination trials. However, annual salt marsh species have persistent seed banks, which suggest that viable seeds would form part of the seed bank in the East Kleinemonde Estuary. S. tegetaria seeds showed a significant decline in seed viability during 2009 only, which was probably due to seed age, unviable seeds and/or fungi development. Riddin and Adams (2009) found 306 and 7 929 seeds m-2 of S. meyeriana and S. tegetaria respectively in the seed bank of the East Kleinemonde Estuary, which would ensure germination once water level declines, as demonstrated in this study.

5.2.4 Phragmites australis In the TOCE, the P. australis habitat experienced water level that ranged predominantly from 40 - 106 cm. Although this species is well adapted to waterlogged conditions (Packham and Willis, 1997), it does not favour a high salinity, as reflected in this study. High water level coupled with high salinity and reduced soils, caused a reduction in live plant cover. Other studies have recorded peak growth in spring, late summer or early autumn (Minchinton, 2002; Scarton et al., 2002), which was observed in this research. Numerous researchers have shown a negative relationship between plant cover and increasing water level and salinity or stable high water level (Yamasaki and Tange, 1981; Hellings and Gallagher, 1992; Adams and Bate, 1994a; Benfield, 1994; Lissner and Schierup, 1997; Adams and Bate, 1999; Clevering, 1999; Riddin and Adams, 2010), which was also reflected in this study. Die-back occured when P. australis plants were inundated for 94 days and salinity was 30 ppt, while growth was inhibited at 20 ppt over a two week period (Adams and Bate, 1999; Benfield, 1994; Lissner and Schierup, 1997). In this study, mean salinity was 22 ppt and mean water level 62 cm. Deegan et al. (2007) state that P. australis prefers moderately fluctuating water levels (± 30 cm amplitude), while Riddin and Adams (2010) indicated that the reduction in reed cover in the East Kleinemonde Estuary was correlated with high water level and salinity. Die-back has also been observed due to reduced substrates (Yamasaki and Tange, 1981; Clevering, 1999) as found in this study (mean Eh of -86 mV). In the POE, a distinctly seasonal growth pattern was evident and the PCA plot identified temperature as one of the main drivers. Sediment pH was also identified as one of main drivers influencing plant cover and height, although it remained fairly constant ranging between 7 and 8; while other studies have shown that this species can tolerate a wide pH range i.e. from very acidic (2.9) to highly alkaline (9.2) (Gucker, 2008). Plant cover was lower in the TOCE compared to the POE due to the prolonged high water level, high salinity and reduced sediment. Mean monthly cover expansion and height increase were more rapid in the TOCE compared to the POE because competition was lower in the TOCE.

P. australis is known to flower in late summer (Auld and Medd, 1987; Stanton, 2005; Greenwood, 2008; Engloner, 2009) and autumn (Ishii and Kadono, 2002; McKee and Richards, 1996). In this study, flowering occurred in early autumn (mean temperature 22°C) and seed output was low in both estuaries. Various studies have reported that seed set rates are generally low and variable (Gustafsson and Simak, 1963; Björk, 1967;

148 Gervais et al., 1993; McKee and Richards, 1996; Ishii and Kadono, 2002). Ishii and Kadono (2002) suggest that self incompatibility may cause low seed productivity, which supports the conclusion from Baldwin et al. (2010). In this study, the potential average number of seeds per inflorescence was 3 187, but the actual seed set was 120 seeds per inflorescence. Other research has indicated 500 - 2 000 (Maheu-Giroux and de Blois, 2007), 350 - 800 seeds (Wijte and Gallagher, 1996) and 0 - 1 000 fertile seeds (McKee and Richards, 1996) per inflorescence. Although seed banks are not necessarily representative of seed production, Phragmites seed banks are considered transient in Europe (Thompson et al., 1997 cited in Baldwin et al., 2010). If the seed banks are transient, the study by Baldwin et al. (2010), which found 284 - 698 seeds m−2 at two high-viability sites compared to 10 seeds m−2 at two low-viability sites, may approximate annual seed production. In this study, the latter coincides with seed production in both the TOCE and POE. During the second reproductive cycle in the TOCE, reproductive output was substantially reduced due to high and saline water; and only five inflorescences or 520 viable seeds developed in the estuary. In comparison, plants in the POE produced 1 080 inflorescences and 10 800 viable seeds per 900 m-2 or one inflorescence and 12 viable seeds m-². Flower and seed production in the POE occurred when salinity increased and rainfall was low, possibly as a response to stress. McKee and Richards (1996) suggest that seed output is positively influenced by warmer temperatures, and that precipitation was positive only during autumn (McKee and Richards, 1996; Engloner, 2009), as supported by studies in Sweden (Gustafsson and Simak, 1963; Björk, 1967). Consequently, flowering was more than likely in response to a combination of warmer temperatures, high salinity and lower rainfall in the POE because it flowered in the typical season of this species. Limited reproduction in the TOCE was due to high water level and salinity due to the closed mouth condition as opposed to the POE where reproduction was substantially higher due to more favouable conditions.

5.2.5 Bolboschoenus maritimus In the TOCE, B. maritiumus was inundated by water levels predominately ranging from 28 – 83 cm. Although adapted to waterlogged conditions, including saline environments (Wilman, 2006), its natural habitat is intertidal (Cowardin et al., 1979; Kantrud, 1996). Further, it does not appear to be adapted to deep water or long periods of daily inundation (Dykyjová, 1986; Coops et al., 1996; Clevering and Hundscheid, 1998; Deegan et al., 2005). Similarly, research by Lentz and Dunson (1998) suggests that the lifespan of Scirpus species is lowered with increased water level. Consequently, prolonged high water level was the main abiotic driver which reduced plant cover. Water salinity (30 - 35 ppt) and sediment salinity (13 - 15 ppt) were also high before die-back occurred. Although not identified as the significant driver, high salinity was related to low cover in the PCA plot. Kantrud (1996) states that growth is negatively affected by increased water salinity, while Kruger and Kirst (1991) recorded reduced growth of B. maritimus at 11 – 13 ppt. Seasonal salinity variations have been purported to affect the growing season of B. maritimus, whilst plant senescence occurs after long periods of high salinity (>15 ppt) (Hootsmans and Wiegman, 1998; Lillebø et al., 2003). Riddin and Adams (2010) also indicated that the reduction in sedge cover in the East Kleinemonde Estuary was correlated with high water level and salinity. Similarly, Lieffers and Shay (1982) recorded improved growth in shallow water with low salinity. In the POE, growth was seasonal, as recorded by Clevering and Hundscheid (1998) and Dykyjová (1986). Although salinity was not strongly correlated with plant cover in the PCA plot, it is most likely also an important abiotic driver, together with temperature. This is because both water column and sediment salinity were high for most of the sampling period i.e. >15 ppt, contributing to the low plant cover. Maximum live plant cover in the TOCE was significantly lower compared to the POE.

149

Mean seed production in B. maritimus during the first reproductive cycle in the TOCE was 11 221 seeds per m- 2, which is similar to a study by Diggory and Parker (2010) for a tidal marsh in the northern San Francisco Bay at 14 152 seeds m-². Flowering usually occurs in spring and summer (Dykyjová, 1986; Diggory and Parker, 2010), but in the TOCE flowering was delayed during the second reproduction period and occured aseasonally in response to a drop in water level. Inflorescence production was substantially lower than the previous flowering cycle due to die-back, which is supported in the research findings by Lentz and Dunson (1998). Dykyjová (1986) noted no flowering in plants submerged in 80 - 90 cm of water. Lieffers and Shay (1982) found that high water level in the previous year prevented seed production (Lieffers and Shay, 1981). Flowering decreased at sites that had been flooded continuously the previous year (Lieffers and Shay, 1982), as reflected in this study. In the POE, the two reproduction cycles were seasonal and coincided with the study by Diggory and Parker (2010). The number of inflorescences ranged from 54 m-2 in the TOCE and 12 - 14 m-2 in the POE, while in a study by Deegan and Harrington (2004), Schoenoplectus triqueter produced 75 - 402 m-2 with a pore water salinity ranging from 1 - 7 ppt respectively. In this study, both estuaries experienced substantially higher sediment salinity i.e. 9 - 27 ppt (TOCE) vs 13 - 23 ppt (POE). Reproductive output during the second reproduction period in the POE was significantly lower than the first reproduction period in the TOCE due to high sediment salinity (13 - 23 ppt) and low fecundity. On dry and highly saline sites, B. maritimus did not flower (Leiffers and Shay, 1982). Deegan et al. (2005) recorded that the reproduction of S. triqueter was significantly reduced at a salinity of 10 ppt. Another study found numerous inflorescences but very limited seed production due to the combined effects of inundation, salinity and sedimentation (Deegan and Harrington, 2004). Research by Coops and Smit (1991) demonstrated that salinity levels of 6 ppt stimulated spikelet and achene formation in B. maritimus. In addition, Charpentier et al. (2000) found that fecundity may be reduced due to clonal growth in small populations, which may also explain the limited achene production during the second flowering period in this study.

Based on the second reproduction cycle of both the estuaries and seed viability, B. maritimus required at least two to three months for flowers to set viable seeds. The germination trials in this study resulted in low germination percentages and no significant monthly changes in percentage germination. Low germination rates are characteristic of Scirpus acutus due to the thick pericarp of the achene, but not low viability as determined by various studies (Harris and Marshall, 1960; O'Neil, 1972; Lacroix and Mosher, 1995; Lombardi et al., 1997). Consequently, germination was low due to the hard pericarp. An overwintering period should enhance germination in the estuary the following spring (Lacroix and Mosher, 1995; Kantrud, 1996) because the germination trials of this study showed an increase from a mean of 3 % (March 2009 to August 2009) germination success to 36 % in seeds harvested in spring (September 2009) from plants that were submerged in the water column during winter.

5.2.6 Ruppia cirrhosa and Chara vulgaris R. cirrhosa displayed an annual life-cycle as it germinated in autumn and died back the following year in autumn, similar to that described by Gesti et al. (2005) of temporarily flooded areas in a coastal Mediterranean lagoon. The tendency to adopt an annual life history is common among aquatic macrophytes subjected to disturbances associated with fluctuating water levels or unpredictable environments (Stearns, 1976; Grime,

150 1979; Kautsky, 1988), such as in TOCEs. High water depth and pH were identified as the main abiotic drivers in the PCA plot. Although water temperature was not identified as a significant driver, high biomass accumulation increased significantly during spring and summer despite water depths being high two months prior, during winter. Germination also occurred when temperatures rose. Temperature is known to affect the germination of seagrasses (Adams and Riddin, 2007). A winter quiescence period was also observed by Menendez (2002) during in Spain and Congdon and McComb (1979) in Australia, where biomass peaked at 503 g DM m-2. Obrador and Petrus (2010) noted the same seasonal pattern and annual life-cycle in R. cirrhosa, with peak biomass ranging from 327 – 919 g DW m-2. The seasonal growth of R. maritimus in southern Brazil was also observed by Costa and Seeliger (1989) and a peak in biomass occurred during summer for R. cirrhosa in Denmark (Kiorboe, 1980). During 2006, the East Kleinemonde Estuary was closed for six months and the mean biomass reached 706 g DM m-2 (Riddin and Adams, 2008b). In this study, a peak biomass of 2 248 g DM m-2 was reached in summer after nine months while the mouth was closed. According to Riddin and Adams (2008b) the germination and expansion of submerged macrophytes is prevented by regular breaching, short periods (<6 months) of mouth closure and low water level (<1.5 m amsl). Water level during this study ranged from 1.6 - 2.5 m amsl and the mouth was closed for the entire sampling period, which promoted biomass accumulation. Riddin and Adams (2008a) related the expansion of R. cirrhosa in the estuary to water level rather than temperature because Ruppia species are known to survive water temperatures of 0 - 38°C and grow exponentially between 10 - 30°C (Verhoeven, 1979). However, in this study significant growth was only observed once temperatures were >17.5°C, despite the fact that water depths increased from 16 - 77 cm during winter, two months prior to reaching this temperature. Similarly, Asaeda (2007) states that in cold years with low spring water temperatures, exponential growth starts later in spring and peak biomass is reached later in Australia. In this study, the increase in temperature followed by the increase in water depth therefore favoured biomass accumulation. A study by Riss et al. (2000) found that Potamogeton communities were associated with high alkalinities, which is related to pH and may explain the significance of pH in this study. Other studies have also found this correlation (Pip, 1979; Catling et al., 1986; Arts et al., 1990; Vestergaard and Sand-Jensen, 2000; Riis and Biggs, 2001). Other studies have shown that R. cirrhosa can complete its life-cycle in three months (Adams and Bate, 1994c) or even two months (Gesti, 2005). However, in this study, the life-cycle was only completed after five months due to low winter temperatures and low water depths after germination. The extended period of mouth closure with high water level therefore enabled R. cirrhosa to complete its life-cycle over a longer period.

A study by Santamaria and Hootsmans (1998) on R. drepanensis found that temperature was important in determining reproduction (Warwick and Brock, 2003). Budding, flowering, germination and subsequent reproduction is also likely to be delayed by a decrease in temperature (Adams and Riddin, 2007). In this study, the reproductive cycle in R. cirrhosa was seasonal as it was related to temperature, as shown in the PCA plot. Although flowering commenced in late winter, it increased significantly from early to mid-spring. Other studies have observed a similar seasonal trend (Costa and Seeliger, 1989; Cho and Porrier, 2005). Cho and Porrier -2 (2005) reported a maximum of 40 inflorescences and 20 390 seeds per m² for 14.3 g DW m compared to this study of 8 097 flowering inflorescences and 26 242 fruit/seeds m-2 for 2 134 g DW m-2. Gesti et al. (2005) recorded substantially higher flower and fruit production for R. cirrhosa growing in temporarily flooded areas than in permanent waters, namely 2 781 - 3 456 flowers m-2 for 439 - 694 g DW m-2 versus 1 400 - 1 600 flowers for 95 - 137 g DW m-2 and a high fruit production (538 - 593 fruit m-2). The reproductive period was also

151 shorter in the temporary environment compared to the permanent environment i.e. 4 weeks versus 8 - 10 weeks. In this study, maximum flower production was 16 193 flowers m-2 for 2 134 g DW m-2, which is similar to that found by Gesti et al. (2005) and Bonis et al. (1993). The reproductive period however lasted much longer, namely 16 (fruiting period) to 42 weeks (flowering period). This may be a reflection of the plants plasticity in terms of taking advantage of favourable conditions because of the unpredictable and fluctuating water levels in a TOCE over the long term (Gesti et al., 2005), especially since the various stands in the estuary were at different phases in the reproductive cycle. Costa and Seeliger (1989) also suggest that high reproductive resource allocation may be required by Ruppia in unstable environments to ensure long term survival where desiccation stress occurs, such as in TOCEs. While Kantrud (1991) writes that annual Ruppia taxa depend on high fecundity to improve the likelihood of reproduction in temporary habitats. Seed emergence studies by Riddin and Adams (2009) on sediment collected from the East Kleinemonde Estuary showed that R. cirrhosa had a maximum germination percentage (11 %) at 35 ppt. Although seeds did not germinate in the germination trials of this study, it is considered unlikely that the seeds produced by the plants were not viable considering the results from Riddin and Adams (2009).

C. vulgaris showed seasonal biomass accumulation due to a significant increase during spring which was associated with the increase in water temperature and water depths, and as depicted in the PCA plot. Several other studies have reported seasonal responses (Stross, 1979; Grillas, 1990; Blindow, 1992; Casanova, 2000; Fernandez-Alaez et al., 2002, Espinar et al., 2002). Fernandez-Alaez et al. (2002) reported that the weather and hydrological regime strongly influenced the seasonal biomass patterns of Chara globularis Thuillier with -2 -2 peak biomass in summer (128 g DW m ) and autumn (165 g DW m ). In Australia, Chara fibrosa var. fibrosa (A. Br.) reached a biomass of 105 g DM m-2 (mean water depth 80 cm). Biomass was significantly lower in depths ≤50 cm than those in >50 cm and height was correlated to water depth (Asaeda et al., 2007). Cassanova and Brock (2002) found that field growth rates increased with increasing water depth. After six months, biomass only reached a maximum of 141 g DW m-2 in summer (mean water depth of 86 cm), compared to 599.5 g DM m-2 during a 6 month closued period in the estuary (Riddin and Adams, 2008b). C. vulgaris is considered non-salt tolerant (Winter and Kirst, 1990; Bonis et al., 1993) and thus the lower biomass was probably a response to the high water salinity (32 - 38.7 ppt) in this study, despite the high water level (Grillas, 1990). Charophytes in the Swartvlei Estuary only grew in areas where salinity was <16 ppt (Howard- Williams and Liptrot, 1980). Studies by Menendez and Sanchez (1998) and Wang et al. (2008) showed that high temperatures in summer decreased C. hispida and C. vulgaris respectively, which coincides with this study when biomass started declining at temperatures between 25 and 29°C in summer. Macrophyte biomass in temperate waters generally increases in spring, peaks in summer, decreases at the end of the summer or in autumn and reaches a minimum in winter (Fernandez-Alaez et al., 2002). Further, the high senescent rate at the end of summer may be attributed to the high light conditions caused by the shallower water (<40 cm) (Stross, 1979; Schwarz et al., 2002; Asaeda et al., 2007). In this study, low water level, high temperature and high salinity caused the die-back of C. vulgaris, while relatively low water depths (28 - 40 cm) combined with high light conditions probably contributed to the die-back of C. vulgaris.

Winter et al. (1987) observed a seasonal reproduction cycle in C. vulgaris with the formation of antheridia and oogenia in summer and the maturation of oogonia two to three months later in autumn. Similarly, this study showed a seasonal pattern with the formation of sexual organs commencing in winter, peaking in spring and

152 continuing into summer. C. australis was also shown to reproduce sexually in spring, summer and autumn (Casanova and Brock, 2000). In Australia, C. fibrosa produced 1 530 oogonia m-2 and 4 120 antheridia m-2 in shallow areas (<1 m) with peaks in late summer (Asaeda et al., 2007). In this study, C. vulgaris produced 196 998 oogonia m-2 and 46 016 antheridia m-2 in late spring (peak oogonia production). The charophyte seed bank in the East Kleinemonde Estuary has been estimated at 31 306 oospores m-2 (Riddin and Adams, 2009) compared to 29 000 - 417 700 oospores m-2 found in temporary marshes in France (Bonis et al., 1995). The latter seed bank correlates better with this study because the persistent seed bank is presumed to be considerably greater than annual production (Capers, 2003). In an experimental study by Bonis et al. (1993), C. aspera, C. canestens and C. contraria produced 12 - 25 g DW of oogonia. Similarly in this study, C. vulgaris produced 20 g DW of oogonia. Although C. vulgaris is not a saline species, and reproduction has been observed to decrease in high salinity (Bonis et al., 1993), the high oogonia and antheridia produced in this study compared to Asaeda et al. (2007) may be a function of the unstable TOCE environment (Kautsky, 1990). Saline macrophytes can produce abundant oogonia in as little as six weeks during floods due to stress, a reproduction strategy that is not often adopted by macrophytes in permanent lakes where there is little stress (Davis and Stevenson, 2007). Charophytes take 24 days to germinate in inundated sediments, which indicate that the water level needs to be stable for at least four months for these species to persist (Riddin and Adams, 2008a). According to Soulié-Marsche (1991a/b), the complete life-cycle of the charophyte, Lamprothamnium papulosum, required a minimum period of three months, but dependent on temperature and light conditions it may be considerably extended (Soulié-Marsche, 2008). Casanova and Brock (1990) found that C. corallina took 63 to 70 days (2 - 2.3 months) to develop oogonia. In this study, C. vulgaris took three months to complete its life-cycle. Further, temperature was an inhibiting factor because biomass accumulation and reproduction were delayed/slowed during a winter quiescence (Verhoeven, 1979), but the copious number of oogonia may be a function of the reproductive strategy of this species in the stressful TOCE environment (Kautsky, 1988; Casanova and Brock, 1990; Brock and Cassanova, 1991; Cassanova and Brock, 2000).

153 6. CHAPTER 6: CONCLUSIONS AND RECOMMENDATIONS

The hypothesis that macrophyte phenology in a TOCE was event driven rather than cyclical, which is due to the unpredictable nature of these estuaries was accepted. In less predictable environments such as TOCEs, especially when coupled with drier climates or regions prone to drought, the depth, duration and season (month) of flooding will affect the germination and establishment of macrophytes, including the completion of life-cycles through to sexual or asexual reproduction (Warwick and Brock, 2003). The other hypotheses were also accepted, namely that environmental conditions in a TOCE are more variable and stressful than environmental conditions in a POE, which are more stable; that macrophytes respond to relatively small water level fluctuations (10 – 20 cm) without the mouth breaching in a TOCE; that macrophytes are able to complete their life-cycles more rapidly in a TOCE compared to a POE; that macrophytes are highly plastic (i.e. adaptable and fexible) in response to fluctuating and stressful environmental conditions in a TOCE; that the intertidal macrophytes S. tegetaria and S. meyeriana require two months and one month respectively to set viable seeds after flowering and four and three months respectively after germination in a TOCE, and that submerged macrophytes require four months of stable high water level to set viable seeds after germination in a TOCE.

The East Kleinemonde Estuary was closed for 17 months which, combined with several sea storm events and reduced freshwater input due to drought, caused hypersaline conditions, high water level and anoxic reduced sediments. These conditions caused the emergent macrophytes to rapidly decline and die-back (PERL 1990; Zedler et al., unknown; Adams and Bate, 1994a; Riddin and Adams, 2008a). J. kraussii did not die-back completely due to its position higher up the elevation gradient. Similarly, in the Great Brak Estuary, freshwater releases from an impoundment on the river were not provided and the mouth closed for a year. Flooding and high water level reduced the cover of the supratidal and intertidal salt marsh species (Adams et al., 1992). In this study, the submerged macrophytes R. cirrhosa and C. vulgaris were able to complete their life-cycles and set viable seeds within five and three months respectively due to stable high water level. By the end of the sampling period, the emergent macrophyte cover was significantly lower than the POE macrophyte cover. Similarly, Adams et al. (1992) found that the diversity and cover of emergent macrophytes in two TOCEs was lower than that for two nearby POEs. This was attributed to fluctuating water levels and periodic hypersaline (>40 ppt) conditions in the TOCE, which reduced the presence of suitable habitat.

The seasonal spring and summer growth of the emergent macrophytes was interrupted due to an extended high water level in the TOCE. Growth resumed when the water retreated exposing bare areas in the supratidal zone from mid-summer to autumn, outside of the usual peak vegetative period (viza viz sping). In contrast, growth in the POE was seasonal and/or cyclical and comparatively stable for the majority of macrophytes, excluding the Sarcocornia hybrid. Plant cover in the Sarcocornia hybrid declined due to reduced sediments during the growth period, which was probably the result of high pH and salinity, including very compact clay soils moistened with rain water. In the TOCE, the intertidal macrophytes, S. tegetaria and S. meyeriana, responded to relatively small water level fluctuations by germinating within the month, and most likely within 3 – 5 days, of the estuary water level dropping from 2.3 to 2.1 m amsl. The water dropped in January 2010 by only 20 cm and exposed bare areas for germination to occur in summer. S. tegetaria and S. meyeriana plants were

154 able to complete their life-cycles within four and three months respectively, producing viable seeds within two and one month respectively after flowering. The study by Riddin and Adams (2008a) in the estuary roughly supports these results as the authors hypothesized that a minimum period of four months would be required for intertidal salt marsh to develop 100 % cover under ideal conditions if local extirpation occurred. In this study, intertidal salt marsh was able to re-establish a reasonable cover in the four months subsequent to the water retreating from the supratidal zone. In contrast, the upper intertidal species, S. decumbens, emerged later due to inundation and high salinity and germinated within either six days to one month after the water dropped in February 2010 by 11 cm during summer and autumn. S. decumbens plants took two months to produce viable seeds after flowering in the POE, but the time from germination to seed formation could not be established due to delayed and slow growth in the TOCE. In comparison, the study by Riddin and Adams (2008a) in the estuary predicted a minimum of three months, without a high water level, for supratidal salt marsh to establish after local extirpation. This means that a five month period should allow S. decumbens to produce viable seeds. Seedling emergence was observed until June 2010 (winter) reflecting an aseasonal germination pattern and it is likely that all three species were representative of the seedlings.

Although the estuary did not open during this study, recession of the water could be equated to a mouth opening event and the water level dropping. This would in turn allow germination, colonisation of bare areas and the completion of life-cycles within a short period. By completing the life-cycle rapidly during favourable conditions, plants avoid future stress which would prevent life-cycle completion (Kautsky, 1988; Blom, 1996). Conversely, the Sarcocornia hybrid and S. meyeriana in the POE took longer to complete their life-cycles compared to the TOCE, namely seven and nine months respectively, while S. tegataria did not germinate in situ but expanded vegetatively, although seed production occurred. The predominance for vegetative expansion in the POE is most likely a function of the relatively stable conditions in the established POE salt marsh compared to the TOCE where a combination of sexual and asexual reproduction ensures persistence (Kautsky, 1988; Desclaux and Roumet, 1996; Adams et al., 1999). Further, as disturbance increases predominance for sexual reproduction increases (Verhoeven, 1979; Kautsky, 1990; Casanova and Brock, 1996; Combroux and Bornette, 2004; Davis and Stevenson, 2007) which would apply to the dynamic TOCE environment as opposed to the more stable POE environment.

The reproduction periods were delayed and shortened due to inundation in the TOCE, but still occurred within the flowering periods typical of the emergent species. However, B. maritimus flowered during autumn and early winter in the supratidal habitat, outside of its normal flowering period viza viz spring and summer during the second reproduction period and was therefore aseasonal. In comparison, the reproduction periods in the POE were longer and also occurred within the expected seasons. During both reproductive cycles, the flowering period of S. meyeriana covered a one to three month period in the TOCE compared to a two month period in the POE. S. tegetaria flowered over a four and five month period in the TOCE compared to a five and six month period in the POE. S. decumbens flowered for four months during the first reproductive cycle in the TOCE and for six months in the POE during the second reproductive cycle. Despite this, S. tegetaria produced significantly more seeds and S. decumbens substantially more seeds in the TOCE compared to the POE. In contrast, S. meyeriana produced significantly more seeds in the POE than the TOCE, but this could be attributed to the morphology of the plants and the lower plant cover. Reproduction was therefore affected by both the depth and

155 duration of water level fluctuations (Brock and Casanova, 1990). The TOCE macrophytes were tolerant of long periods of inundation due to the ability to reproduce both sexually and vegetatively (Mony, 2010). Water regime therefore played a major role in macrophyte phenology in the TOCE. In contrast, the water regime in the POE played a minor role in the macrophyte phenology. The vegetative and sexual reproduction of the reed and sedge, P. australis and B. maritimus, were both significantly reduced in the TOCE compared to the POE, although seed production of B. maritimus in the POE was low due to high sediment salinity and low fecundity. The submerged macrophytes, R. cirrhosa and C. vulgaris, which were only studied in the TOCE on the other hand had long reproduction periods and seed production was significant during the study period due to the protracted period of stable high water levels. It can therefore be concluded that macrophyte phenology in the TOCE was significantly influenced by the closed mouth condition, high water level and relatively small water level fluctuations (2.3 - 2.1 m amsl). Consequently, macrophyte phenology was event driven in the TOCE compared to the POE were macrophyte phenology was cyclical, seasonal and/or comparatively stable. Although the phenology of the submerged macrophytes was seasonal, the typical seasonal nature of the phenological cycles corresponded with the high estuary water level due to closed mouth conditions and was extended due to these conditions.

The emergent macrophytes, P. australis, J. kraussii, S. virginicus, S. decumbens and S. tegetaria were able to withstand the extended period of inundation despite the significant reduction in plant cover. Once the water level retreated slightly (11 – 20 cm), the salt marsh plants resumed vegetative growth from plants that were submerged in high saline waters for the inundated period. Although all the new B. maritimus growth occurred in the exposed supratidal areas, one B. maritimus plant emerged in the inundated waters despite the high salinity and water level (29 cm) in June 2010 (i.e. at the end of the sampling period), after being absent from the flooded quadrats for one year. The macrophytes in the TOCE therefore demonstrated considerable plasticity in response to fluctuating environmental conditions and a protracted period of unfavourable and stressful conditions i.e. high water level and salinity. In contrast to the TOCE, competition with other species limited cover expansion because there were limited bare areas to expand to, due to the dense matrix of salt marsh cover and the stability of the POE environment (Snow and Vince, 1984; Bertness and Ellison, 1987; Pennings and Callaway, 1992; Streevers and Genders, 1997). The macrophyte phenology was therefore typically seasonal, cyclical and/or comparatively stable in comparison to the dynamic responses found in the TOCE.

Where TOCEs are managed to keep the mouth open at certain times of the year, two components that must be considered to ensure that the salt marsh and submerged macrophytes are kept in a healthy state, are (1) water level; and (2) the season (timing) and duration of a particular water level. Firstly, the estuary must be at a water level that will provide sufficient exposed intertidal and/or supratidal areas for emergent macrophytes to complete their life-cycles i.e. macrophytes must not be inundated for a protracted period otherwise they will not be able to flower and produce viable seeds. In contrast, the submerged macrophytes require a stable high water level for a particular length of time to complete their life-cycles. The second component, namely the season (timing) and duration of a particular water level, must be considered because macrophytes require a certain period (duration) of warmer temperatures (season) to ensure that adequate vegetative production and sexual reproduction take place to complete their life-cycles and produce viable seeds.

156 Vegetative production and sexual reproduction in all the emergent macrophytes overlapped over the warmer months i.e. from November to March, therefore the timing of lower water level should occur over this period. Although seeds germinated from summer to winter in this study, the best time for seed germination is more than likely during spring/summer (Riddin and Adams, 2008b). Warmer temperatures in spring/summer should also enhance vegetative growth. Further, the intertidal salt marsh in the East Kleinemonde Estuary should take approximately four months to become re-established (Riddin and Adams, 2008a), which coincides with this study because S. tegetaria and S. meyeriana took a maximum of four and three months respectively to complete their life-cycles and attain good plant cover. Although this study was unable to determine the time from germination to seed formation in S. decumbens, the recommended time period of four to five months should also ensure that the supratidal salt marsh recovers. This is because Riddin and Adams (2008a) report that supratidal salt marsh should take a minimum of three months to re-establish, after complete loss without high water level, in the East Kleinemonde Estuary. The recommended time period for maintaining lower water level will provide a maximum five month period for S. decumbens to re-establish cover in order to produce viable seeds after at least two months of flowering commencing. This period will also provide a window of opportunity for seed germination to occur in S. decumbens, S. tegetaria and S. meyeriana.

The submerged macrophyte R. cirrhosa commenced flowering in early spring/early September and sexual reproduction increased significantly during spring from September to October, as temperatures rose. Oospore production of C. vulgaris commenced in late winter/early August and showed substantial increases during spring as temperature increased. Both peak seed and oospore production occurred in late spring/early November. Short periods (<6 months) of closed mouth conditions and regular breaching (>2.6 times a year) with water level <1.5 m amsl prevent the germination and expansion of submerged macrophytes (Riddin and Adams, 2008b). Consequently, the estuary, under normal conditions, is closed for most of the year and needs to be closed for more than six months at a high water level for submerged species to establish. According to Riddin and Adams (2008b), at water levels >1.5 m msl the submerged habitat becomes established after an inundation period of one to two months. R. cirrhosa has been observed to germinate and rapidly complete their life-cycle within three months (Adams and Bate, 1994c; Riddin and Adams, 2010), while Gesti (2005) found that R. cirrhosa completed its life-cycle in less than two months. However, budding, flowering, germination and subsequent reproduction in R. cirrhosa are also likely to be delayed by a decrease in temperature (Adams and Riddin, 2007). In this study, R. cirrhosa germinated in mid-autumn (April), took five months to produce 69 seeds m-² (early September) and seven months to reach peak seed production at 26 242 seeds m-² (early November). C. vulgaris geminated in late autumn (May), took three months to produce 1 206 oogonia m-² (early August) and six months to reach peak oogonia production at 196 998 oogonia m-² (early November). Closed mouth conditions and an associated high water level therefore favoured the submerged macrophytes. The main abiotic driver was water level, but because it was constant for a protracted period as a result of the closed mouth conditions, water temperature was also an important abiotic driver.

157 Consequently, when the salt marsh and reed and sedge habitats in TOCEs have been extirpated or severely degraded due to a protracted closed mouth condition caused by anthropogenic disturbances (e.g. dam construction), the mouth should be opened to ensure a low water level for a minimum of four to five months in order for the habitats to recover. It is therefore recommended that the following management protocols be applied to anthropogenically impacted TOCEs:

(a) When the estuary water level inundates the intertidal and supratidal habitat for more than 8 - 12 months, and the macrophytes have been extirpated or severely degraded, the mouth should be opened. One breach event is recommended to reduce the water level that inundates the intertidal habitat (in the East Kleinemonde Estuary this threshold was >1.5 - 1.6 m amsl) and supratidal habitat (in the East Kleinemonde Estuary this threshold was >1.8 – 2.1 m amsl).

(b) Low water level should be maintained for at least four to five months from late spring to early autumn i.e. mid November to March in order for intertidal and supratidal salt marsh plants to germinate and produce viable seeds. The intertidal macrophytes, S. tegetaria and S. meyeriana required four and three months respectively to germinate from seed, establish as seedlings and reach reproductive maturity in this study. Increases in water level and inundation of seedlings will prevent their establishment given their susceptibility to the slightest inundation, for example in this study a rise in water level of 1 cm killed two-leaf stage seedlings (height of ± 0.5 - 1 cm).

(c) If necessary, a maximum of two breach events can be implemented if a high water level is reached again during the recommended period (e.g. due to high rainfall and storm surges).

(d) As a minimum requirement, and if submerged macrophytes require more time to produce viable seeds, as reflected in this study, water level should not be higher than the supratidal areas during the recommended period. This is because the emergent macrophytes should be able to re-colonize the areas higher upper the elevation gradient, as demonstrated in this study (in the East Kleinemonde Estuary this threshold was ≤2.1 m amsl). This will also prevent the deterioration of other supratidal macrophytes, such as J. kraussii and S. virginicus and the reed and sedge macrophytes, P. australis and B. maritimus.

(e) The estuary water level should be allowed to inundate the supratidal habitat occasionally, or at least once a year, particularly if drought is experienced and sediment salinity is high. This, coupled with rainfall, is important to ensure that accumulated salts are flushed out from the upper marsh area (Adams et al., 1999; CSIR, 2003).

158 7. REFERENCES

Adam, P. 1990. Salt Marsh Ecology. Cambridge University Press, Cambridge, UK. 461 pp.

Adam, P. 2002. Salt marshes in a time of change. Environmental Conservation 29(1) 39–61.

Adams, J.B. 1991. The distribution of estuarine macrophytes in relation to freshwater in a number of eastern Cape estuaries. MSc Thesis. University of Port Elizabeth, South Africa.177 pp.

Adams, J.B., Knoop, W.T. and Bate, G.C. 1992. The distribution of estuarine macrophytes in relation to freshwater. Botanica Marina 35: 215-226.

Adams, J.B. and Bate, G.C. 1994a. The effect of salinity and inundation on the estuarine macrophyte Sarcocornia perennis. Aquatic Botany 47: 341-348.

Adams, J.B. and Bate, G.C. 1994b. The freshwater requirements of estuarine plants incorporating the development of an estuarine decision support systems. Report to the Water Research Commission by the Department of Botany, University of Port Elizabeth.

Adams, J.B. and Bate, G.C. 1994c. The tolerance to desiccation of the submerged macrophytes Ruppia cirrhosa Petagna (Grande) and Zostera capensis Setchell. Journal of Experimental Marine Biology and Ecology 183: 53-62.

Adams, J.B. 1994. The importance of freshwater to the survival of estuarine plants. PhD Thesis. University of Port Elizabeth, Port Elizabeth. 172 pp.

Adams, J.B., Bate, G.C. and O‘Callaghan, M. 1999. Primary producers. In: Allanson B.R., and Baird, D. (Eds.), Estuaries of South Africa. Cambridge University Press: 91-112.

Adams, J.B. and Bate, G.C. 1999. Growth and photosynthetic performance of Phragmites australis in estuarine waters: A field and experimental evaluation. Aquatic Botany 64: 359–367.

Adams J.B. and Riddin, T. 2007. Macrophytes. In: Whitfield, A. and Bate, G.C. 2007. A review of information on temporarily open/closed estuaries in the warm and cool temperate biogeographic regions of South Africa with particular emphasis on the influence of river flow on these systems. WRC Report 1581/1/07. Water Research Commission, Pretoria, South Africa.

Al-Busaidi, A.S. and P. Cookson, 2003. Salinity-pH relationship in calcareous soils. SQU Journal for Scientific Research: Agricultural and Marine Sciences 8(1): 41-46.

Aldous, D.E. 2003. Accessions to elevated salt concentrations. ISHS Acta Horticulturae 661: I International Conference on Turfgrass Management and Science for Sports Fields.

Alexander, H. and Dunton, K.H. 2002. Freshwater inundation effects on emergent vegetation of a hypersaline salt marsh. Estuaries 25(6): 1426-1435. 159 Allanson, B.R. and Whitfield, A.K. 1983. Limnology of the Touw River floodplain. Cooperative Scientific Programmes, CSIR, SANSP Report 79. pp 40.

Allanson, B. and Baird, D. 1999. Estuaries of South Africa. Cambridge University Press. 340 pp.

Allanson, B.R. 2000. The Knysna Basin Project reviewed: Research findings and implications for management. In Hodgson, A.N. and Allanson, B.A. (eds). The Knysna Basin Proiect 1995-1998: A scientific report on the Knysna Estuary. Transactions of the Royal Society of South Africa 55: 97-105.

Allison, S.K. 2006. Recruitment and establishment of salt marsh plants following disturbance by flooding. The American Midland Naturalist 136: 232-247.

Álvarez-Rogel, J., Alcaraz, F. and Ortiz, R. 2000. Soil and moisture gradients and plant zonation in Mediterranean salt marshes of south-east Spain. Wetlands 20: 357-372.

Andersen, J.M. 1976. An ignition method for determination of total phosphorus in lake sediments. Water Research 10: 329–331.

Anderson, M.R. and Kalff, J. 1988. Submerged aquatic macrophyte biomass in relation to sediment characteristics in ten temperate lakes. Freshwater Biology 19: 115–121.

Armstrong, W., Wright, E.J., Lythe S. and Gaynard T.J. 1985. Plant zonation and the effects of the spring-neap tidal cycle on soil aeration in a Humber salt marsh. Journal of Ecology 73: 3232-3393.

Armstrong, J., Afreen-Zobayed, F. and Armstrong, W. 1996. Phragmites die-back: Sulphide and acetic acid induced bud and root death, lignifications and blockages within aeration and vascular systems. New Phytologist 134: 601-614.

Arts, G.H.P., Roelofs, J.G.M. and de Lyon, M.J.H. 1990. Differential tolerances among soft-water macrophyte species to acidification. Canadian Journal of Botany 68: 2127–2134.

Asaeda, T., Rajapakse, L. and Sanderson, B. 2007. Morphological and reproductive acclimations to growth of two charophyte species in shallow and deep water. Aquatic Botany 86: 393–401.

Ashraf, M. 2004. Some important physiological selection criteria for salt tolerance in plants. Flora 199: 361-376.

Auld, B.A. and Medd, R.W. 1987. Weeds: An Illustrated Botanical Guide to the Weeds of Australia. Inkata Press, Melbourne, Australia. 255 pp.

Austenfeld, F.A. 1988. Seed dimorphism in Salicornia europaea: Nutrient reserves. Physiologia Plantarium 73: 502–504.

Badenhorst, P. 1988. Report on the dynamics of the Kleinemonde West and East estuaries (CSE 13 and 14). CSIR Report EMA-T 8805. 31 pp.

160 Bai, J., Ouyang, H., Deng, W., Zhu, Y., Zhang, X. and Wang, Q. 2005. Spatial distribution characteristics of organic matter and total nitrogen of marsh soils in river marginal wetlands. Geoderma 124: 181-192.

Baird, D., Marais, J.F.K. and Martin, A.P. (Editors). 1988. Swartkops Estuary. Proceedings of a symposium: 14- 15 September 1987. University of Port Elizabeth. National Scientific Programmes Unit: CSIR, SANSP Report 156, pp 112.

Baldwin, A.H., McKee, K.L. and Mendelssohn, I.A. 1996. The influence of vegetation, salinity, and inundation on seed banks of oligohaline coastal marshes. American Journal of Botany 83: 470–479.

Baldwin, A.H. and Mendelssohn, I.A. 1998. Effects of salinity and water level on coastal marshes: An experimental test of disturbance as a catalyst for vegetation change. Aquatic Botany 61: 255–268.

Baldwin, A.H., Kettenring, K.M., and Whigham, D.F. 2010. Seed banks of Phragmites australis-dominated brackish wetlands: Relationships to seed viability, inundation, and land cover. Aquatic Botany 93: 163–169.

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., Adams, M.S. and Clesceri, N.L. 1986. Environmental factors and their consideration in the management of submersed aquatic vegetation. Journal of Management 24: 1-10.

Barko, J.W. and Smart, R.M. 1986. Sediment-related mechanisms of growth limitation in submersed macrophytes. Ecology 67(5): 1328-1340.

Barnard, R.O. 1990. Handbook of standard sediment testing methods for advisory purposes. Sediment Science Society of South Africa, Pretoria. 35 pp.

Baskin, C.C. and Baskin, J.M. 1979. The seed bank in a population of an endemic plant species and its ecological significance. Biological Conservation 14: 125-130.

Baskin, C.C. and Baskin, J.M. 1983. Germination ecology of Veronica arvensis. Journal of Ecology 71: 57-68.

Baskin, C.C. and Baskin, J.M. 1984. Role of temperature in regulating timing of germination in soil seed reserves of Lamium purpureum L. Weed Research 24: 341-349.

Baskin, C.C. and Baskin, J.M. 1985. The annual dormancy cycle in buried weed seeds: A continuum. BioScience 35: 492-498.

Baskin, C.C. and Baskin, J.M. 1986. Change in dormancy status of Frasera caroliniensis seeds during overwintering on the parent plant. American Journal of Botany 73(1): 5-10.

Baskin, C.C., Baskin, J.M. and Spooner, D.M. 1989. Role of temperature, light and date: Seeds were exhumed from soil on germination of four wetland perennials. Aquatic Botany 35: 387-394.

161 Baskin, C.C. and Baskin, J.M. 1998. Seeds: Ecology, Biogeography, and Evolution of Dormancy and Germination. Academic Press, San Diego, CA, USA.

Baskin, C.C. and Baskin, J.M. 2004. Determining dormancy-breaking and germination requirements from the fewest seeds. pp. 162-179. In: Guerrant, E., Havens, K. and Maunder, M. (Eds). Strategies for Survival. Island Press, Washington. 504 pp.

Beck, N.G. and Bruland, K.W. 2000. Diel biogeochemical cycling in a hyperventilating shallow estuarine environment. Estuaries 23: 177–187.

Beeftink, W.G. 1985. Population dynamics of annual Salicornia species in the tidal salt marshes of the Oosterschelde, the Netherlands. Vegetatio 61 (1/3): 127-136.

Begg, G.W. 1984. The Estuaries of Natal. Part 2. Natal Town and Regional Planning Report 55. 631 pp.

Bell, H.L. and O'Leary, J.W. 2003. Effects of salinity on growth and cation accumulation of Sporobolus virginicus (Poaceae). American Journal of Botany 90(10): 1416-1424.

Benfield, M.C. 1984. Some factors influencing the growth of Phragmites australis (Cav.) Trin ex Steudal. MSc. Thesis. University of Natal, Durban. 84 pp.

Bennett, B.A. 1989. A comparison of the fish communities in nearby permanently open, seasonally open and normally closed estuaries in the south-western Cape, South Africa. South African Journal of Marine Science 8: 43-55.

Bertness, M.D. and Ellison, A.M. 1987. Determinants of pattern in a New England salt marsh plant community. Ecological Monographs 57:129-147.

Bertness, M.D. 1991. Interspecific interactions among high marsh perennials in the New England salt-marsh. Ecology 72: 125-137.

Bertness, M.D., Gough, L. and Shumway, S.W. 1992. Salt tolerances and the distribution of fugitive salt marsh plants. Ecology 73: 1842– 1851.

Bertness, M.D. and Shumway, S.W. 1993. Competition and facilitation in marsh plants. American Naturalist 142: 718-724.

Black, C.B. 1965. Methods of Soil Analysis. Society of Agronomy, Inc., Publisher, Madison. 1000 pp.

Blanch, S.J., Ganf, G.G. and Walker, K.F. 1999. Growth and resource allocation in response to flooding in the emergent sedge Bolboschoenus medianus. Aquatic Botany 63: 145–160.

Blindow, I. 1992. Decline of charophytes during eutrophication: Comparison with angiosperms. Freshwater Biology 28: 9–14.

162 Blindow, I., Dietrich, J., Mollmann, N. and Schubert, H. 2003. Growth, photosynthesis and fertility of Chara aspera under different light and salinity conditions. Aquatic Botany 76: 213–234.

Blindow, I., Möllmann, N., Boegle, M.G. and Schütte, M. 2009. Reproductive isolation in Chara aspera populations. Aquatic Botany 91(3): 224-230.

Bliss, S.A and Zedler, P.H. 1998. The germination of seedlings in vernal pools: Sensitivity to environmental conditions and effects on community structure. Oecologia 113: 67-73.

Blom, C.W.P.M. and Voesenek, L.A.C.J. 1996. Flooding: The survival strategies of plants. Tree 2(7): 290-295.

Biggs, B.J.F. 1996. Hydraulic habitat of plants in streams. Regulated Rivers: Research and Management 12 (2- 3): 131-144.

Bockelmann, A.C., Bakker, J.P., Neuhaus, R. and Lage, J. 2002. The relation between vegetation zonation, elevation and inundation frequency in a Wadden Sea salt marsh. Aquatic Botany 73: 211-221.

Boedeltje, G., Ter Heerdt, G.N.J. and Bakker, J.P. 2002. Applying the seedling-emergence method under waterlogged conditions to detect the seed bank of aquatic plants in submerged sediments. Aquatic Botany 72: 121–128.

Boedeltjie, G., Bakker, J.P., Ten Brinke, A., van Groenendael, J.M. and Soesbergen, M. 2004. Dispersal phenology of hydrochorous plants in relation to discharge, seed release time and buoyancy of seeds: The flood pulse concept supported. Journal of Ecology. 92: 786–796.

Boeglea, M.G., Schneiderb, S.C., Schubertc, H. and Melzera. A. 2010. Chara baltica Bruzelius 1824 and Chara intermedia A. Braun 1859—Distinct species or habitat specific modifications? Aquatic Botany 93: 195–201.

Bok, A.H. 1983. The demography, breeding biology and management of two mullet species (Pisces: Mugilidae) in the eastern Cape, South Africa. PhD Thesis. Rhodes University. 268 pp.

Bolen, E. 1964. Plant ecology of spring-fed salt marshes in western Utah. Ecology Monographs 34: 143–166.

Bonis, A., Grillas, P., van Wijck, C. and Lepart, J. 1993. The effect of salinity on the reproduction of coastal submerged macrophytes in experimental communities. Journal of Vegetation Science, 4 (4): 461-468.

Bonis, A. and Lepart, J. 1994. Vertical structure of seed banks and the impact of depth of burial on recruitment in two temporary marches. Vegetation 112: 127–139.

Bonis, A., Lepart, J. and Grillas, P. 1995. Seed bank dynamics and coexistence of annual macrophytes in a temporary and variable habitat. Oikos 74: 81-92.

Bonis, A., and Grillas, P. 2002. Deposition, germination and spatio-temporal patterns of charophyte propagule banks: A review. Aquatic Botany 72: 235–248.

163 Boorman, L.A. 1999. Salt marshes - Present functioning and future change. Mangrove Salt Marshes 3: 227- 241.

Bornman, T.G. 2002. Freshwater requirements of supratidal and floodplain salt marsh on the west coast, South Africa. PhD Thesis. University of Port Elizabeth, South Africa. 243 pp.

Bornman. T.G., Adams, J.B. and Bate, G. 2004. The influence of floodplain geohydrology on the distribution of Sarcocornia pillansii in the Olifants Estuary on the west coast, South Africa. Journal of Arid Environments 56: 603-625.

Bornman, T.G., Adams, J.B. and Bate, G.C. 2008. Environmental factors controlling the vegetation zonation patterns and distribution of vegetation types in the Olifants Estuary, South Africa. Journal of Botany 74: 685–695.

Bouzille, J. B., Bonis, A., Clement, B. and Godeau, M. 1997. Growth patterns of Juncus gerardi clonal populations in a coastal habitat. Plant Ecology 132: 39–48.

Bouzillé, J.B., Kernéis, E. and Touzard, B. 2001. Vegetation and ecological gradients in abandoned salt pans in western France. Journal of Vegetation Science 12: 269-278.

Bray, S. and Officer, D. 2007. Weedy Sporobolus grasses. Best Practice Manual. Department of Primary Industries and Fisheries. Brisbane. Australia.

Breen, C.M., Everson, C. and Rogers, K. 1977. Ecological studies on Sporobolus virginicus (L.) Kunth with particular reference to salinity and inundation. Hydrobiologia 54: 135–140.

Breen, C.M. and McKenzie, M. (Eds). 2001. Managing Estuaries in South Africa: An Introduction. Institute of Natural Resources. Pietermaritzberg. 66 pp.

Brewer, J.S., Levine, J.M. and Bertness, M.D. 1997. Effects of biomass removal and elevation on species richness in a New England salt marsh. Oikos 80(2): 333-341.

Brix, H. and Schierup, H. 1989. The use of aquatic macrophytes in water pollution control. Ambio 18: 100–107.

Briggs, D. 1977. Sources and Methods in Geography. Butterworths, London. 123 pp.

Britton, D.L. and Brock, M.A. 1994. Seasonal germination from wetland seed banks. Australian Journal of Marine Freshwater Research 45: 1445–1458.

Brock, M.A. 1982. Biology of the salinity tolerant genus Ruppia L. in saline lakes in south Australia II. Population ecology and reproductive biology. Aquatic Botany 19: 249-268.

Brock, T.C.M., van der Velde, G. and Van de Steeg, H.M., 1987. The effects of extreme water level fluctuations on the wetland vegetation of a nymphaeid-dominated oxbow lake in The Netherlands. Archiv für Hydrobiologia 27: 57-73.

164 Brock, M.A. and Casanova, M.T. 1991. Plant survival in temporary waters: A comparison of charophytes and angiosperms. Verhandlungen des Internationalen Verein Limnologie (Limnology) 24: 2668–2772.

Brock, M.A. and Rogers, K.H. 1998. The regeneration potential of the seed bank of an ephemeral floodplain in South Africa. Aquatic Botany 61: 123-135.

Brock, M.A., Nielsen, D.L. and Russell J.S. 2003. Drought and aquatic community resilience: The role of eggs and seeds in sediments of temporary wetlands. Freshwater Biology 48(7): 1207–1218.

Brown, R.A. 1990. Strangford Lough. The Wildlife of an Irish Sea Lough. The Institute of Irish Studies, Queens University of Belfast. 228 pp.

Brown, K. and Brooks, K. 2002. Bushland Weeds: A practical guide to their management. Environmental Weeds Action Network, Greenwood.

Bulthuis, D.A. 1983. Effects of temperature on the photosynthesis-irradiance curve of the Australian seagrass, Heterozostera tasmanica. Marine Biology Letters 4: 47-57.

Bunn, S.E. and Arthington, A.H. 2002. Basic principles and ecological consequences of altered flow regimes for aquatic biodiversity. Environmental Management 30(4) 492–507.

Bush, J.K. 2006. The role of soil moisture, salinity, and oxygen on the growth of Helianthus paradoxus (Asteraceae) in an inland salt marsh of west Texas. Journal of Arid Environments 64: 22–36.

Björk, S. 1967. Ecological investigations of Phragmites communis: Studies in theoretical and applied limnology. Folia Limnologica Scandinavica 14: 1–248.

Calado, G. and Duarte, P. 2000. Modeling growth of Ruppia cirrhosa. Aquatic Botany 68: 29–44.

Callaway, R.M. 1994. Facilitative and interfering effects of Arthrocnemum subterminale on winter annuals. Ecology 75: 681-686.

Callaway, R.M. and Subraw, C.S. 1994. Effects of variable precipitation on the structure and diversity of a California salt marsh community. Journal of Vegetation Science 5: 433–438.

Campbell, J.J. 2005. Influence of environmental factors on the seed ecology of Vallisneria americana. MSc. Thesis. The College of William and Mary in Virginia, Virginia. 167 pp.

Cantero, J.J., Leon, R., Cisneros, J.M. and Cantero, A. 1998. Habitat structure and vegetation relationships in central Argentina salt marsh landscapes. Plant Ecology 137: 79-100.

Capers, R.S. 2003. Macrophyte colonization in a freshwater tidal wetland (Lyme, CT, USA). Aquatic Botany 77: 325–338.

165 Carruthers, T.J.B. and Walker, D.I. 1999. Sensitivity of transects across a depth gradient for measuring changes in aerial coverage and abundance of Ruppia megacarpa Mason. Aquatic Botany 65: 281–292.

Casanova, M.T. and Brock, M.A. 1990. Charophyte germination and establishment from the seed bank of an Australian temporary lake. Aquatic Botany 36(3): 247-254.

Casanova, M.T. 1994. Vegetative reproduction response of charophytes to water-level fluctuations in permanent and temporary wetlands in Australia. Australian Journal of Marine and Freshwater Research 45: 1409–1419.

Casanova, M.T. and Brock, M.A. 1996. Can oospore germination patterns explain Charophyte distribution in permanent and temporary flooded wetlands? Aquatic Botany 54: 297–312.

Casanova, M.T. and Brock, M.A. 1999. Life histories of charophytes from permanent and temporary wetlands in eastern Australia. Australian Journal of Marine and Freshwater Research 47: 383–397.

Casanova, M.T. and Brock, M.A. 2000. How do depth, duration and frequency of flooding influence the establishment of wetland plant communities? Plant Ecology 147: 237–250.

Castellanos, E.M., Figueroa, M.E. and Davy, A.J. 1994. Nucleation and facilitation in salt-marsh succession: interactions between Spartina maritima and Arthrocnemum perenne. Journal of Ecology 82: 239–248.

Castillo, J. M., Fernandez-Baco, L., Castellanos, E.M., Luque, C.J., Figueroa, M.E. and Davy, A.J. 2000. Lower Limits of Spartina densiflora and S. maritima in a Mediterranean salt marsh determined by different eco-physiological tolerances. The Journal of Ecology 88(5): 801-812.

Castroviejo, S. 1990. LII. Chenopodiaceae. pp. 476–553. In: Castroviejo, S., Lainz, M., López González, G., Montserrat, P., Muñoz Garmendia, F., Paiva, J. and Villar, L. (eds.). Flora Iberica 2. Real Jardín Botánico, Madrid.

Catling, P.M., Freedman, B., Stewart, C., Kerekes, J.J. and Lefkovitch, L.P. 1986. Aquatic plants of acidic lakes in Kejimkujik National Park, Nova Scotia: Floristic composition and relation to water chemistry. Canadian Journal of Botany 64: 724–729.

Chambers, P.A. and Kalff, J. 1985. Depth distribution and biomass of submerged aquatic macrophyte communities in relation to secchi depth. Canadian Journal of Fisheries and Aquatic Science 42: 701–709.

Chambers R.M., Mozdzer T.J. and Ambrose J.C. 1998. Effects of salinity and sulfide on the distribution of Phragmites australis and Spartina alterniflora in a tidal salt marsh. Aquatic Botany 62: 161-169.

Chambers, R.M., Osgood, D. T., Bart, D.J. and Montalto, F. 2003. Phragmites australis invasion and expansion in tidal wetlands: Interactions among salinity, sulfide, and hydrology wetlands: Estuaries 26(2B) 398-406.

Chapin, F.S., Bloom, A.J., Field, C.B. and Waring, R.H. 1987. Plant responses to multiple environmental factors. BioScience 37: 49–57.

166 Charpentier, A., Grillas, P. and Thompson, J.D. 2000. The effects of population size limitation on fecundity in mosaic populations of the clonal macrophyte Scirpus maritimus (Cyperaceae). American Journal of Botany 87(4): 502-507.

Charpentier, A., Mesléard, F. and Grillas, P. 2009. The role of water level and salinity in the regulation of Juncus gerardi populations in former rice fields in southern France. Journal of Vegetation Science 9 (3): 361-370.

Chapman, V.J. 1974. Salt Marshes and Salt Deserts of the World. Second Edition. Lehre, Germany. 392 pp.

Chapman, V.J. 1976. Coastal Vegetation. Second Edition. Permagon Press, New York. 292 pp.

Chen, H., Qualls, R.G. and Miller, G.C. 2002. Adaptive responses of Lepidium latifolium to soil flooding: biomass allocation, adventitious rooting, aerenchyma formation and ethylene production. Environmental and Experimental Botany 48: 119–128.

Cho, H.J. and Poirrier, M.A. 2005. Seasonal growth and reproduction of Ruppia tilizat L. in Lake Pontchartrain, Louisiana, USA. Aquatic Botany 81: 37–49.

Chuwen, B.M., Hoeksema, S.D. and Potter, I.C. 2009. The divergent environmental characteristics of permanently-open, seasonally-open and normally-closed estuaries of south-western Australia. Estuarine, Coastal and Shelf Science 85:12–21.

Cizkova-Koncalova, H., Kvet, J. and Thompson, K. 1992. Carbon starvation: A key to reed decline in eutrophic lakes. Aquatic Botany 43: 105-113.

Clarke, L.D. and Hannon, N.J. 1970. The mangrove swamp and salt marsh communities of the Sydney district. III Plant growth in relation to salinity and waterlogging. Journal of Ecology 56: 351-369.

Clarke, L.D. and Hannon, N.J. 1971. The mangrove swamp and salt marsh communities of the Sydney district. IV. The significance of species interactions. Journal of Ecology 59: 535-553.

Clarke, P.J. and Jacoby, C.A. 1994. Biomass and above-ground productivity of salt-marsh plants in south- eastern Australia. Australian Journal of Marine and Freshwater Resources 45: 1521-1528.

Clevering, O.A. 1995. Germination and seedling emergence of Scirpus lacustris L. and Scirpus maritimus L. with special reference to the restoration of wetlands. Aquatic Botany 50: 63–78.

Clevering, O.A., van Vierssen, W. and Blom, C.W.P.M. 1995. Growth, photosynthesis and carbohydrate utilization in submerged Scirpus maritimus L. during spring growth. New Phytologist 130: 105-116.

Clevering, O.A., Blom, C.W.P.M. and van Vierssen, W. 1996. Growth and morphology of Scirpus lacustris and S. maritimus seedlings as affected by water level and light availability. Functional Ecology 11: 289-296.

167 Clevering, O.A. and van Gulik, W.M.G. 1997. Restoration of Scirpus lacustris and Scirpus maritimus stands in a former tidal area. Aquatic Botany 55: 229-246.

Clevering, O.A. and Hundscheid, M.P.J. 1998. Plastic and non-plastic variation in growth of newly established clones of Scirpus (Bolboschoenus) maritimus L. grown at different water depths. Aquatic Botany 62: 1-17.

Clevering, O.A. 1999. The effects of litter on growth and plasticity of Phragmites australis clones originating from infertile, fertile or eutrophicated habitats. Aquatic Botany 64: 35–50.

Coetzee, J.C., Adams, J.B. and Bate, G.C. 1997. A botanical importance rating of selected Cape estuaries. Water SA 23: 81 - 93.

Coleman, J.S., McConnaughay, K.D.M. and Ackerly, D.D. 1994. Interpreting phenotypic variation in plants. Tree 9: 187–191.

Colmer, T.D. and Flowers, T.J. 2008. Flooding tolerance in halophytes. New Phytologist, 179: 964–974.

Collins, C.D., Sheldon, R.B. and Boylen, C.W. 1987. Littoral zone macrophyte community structure: Distribution and association of species along physical gradients in Lake George, New York, U.SA. Aquatic Botany 29: 177-194.

Combroux, I.C.S. and Bornette, G. 2004. Propagule bank regeneration strategies of aquatic plants. Journal of Vegetation Science 15: 13–20.

Congdon, R.A. and McComb, A.J. 1979. Productivity of Ruppia: Seasonal changes and dependence on light in an Australian estuary. Aquatic Botany 6: 121–132.

Congdon, R.A. and McComb, A.J. 1980. Productivity and nutrient content of Juncus kraussii in an estuarine marsh in south-western Australia. Australian Journal of Ecology (1980) 5: 221-234.

Cooper, A. 1982. The effects of salinity and waterlogging on the growth and cation uptake of salt marsh plants. New Phytologist 90: 263-275.

Cooper, J.A.G. 2001. Geomorphological variability among microtidal estuaries from the wave dominated South African coast. Geomorphology 40: 99–122.

Coops, H. and Smit, H. 1991. Effects of various water depths on Scirpus maritimus L.: Field and experimental observations. Internationale Vereinigung fuer Theoretische und Angewandte Limnologie Verhandlungen 24: 2706-2710.

Coops, H., van den Brink, F.W.B. and van der Velde, G. 1996. Growth and morphological responses of four helophyte species in an experimental water-depth gradient. Aquatic Botany 54: 11-24.

Coops, H. 2002. Editorial. Ecology of charophytes: An introduction. Aquatic Botany 72: 205–208.

168 Cook, C.D.K. 2004. Aquatic and Wetland Plants of Southern Africa. Backhuys Publisher. Netherlands. 281 pp.

Costa, C.S.B. and Seeliger, O. 1989. Vertical distribution and resource allocation of Ruppia maritime L. in a southern Brazilian estuary. Aquatic Botany 33: 123-129.

Costa, C.S.B., Marangoni, J.C. and Azevedo, A.M.G. 2003. Plant zonation in irregularly flooded salt marshes: relative importance of stress tolerance and biological interactions. Journal of Ecology 91: 951-965.

Cowardin, L.M., Carter, V. Golet, F.C. and La Roe, E.T. 1979. Classification of wetlands and deepwater habitats of the United States. U.S. Fish and Wildlife Service FWS/OBS-79/31. 103 pp.

Cowley, P.D. 1998. Fish population dynamics in a temporarily open/closed South African estuary. PhD Thesis, Rhodes University, Grahamstown. 145 pp.

Cowley, P.D. Wood, A.D., Corroyer B., Nsubuga Y. and Chalmers R. 2000. A survey of fishery resource utilization on four Eastern Cape Estuaries (Great Fish, West Kleinemonde, East Kleinemonde and Kowie). South African Institute for Aquatic Biodiversity. Grahamstown.

Cowley, P.D. and Whitfield, A.K. 2001. Ichthyofaunal characteristics of a typical temporarily open/closed estuary on the south east coast of South Africa. Ichthyological Bulletin 71: 1-17.

Cowley, P.D. and Daniel, C. 2001. Estuaries of the Ndlambe municipality (EC 105). Great Fish, Klein Palmiet (Brak), East Kleinemonde, West Kleinemonde, Riet, Rufanes, Kowie, Kasouga, Kariega, Bushmans and Boknes estuaries. Special report prepared for the Institute of Natural Resources. 63 pp.

Cowley, P.D., Wood, A.D., Corroyer, B., Nsubuga, Y. and Chalmers, R. 2003. A survey of fishery resource utilization on four Eastern Cape estuaries (Great Fish, West Kleinemonde, East Kleinemonde and Kowie). Protocols contributing to the management of estuaries in South Africa, with a particular emphasis on the Eastern Cape Province Volume III Project C, Supplementary Report C5. Grahamstown. 165 pp.

Cross, D.H. and Fleming, K.L. 1989. Control of Phragmites or common reed. Fish and Wildlife Leaflet 13.4.12. Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. 5 pp.

Crossle, K. and Brock, M. 2002. How do water regime and clipping influence wetland plant establishment from seed banks and subsequent reproduction? Aquatic Botany 74: 43–56.

Crum, G.H. and Bachmann, R.W. 1973. Submerged aquatic macrophytes of the Iowa Great Lakes region, Iowa State. Journal of Research 48: 147--173.

CSIR. 1992. Great Brak management programme: Report on the monitoring results for the period April 1990 to March 1992. CSIR Report EMAS-C 92083, Interim Report. Stellenbosch.

CSIR. 2003. Great Brak Estuary Management Programme: Review Report. CSIR Report ENV-S-C-2003-092. Stellenbosch.

169 Curcó, A., Ibanez, C., Day, J.W. and Prat, N. 2002. Net primary production and decomposition of salt marshes of the Ebre Delta (Catalonia, Spain). Estuaries 25: 309–324.

Daehler, C.C. and Strong, D.R. 1994. Variable reproductive output among clones of Spartina alterniflora (Poaceae) invading San Francisco Bay, California: the influence of herbivory, pollination, and establishment site. American Journal of Botany 81: 307–313.

Davis, B.A.S. and Stevenson, A.C. 2007. The 8.2 ka event and Early–Mid Holocene forests, fires and flooding in the Central Ebro Desert, NE Spain. Quaternary Science Reviews 26: 1695–1712.

Davison, D.M. and Hughes, D.J. 1998. Zostera Biotopes volume I. An Overview of Dynamics and Sensitivity Characteristics for Conservation Management of Marine SACs. Scottish Association for Marine Science UK Marine SACs Project. 95 pp.

Davison, D.M. and Hughes, D.J. 1998. Zostera Biotopes (Volume I). An overview of dynamics and sensitivity characteristics for conservation management of marine SACs. Scottish Association for Marine Science (UK Marine SACs Project).

Davy, A. J., Noble, S.M. and Oliver, R.P. 1990. Genetic variation and adaptation to flooding in plants. Aquatic Botany 38: 91–108.

Davy, A.J. 2000. Development and structure of salt marshes: community patterns in time and space. In: (M.P. Weinstein and D.A. Kreeger, eds.). Concepts and controversies in tidal marsh ecology. Kluwer Academic Publishers. Dordrecht, the Netherlands. 137-156.

Davy, A.J., Bishop G.F. and Costa, C.S.B. 2001. Salicornia L. (Salicornia pusilla J. Woods, S. ramosissima J. Woods, S. europaea L., S. obscura P.W. Ball and Tutin, S. nitens P.W. Ball and Tutin, S. fragilis P.W. Ball and Tutin and S. dolichostachya Moss). Journal of Ecology 89(4): 681-707.

Davy, A.J., Bishop, G.F., Mossman, H., Redondo-Gómez, S., Castillo, J.M., Castellanos, E.M., Luque, T. and Figueroa, M.E. 2006. Biological Flora of the British Isles: Sarcocornia perennis (Miller) A.J. Scott. Journal of Ecology 94: 1035–1048.

Dawson, F.H. 1988. Water flow and the vegetation of running waters. In: Symoens, J.J. (Ed.), Vegetation of Inland Waters. Handbook of Vegetation Science, vol. 15/1. Kluwer Academic Publishers, Dordrecht.

Day, J.H. 1981. Estuarine Ecology with Particular Reference to Southern Africa. A.A. Balkema, Cape Town. 411 pp. de Bakker, N.V.J., Beem, A.P., van de Staaij, J.W.M., Rozema, J. and Aerts, R. 2001. Effects of UV-B radiation on a charophycean alga, Chara aspera. Plant Ecology 154: 237–246.

Deegan, B. and Harrington, T.J. 2004. The Distribution and Ecology of Schoenoplectus triqueter in the Shannon Estuary. Biology and Environment: Proceedings of the Royal Irish Academy 104B(2): 107-117.

170 Deegan, B., Harrington, T.J. and Dundon, P. 2005. Effects of salinity and inundation regime on growth and distribution of Schoenoplectus triqueter. Aquatic Botany 81: 199–211.

Deegan, B.M., White, S.D. and Ganf, G.G. 2007. Nutrients and water level fluctuations: A study of three aquatic plants. River Research and Applications 21(8): 899–908. de Laune, R.D., Smith, C.J. and Patrick, W.H.Jr. 1983. Relationship of marsh elevation, redox potential and sulfide to Spartina alternifolia productivity. Soil Science Society of America 47: 930-935. de Laune R.D., Pezeshki, S.R. and Patrick W.H. Jr. 1987. Response of coastal plants to increase in submergence and salinity. Journal of Coastal Research 3(4): 535-546.

Denslow, J.S. and Battaglia, L.L. 2002. Stand competition and structure across a changing hydrological gradient: Jean Lafitte National Park, Louisiana, USA. Wetlands 22: 738-752.

Department of Nature and Conservation (DNC). 2006. Managing Weeds in Bushland. Department of Nature and Conservation (DNC). Urban Nature Programme. Bentley, Australia. 1-2.

Desclaux, D. and Roumet, P. 1996. Impact of drought stress on the phenology of two soybean cultivars. Field Crops Research 46: 61–70. de Winton, M.D., Clayton, J.S. and Champion, P.D. 2000. Seedling emergence from seed banks of 15 New Zealand lakes with contrasting vegetation histories. Aquatic Botany 66: 181–194. de Winton, M.D., Casanova, M.T. and Clayton, J.S. 2004. Charophyte germination and establishment under low irradiance. Aquatic Botany 79: 175–187.

Dietert, M.F. and Shontz, J.P. 1978. Germination ecology of a Maryland population of salt marsh Bulrush (Scirpus robustus). Estuaries 1(3): 164-170.

Diggory, Z.E. and Parker, V.T. 2010. Seed supply and revegetation dynamics at restored tidal marshes, Napa River, California. Restoration Ecology 10: 1526-2009.

Doerr, A.H. 1990. Fundamentals of Physical Geography. Wm.C. Brown Publishers, Alaska. pp 281-302.

Drabsch, J.M., Parnell, K.E., Hume, T.M. and Dolphin, T. J. 1999. The capillary fringe and the water table in an intertidal estuarine sand flat. Estuarine, Coastal and Shelf Science 48: 215-222.

Dreyer, L.L., Esler, K.J. and Zietsman, J. 2006. Flowering phenology of South African Oxalis: Possible indicator of climate change? South African Journal of Botany 72: 150 – 156.

Dykyjová, D. 1986. Production ecology of Bolboschoenus maritimus (L.) Palla (Scirpus maritimus L. s.l.). Folia Geobotanica and Phytotaxonomica 21(1) 27-64.

171 Duarte, C.M. and Kalff, J. 1987. Weight-density relationships in submerged macrophytes. The importance of light and plant geometry. Oecologia 72: 612-617.

Duarte, C.M. and Kalff, J. 1990. Patterns in the submerged macrophyte biomass of lakes and the importance of the scale of analysis in the interpretation. Canadian Journal of Fisheries and Aquatic Science 47: 357–363.

Duarte, C.M. 1991. Seagrass depth limits. Aquatic Botany 40: 363-377.

Dunton K.H., Hardegree, B. and Whitledge, T.E. 2001. Response of estuarine marsh vegetation to interannual variations in precipitation. Estuaries 24(6A): 851–861

Dugdale, T.M., de Winton, M.D. and Clayton, J.S. 2001. Burial limits to the emergence of aquatic plant propagules. New Zealand Journal of Marine and Freshwater Research 35: 147-153.

Duke, J.A. 1978. The quest for tolerant germplasm. In: ASA Special Symposium 32, Crop tolerance to suboptimal land conditions. American Society of Agronomics. Madison, WI. pp. 1–61.

Duke, J.A. 1979. Ecosystematic data on economic plants. Quarterly Journal of Crude Drug Research 17(3–4): 91–110.

Dutta, A.C. 1979. Botany for degree students (5th Edition).Oxford University Press. Calcutta.

Duvauchelle, D. and Magee, P. 2007. Plant Fact Sheet: Sporobolus virginicus (L.) Kunth. U.S. Department of Agriculture.

DWAF. 2005. Kromme/Seekoei Catchments Reserve Determination Study: Technical Component. Kromme. Prepared by CSIR for CES. Report No. KSCR-IR-0001.

Dwire, K.A., Boon Kauffman, J. and Baham, J.E. 2006. Plant species distribution in relation to water-table depth and soil redox potential in montane riparian meadows. Wetlands 26(1):131–146.

Egan, T.P. and Ungar, I.A. 1999. The effects of temperature and seasonal change on the germination of two salt marsh species, Atriplex prostrata and Salicornia europaea, along a salinity gradient. International Journal of Plant Sciences 160(5): 861-867.

Engloner, A.I. 2009. Structure, growth dynamics and biomass of reed (Phragmites australis) – A review. Flora 204: 331–346.

Ehrenfeld, J.G., Han, X., Parsons, W.F.J. and Zhu, W. 1997. On the nature of environmental gradients: Temporal and spatial variability of soils and vegetation in the New Jersey Pinelands. Journal of Ecology 85(6): 785-798.

Eid, E.M., Shaltout, K.H., Al-Sodany, Y.M. and Jensen, K. 2010. Effects of abiotic conditions on Phragmites australis along geographic gradients in Lake Burullus, Egypt. Aquatic Botany 92: 86–92.

172 Ekstam, B. and Forseby, A. 1999. Germination response of Phragmites australis and Typha latifolia to diurnal fluctuations in temperature. Seed Science Research 9: 157-163.

Eleuterius, L.N. and Caldwell, J.D. 1984. Reproductive phenology of tidal marsh plants in Mississippi. Castanea 49(4): 172-179.

Epling, C., Lewis, H. and Ball, F.M. 1960. The breeding group and seed storage: A study in population dynamics. Evolution 14: 238-55.

Ervin, G.N. and Wetzel, R.G. 2001. Seed fall and field germination of needlerush, Juncus effusus L. Aquatic Botany 71: 233–237.

Espinar, J.L., García, L.V., García Murillo, P. and Toja, J. 2002. Submerged macrophyte zonation in a Mediterranean salt marsh: A facilitation effect from established helophytes? Journal of Vegetation Science 13: 831-840.

Espinar, J.L., García, L.V., Figuerola, J., Green, A.J. and Clemente, L. 2004. Helophyte germination in Mediterranean wetlands: Gut-passage by ducks changes seed response to salinity. Journal of Vegetation Science 15: 313–320.

Espinar, J.L., Luis, V., García, L.V. and Clemente, L. 2005. Seed storage conditions change the germination pattern of clonal growth plants in Mediterranean salt marshes. American Journal of Botany 92(7): 1094– 1101.

Evans, C.E. and Etherington, J.R. 1990. The effect of soil water potential on seed germination of some British plants. New Phytologist 115(3): 539-548.

Fang, X., Subudhi, P.K., Venuto, B.C., Harrison, S.A. and Ryan, A.B. 2004. Influence of flowering phenology on seed production in smooth cordgrass (Spartina alterniflora Loisel.). Aquatic Botany 80: 139–151.

Fenner, M. 1985. Seed Ecology. Chapman and Hall, London. 214 pp.

Fenner, M. and Thompson, K. 2005. The ecology of seeds. Press syndicate of the University of Cambridge. Cambridge. 250 pp.

Fernandez-Alaez, M., Fernandez-Alaez, C. and Rodriguez, S. 2002. Seasonal changes in biomass of charophytes in shallow lakes in the northwest of Spain. Aquatic Botany 72: 335–348.

Fernández-Illescas, F., Javier, F., Nieva J., Silva, I., Tormo, R. and Muñoz, A.F. 2010. Pollen production of Chenopodiaceae species at habitat and landscape scale in Mediterranean salt marshes: An ecological and phenological study. Review of Palaeobotany and Palynology 161: 127–136.

Flindt, M.R. 1992. Measurements of nutrient fluxes and mass balances by on-line in situ dialysis in a Zostera marina L. bed culture. Verhandlungen des Internationalen Verein Limnologie (Limnology) 25: 2259–2264.

173 Flindt, M.R., Pardal, M.A., Lillebø, A.I., Martins, I. and Marques, J.C. 1999. Nutrient cycling and plant dynamics in estuaries: A brief review. Acta Oecologica 20 (4): 237−248.

Figueroa, M.E., Castillo, J.M., Redondo, S., Luque, T., Castellanos, E.M., Nieva, F.J., Luque, C.J., Rubio- Casal, A.E. and Davy, A.J. 2003. Facilitated invasion by hybridization of Sarcocornia species in a salt- marsh succession. Journal of Ecology 91(4) 616-626.

Forbes, V.R. 1998. Recreational and resource utilisation of Eastern Cape estuaries and development towards a management strategy. MSc Thesis. University of Port Elizabeth, Port Elizabeth.

Fonseca, M.S., Fisher, J.S., Zieman, J.C. and Thayer, G.W. 1982. Influence of the seagrass Zostera marina L., on current flow. Estuarine, Coastal and Shelf Science 15: 351-364.

Forsberg, C. 1965. Sterile germination of Chara and seeds of Najas marina. Physiologia Plantarium. 18: 129– 137.

Freitag, H., Golub, V.B. and Yuritsyna, N.A. 2001. Halophytic plant communities in the northern Caspian lowlands: Annual halophytic communities. Phytocoenologia 31: 63–108.

Froend, R.H. and McComb, A.J. 1994. Emergent macrophyte distribution, productivity and reproduction phenology relative to water regime at wetlands of south west Australia. Australian Journal of Marine and Freshwater Research. 45: 1491-1508.

Geissler, K. and Gzik, A. 2008. The impact of flooding and drought on seeds of Cnidium dubium, Gratiola officinalis, and Juncus atratus, three endangered perennial river corridor plants of Central European lowlands. Aquatic Botany 89: 283–291.

Gervais, C., Trahan, R., Moreno, D. and Drolet, A.M. 1993. Phragmites australis in Quebec: geographic distribution, chromosome numbers and reproduction. Canadian Journal of Botany 71: 1386–1393.

Gesti, J., Badosa, A. and Quintana X.D. 2005. Reproductive potential in Ruppia cirrhosa (Petagna) Grande in response to water permanence. Aquatic Botany 81: 191–198.

Ghosh, S.K., Santra, S.C. and Mukherjee, P.K. 1993. Phenological studies in aquatic macrophytic plants of lower Gangetic Delta, West Bengal, India. Feddes Repertorium 104(1-2): 93–111.

Gomez-Plaza, M., Martinez-Mena, M., Albaladejo, J. and V.M. Castillo. 2001. Factors regulating spatial distribution of soil water content in small semiarid catchments. Journal of Hydrology 253: 211-226.

González-Alcaraz, M.N., Conesa, H.M., Tercero, M.C., Schulin, R., lvarez-Rogel, J.A. and Egea, C. 2010 (In press). The combined use of liming and Sarcocornia fruticosa development for phytomanagement of salt marsh soils polluted by mine wastes. Journal of Hazardous Materials. 32 pp.

Goodman, A.M., Ganfa, G.G., Maierb, H.R. and Dandy, G.C. 2011. The effect of inundation and salinity on the germination of seed banks from wetlands in South Australia. Aquatic Botany 94(2): 102-106.

174 Gordon-Gray, K.D., Baijnath, H., Ward, C.J. and Wragg, P.D. 2009. Studies in Cyperaceae in southern Africa 42: Pseudo-vivipary in South African Cyperaceae. South African Journal of Botany 75: 165–171.

Grace, J.B. and Wetzel, R.G. 1982. Variations in growth and reproduction within populations of two rhizomatous plant species: Typha latifolia and Typha angustifolia. Oecologia 53: 258—263.

Grace, J.B. 1987. The effects of pre-emption on the zonation on two Typha species along lakeshores. Ecological Monographs 57: 283-303.

Grace, J.B. 1988. The effects of nutrient additions on mixtures of Typha latifolia L. and Typha domingensis Pers. along a water-depth gradient. Aquatic Botany 31: 83–92.

Grace, J.B. 1989. Effects of water depth on Typha latifolia and Typha domingensis. American Journal of Botany 76(5): 762-768.

Gray, A.J. 1985. Adaptation in perennial coastal plants with particular reference to heritable variation in Puccinellia maritima and Ammophila arenaria. Vegetatio 61: 179-188.

Greenway, H. and Munns, R. 1980. Mechanisms of salt tolerance in non halophytes. Annual Review of Plant Physiology 31: 149-190.

Greenwood, M.E. and du Bowy, P.J. 2005. Germination characteristics of Zannichellia palustris from New South Wales, Australia. Aquatic Botany 82: 1–11.

Greenwood, M.E. 2008. Predicting the effects of salinity on three dominant macrophytes: An anticipatory approach to the restoration of degraded coastal wetlands in NSW, Australia. PhD Thesis. The University of Newcastle, NSW, Australia. 219 pp.

Greenwood, M.E. and MacFarlane, G.R. 2009. Effects of salinity on competitive interactions between two Juncus species. Aquatic Botany 90: 23-29.

Gribsholt, B. and Kristensen, E. 2003. Benthic metabolism and sulfur cycling along an inundation gradient in a tidal Spartina anglica salt marsh. Limnology and Oceanography 48(6): 2151–2162.

Gries, C., Kappen, L. and Losch, R. 1990. Mechanism of flood tolerance in reed, Phragmites australis (Cav.) Trin. ex Steudel. The New Phytologist 114: 589-593.

Griffiths, S.P. 2001. Factors influencing fish composition in an Australian intermittently open estuary. Is stability salinity dependent? Estuarine, Coastal and Shelf Science 52: 739–751.

Grillas, P. 1990. Distribution of submerged macrophytes in the Camargue in relation to environmental factors. Journal of Vegetation Science 1: 393-402.

Grime, J.P. 1979. Plant strategies and vegetation processes. John Wiley and Sons. New York. 222 pp.

175 Grime, J.P., Mason, G. and Curtis, A. 1981. A comparative study of germination characteristics in a local flora. Journal of Ecology 69: 1017-1059.

Grise, D., Titus, J.E. and Wagner, D.J. 1986. Environmental pH influences the growth and tissue chemistry of the submerged macrophyte Vallisneria americana. Canadian Journal of Botany 64: 306-310.

Groenendijk, A.M. 1985. Ecological consequences of tidal management for the salt-marsh vegetation. Vegetatio 62: 415-424.

Gucker, C.L. 2008. Phragmites australis. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer).

Gul, B. and Weber, D.J. 1999. Effect of salinity, light, and thermoperiod on the seed germination of Allenrolfea occidentalis. Canadian Journal of Botany 77: 1-7.

Gulzar, S. and Khan, M.A. 2001. Seed germination of a halophytic grass Aeluropus lagopoides. Annals of Botany 87: 319-324.

Gumenge, P. 2010. Eastern Cape reels as drought persists. Crocotts Mail. Grahamstown.

Güsewell, S. and Klötzli, F. 2000. Assessment of aquatic and terrestrial reed (Phragmites australis) stands. Wetlands Ecology and Management 8: 367–373.

Gustafsson, A. and Simak, M. 1963. X-ray photography and seed sterility in Phragmites communis Trin. Hereditas 49: 442–450.

Haramis, G.M. and Carter, V. 1983. Distribution of submersed aquatic macrophytes in the tidal Potomac River. Aquatic Botany 15: 65-79.

Harwell, M.C. and Havens, K.E. 2003. Experimental studies on the recovery potential ofsubmerged aquatic vegetation after flooding and desiccation in a large subtropical lake. Aquatic Botany 77: 135–171.

Harley, C.D.G. and Bertness, M.D. 1996. Structural interdependence: An ecological consequence of morphological responses to crowding in marsh plants. Functional Ecology 10(5): 654-661.

Harris, S.W. and Marshall, W.H. 1960. Germination and planting experiments on soft-stem and hard-stem bulrush. Journal of Wildlife Management 24: 134-139.

Harrison, T.D. and Whitfield, A.K. 1995. Fish community structure in three temporarily open/closed estuaries on the Natal coast. Ichthyological Bulletin of the JLB Smith Institute of Ichthyology 64: 1-80.

Harrison, T.D. 2003. Biogeography and community structure of fishes in South African estuaries. PhD Thesis. Rhodes University, Grahamstown. 229 pp.

176 Harrison, T.D. 2004. Physico-chemical characteristics of South African estuaries in relation to the zoogeography of the region. Estuarine, Coastal and Shelf Science. 61: 73-87.

Haslam, S.M. 1971. The performance of Phragmites communis in relation to water supply. Annals of Botany 34: 867-877.

Haslam, S.M. 1972. Biological Flora of the British Isles, Phragmites communis Trin. Journal of Ecology 60: 585–610.

Hastie, B.F. and Smith, S.D.A. 2006. Benthic macrofaunal communities in intermittent estuaries during a drought: Comparisons with permanently open estuaries. Journal of Experimental Marine Biology and Ecology 330: 356–367.

Havens, K.E., Sharfstein, B., Brady, M.A., East, T. L., Harwell, M.C., Maki, R.P. and Rodusky, A.J. 2004. Recovery of submerged plants from high water stress in a large subtropical lake in Florida, USA. Aquatic Botany 78: 67–82.

Hellings, S.E. and Gallagher, J.L. 1992. The effects of salinity and flooding on Phragmites australis. Journal of Applied Ecology 29: 41–49.

Henninger, T.O., Froneman, P.W. and Hodgson, A.N. 2008. The population dynamics of the estuarine isopod Exosphaeroma hylocoetes within three temporarily open/closed southern African estuaries. African Zoology 43(2): 202-217.

Heydorn, A.E.F. and Tinley, K.L. 1980. Estuaries of the Cape. Part 1. Synopsis of the Cape coast natural features, dynamics and utilization. CSIR Research Report 380: 1-97.

Heydorn, A.E.F. and Grindley, J.A. 1982. Estuaries of the Cape. Part II: Synopsis of available information on individual systems. Report no. 10: Kowie (CSE 10). CSIR Research Report 409. National Research Institute for Oceanography. Council for Scientific and Industrial Research. Stellenbosch.

Hillel, D. 1971. Soil and water: Physical principles and processes. In: Physiological Ecology: A series of monographs, texts and treatises. Kozlowski, T.T. (eds). Academic Press Inc., New York. 288 pp.

Hill, B.J. and Allanson, B.R. 1971. Temperature tolerance of the estuarine prawn Upogebia africana (Anomura, Crustacea). Marine Biology 11: 337-343.

Holley, D. 2009. Soil nutrients and pH affect plant growth. The availability of minerals and soil pH influence growth in plants. pp 1.

Hogeland, A.M. and Killingbeck, K.T. 1985. Biomass, productivity and life history traits of Juncus militaris Bigel. in two Rhode Island (U.S.A.) freshwater wetlands. Aquatic Botany 22: 335-346.

Hootsmans, M.J.M. and Wiegman, F. 1998. Four helophyte species growing under salt stress: Their salt of life? Aquatic Botany 62: 81-94.

177 Howard-Williams, C. and Allanson, B.R. 1979. The ecology of Swartvlei: Research for planning and future management. Water Research Commission Report. 26 pp.

Howard-Williams, C. 1980. Aquatic macrophytes communities of the Wilderness Lakes: Community structure and associated environmental conditions. Journal of the Limnological Society of southern Africa 6(2): 85- 92.

Howard-Williams, C. and Liptrot, M.R.M. 1980. Submerged macrophyte communities in a brackish South African estuarine-lake system. Aquatic Botany 9:101-116.

Howard-Williams, C. and Allanson, B.R. 1981. Phosphorous cycling in a dense Potamogeton pectinatus L. bed. Oecologia 49: 56-66.

Hroudová, Z., Zákravský, P. and Frantik, T. 1999. Ecological differentiation of Central European Bolboschoenus taxa and their relationship to plant communities. Folia Geobotanica 34: 77-96.

Hroudová, Z., Zákravský, P., Duchásek, M. and Marhold, K. 2007. , distribution and ecology of Bolboschoenus in Europe. Annals of Botany Fennici 44: 81-122.

Hu, B., Xie, S., Feng, J. and Zhang, M. 2008. Study on the phenology of Chara vulgaris in Xin‘an Spring, north China. Frontiers of Biology in China 4(2) 207-213.

Huiskes A.H.L., Koutstaal B.P., Herman P.M.J., Beeftink W.G., Markusse M.M. and de Munck W. 1995. Seed dispersal of halophytes in tidal salt marshes. Journal of Ecology 83: 559-567.

Human, L. 2010. Submerged macrophytes, reeds and macroalgae as indicators of nutrient enrichment in the East Kleinemonde Estuary. MSc Thesis. Nelson Mandela Metropolitan University. Port Elizabeth. 109 pp.

Hutchinson, G.E. 1975. A treatise on limnology. Version 3. Limnological Botany. John Wiley and Sons, New York. 600 pp.

Invers, O., Romero, J. and Perez, M. 1997. Effects of pH on seagrass photosynthesis: A laboratory and field assessment. Aquatic Botany 59: 185-194.

Ishii, J. and Kadono, Y. 2002. Factors influencing seed production of Phragmites australis. Aquatic Botany 72: 129–141.

Jackson, M.D. and Drew, M.C. 1984. Effects of flooding on herbaceous plants. In: Flooding and Plant Growth (ed. Kozlowski, T.T.). Academic Press Inc. New York, London. 47-128 pp.

James, R.T., Jones, B.L. and Smith, V.H. 1995. Historical trends in the Lake Okeechobee ecosystem. II. Nutrient budgets. Archiv für Hydrobiologia 107: 25–47.

Jarvis, J.C. and Moore, K.A. 2008. Influence of environmental factors on Vallisneria americana seed germination. Aquatic Botany 88: 283–294

178 Javier, F., Nieva, J., Castellanos, E.M., Castillo, J.M. and Figueroa. M.E. 2005. Clonal growth and tiller demography of the invader cordgrass Spartina densiflora Brongn. at two contrasting habitats in SW European salt marshes. Wetlands 25(1): 122–129.

Jefferies, R.L. and Gottilieb, L.D. 1982. Genetic differentiation of the microspecies Salicornia europaea L. (sensu stricto) and S. ramosissima J. Woods. New Phytologist 92:123–129.

Jefferies, R.L., Jensen, A. and Bazely, D. 1983. The biology of the annual Salicornia europaea agg. at the limits of its range in Hudson Bay. Canadian Journal of Botany 61: 762–773.

Jefferies, R.L. and Rudmik, T. 1991. Growth, reproduction and resource allocation in halophytes. Aquatic Botany 39: 3-16.

Jennings, M.E. 2006. Nutrient dynamics in and offshore of two permanently open South African estuaries with contrasting fresh water inflow. MSc Thesis. Rhodes University. Grahamstown. 155 pp.

Ji, Y., Zhou, G. and New, T. 2009. Abiotic factors influencing the distribution of vegetation in coastal estuary of the Liaohe Delta, Northeast China. Estuaries and Coasts 32(5): 937-942.

Jones, V. and Richards, P.W. 1954. Juncus acutus L. Journal of Ecology 42: 639-650.

Jordan, T.E., Correll, D.L. and Whigham, D.F. 1983. Nutrient flux in the Rhode River: Tidal exchange of nutrients by brackish marshes. Estuarine Coastal and Shelf Science 17: 651-667.

Jupp, B.P. and Spence, D.H.N. 1977. Limitations of macrophytes in a eutrophic lake, Loch Leven. I. Effects of phytoplankton. Journal of Ecology 65: 175-186.

Jury, M.R. and Levy, K. 1993. The Eastern Cape drought. Water SA 19(2): 133-137.

Kadereit, G., Mucina, L. and Freitag, H. 2006. Phylogeny of Salicornioideae (Chenopodiaceae): Diversification, biogeography, and evolutionary trend. Taxon 55(3): 617-642.

Kadereit, G., Ball, P., Beer, S., Mucina, L., Sokoloff. D., Teege, P., Yaprak, A.E. and Freitag, H. 2007. A taxonomic nightmare comes true: phylogeny and biogeography of glassworts (Salicornia L., Chenopodiaceae). Taxon 56 (4): 1143–1170.

Kahn, M.A. and Ungar, I.A. 1984. The effect of salinity and temperature on the germination of polymorphic seeds and growth of Atriplex triangularis Willd. American Journal of Botany 71(4): 481-489.

Kahn, M.A., Gul, B., Darrell J. and Weber, D.J. 2000. Germination responses of Salicornia rubra to temperature and salinity. Journal of Arid Environments 45: 207-214.

Kahn, M.A. and Gul, B. 2006. Halophyte seed germination. In: Kahn, M.A., Weber, D.J. (Eds.), Ecophysiology of High Salinity Tolerant Plants. Springer, The Netherlands. 11–30 pp.

179 Kaligaric, M., Bohanec, B., Simonovik, B. and N. Sajna. 2008. Genetic and morphologic variability of annual glassworts (Salicornia L.) from the Gulf of Trieste (Northern Adriatic). Aquatic Botany 89: 275–282.

Kalin, M. and Smith, M.P. 2007. Germination of Chara vulgaris and Nitella flexilis oospores: What are the relevant factors triggering germination? Aquatic Botany 87: 235–241.

Kantrud, Harold A. 1991. Wigeongrass (Ruppia maritima L.): A literature Review. U.S. Fish and Wildlife Service, Fish and Wildlife Research 10: 58 pp.

Kantrud, H.A. 1996. The alkali (Scirpus maritimus L.) and salt marsh (S. robustus Pursh) bulrushes: A literature review. National Biological Service, Information and Technology Report 6. Jamestown, ND: Northern Prairie Wildlife Research Center Online. http://www.npwrc.usgs.gov/resource/plants/bulrush/index.htm (Version 16JUL97).

Kassas, M. 2002. Management Plan for Burullus Protectorate Area. MedWetCoast,Global Environmental Facility (GEF) and Egyptian Environmental Affairs Agency (EEAA), Cairo.

Katembe, W.J., Ungar, I.A. and Mitchell, J.P. 1998. Effect of salinity on germination and seedling growth of two Atriplex species (Chenopodiaceae). Annals of Botany 82: 167 – 175.

Kautsky, L. 1988. Life strategies of aquatic soft bottom macrophytes. Oikos 53: 126-135.

Kautsky, L. 1990. Seed and tuber banks of aquatic macrophytes in the Asko area, northern Baltic proper. Holarctic Ecology 13: 143–148.

Keddy, P.A. and Reznicek, A.A. 1982. The role of seed banks in the persistence of Ontario‘s coastal plain flora. American Journal of Botany 69: 13–22.

Keller, B., Lajtha, K. and Christofor, S. 1998. Trace metal concentrations in the sediments and plants of the Danube delta, Romania. Wetlands 18: 42-50.

Kemp, W.M. and Murray, L. 1986. Oxygen release from roots of the submersed macrophyte Potamogeton perfoliatus L.: Regulating factors and ecological implications. Aquatic Botany 26: 271-283.

Kennish, M.J. 2002. Environmental threats and environmental future of estuaries. Environmental Conservation 29: 78-107.

Kercher, S.M. and Zedler, J.B. 2004. Flood tolerance in wetland angiosperms: a comparison of invasive and noninvasive species. Aquatic Botany 80: 89–102.

Kettenring, K.M. and Whigham, D.F. 2009. Seed viability and seed dormancy of non-native Phragmites australis in suburbanized and forested watersheds of the Chesapeake Bay, USA. Aquatic Botany 91: 199– 204.

180 Kettenring, K.M., McCormick, M.K., Baron, H.M. and Whigham, D.F. 2010. Phragmites australis (common reed) invasion in the Rhode River sub-estuary of the Chesapeake Bay: Disentangling the effects of foliar nutrients, genetic diversity, patch size, and seed viability. Estuaries Coasts 33: 118–126.

Khan, M.A. and Gul, B. 1998. High salt tolerance in germinating dimorphic seeds of Arthrocnemum indicum. International Journal of Plant Sciences 159(5): 826-832.

Khan, M.A. and Ungar, I.A .1984a. The effect of salinity and temperature on germination of polymorphic seeds and growth of Atriplex triangularis. American Journal of Botany 71: 481 – 489.

Khan, M.A. and Ungar, I.A .1984b. Seed polymorphism and germination responses to salinity stress in Atriplex triangularis Willd. Botanical Gazette 145(4): 487-494.

Khan, M.A. and Ungar, I.A. 1985. The role of hormones in regulating the germination of polymorphic seeds and early seedling growth of Atriplex triangularis Willd. under saline conditions. Physiologia Plantarium 63:109– 113.

Khan, M.A., Gul, B. and Webber, D.J. 2000a. Germination responses of Salicornia rubra to temperature and salinity. Journal of Arid Environments 45: 207-214.

Khan, M.A., Gul, B. and Webber, D.J. 2000b. The effect of salinity on the growth, water status, and ion content of a leaf succulent perennial halophyte, Suaeda fruticosa (L.) Forssk. Journal of Arid Environments 45: 73- 84.

Khan, M.A. and Gulzar, S. 2003. Germination responses of Sporobolus ioclados: A potential forage grass. Journal of Arid Environments 53: 387-394.

Kiorboe, T. 1980. Distribution and production of submerged macrophytes in Tipper Grnnd (Ringkobing fjord, Denmark) and the impact of waterfowl grazing. Journal of Applied Ecology 17: 675-687.

Kludze, H.K. and de Laune, R.D. 1994. Methane emission and growth of Spartina patens in response to soil redox intensity. Soil Science Society Am. J. 58: 1838–1845.

Kludze, H.K. and de Laune, R.D. 1995. Straw application effects on methane and oxygen exchange and growth in rice. Soil Science Society of America Journal. 59: 824–830.

Knevel, I.C. and Lubke, R.A. Reproductive phenology of Scaevola plumieri; A key coloniser of the coastal foredunes of South Africa. Plant Ecology 175 (1): 137 – 145.

Koch, M.S. and Mendelssohn, I.A. 1989. Sulphide as a soil phytotoxin: differential responses in two marsh species. Journal of Ecology 77: 565-578.

Koch, M.S., Mendelssohn, I.A. and McKee, K.L. 1990. Mechanism for the hydrogen sulfide-induced growth limitation in wetland macrophytes. Limnology and Oceanography 35: 399-408.

181 Kongchum, M. 2005. Effect of plant residue and water management practices on soil redox chemistry, methane emission, and rice productivity. PhD. Louisiana State University and Agricultural and Mechanical College. 189 pp.

Kopke, D. 1988 The climate of the Eastern Cape. In Towards an environmental plan for the Eastern Cape (Bruton, M. N. & Gess, F. W. eds). Rhodes University, Grahamstown. 44-52 pp.

Kruger, L. and Kirst, G.O. 1991. Field studies on the ecology of Bolboschoenus maritimus (L.) Palla. (Scirpus maritimus L.s.l.). Folia Geobotanica Et Phytotaxonomica 26: 277-286.

Kufel, L. and Kufel, I. 2002. Chara beds acting as nutrient sinks in shallow lakes - a review. Aquatic Botany 72: 249–260.

Kuhn, N.L. and Zedler, J.B. 1997. Differential effects of salinity and soil saturation on native and exotic plants of a coastal salt marsh. Estuaries 20: 391–403.

Kyuma, K. 2004. Paddy Soil Sciences. Kyoto University Press and Transpacific Press, Kyoto and Melbourne. 37-59 pp.

Laegdsgaad, P. 2006. Ecology, disturbance and restoration of coastal saltmarsh in Australia: a review. Wetlands Ecology and Management: 14:379-399.

Lammens, E.H.R.R. 1989. Causes and consequences of the success of bream in Dutch eutrophic lakes. Hydrobiologia 23: 11–18.

Larcher, W. 1995. Physiological plant ecology. Springerling Verlag, Berlin. 493 pp.

Leithead, H.L., Yarlett, L.L. and Shiflett, T.N. 1976. 100 native forage grasses in 11 southern states. USDA SCS Agriculture Handbook No. 389, Washington, DC.

Lenssen, J.P.M., Ten Dolle, G.E. and Blom, C.W.P.M. 1998. Flooding and the recruitment of reed marsh and tall forb plant species. Plant Ecology 139: 13–23.

Lenssen, J.P.M., Van Kleunen, M., Fischer, M. and De Kroon, H. 2004. Local adaptation of the clonal plant Ranunculus reptans to flooding along a small-scale gradient. Journal of Ecology 92: 696–706.

Lentz, K.A. and Dunson, W.A. 1998. Water level affects growth of endangered northeastern bulrush, Scirpus ancistrochaetus Schuyler. Aquatic Botany 60: 213–219.

Leps, J. and Smilauer, P. 1999. Multivariate analysis of ecological data. Faculty of Biological Sciences, University of South Bohemia, Ceské Budejovice.110 pp.

Li, J., Huang, P. and Zhang, R. 2010. Modelling the refuge effect of submerged macrophytes in ecological dynamics of shallow lakes: A new model of fish functional response. Ecological Modelling 221: 2076–2085

182 Lieffers, V.J. and Shay, J.M., 1981. The effects of water level on growth and reproduction of Scirpus maritimus var. paludus. Canadian Journal of Botany 59: 118–121.

Lieffers, V.J. and Shay, J.M. 1982. Distribution and variation in growth of Scirpus maritimus var. paludosus on the Canadian prairies. Canadian Journal of Botany 60(10): 1938–1949.

Lillebø, A.I., Pardal, M.A., Neto, J.M. and Marques, J.C. 2003. Salinity as the major factor affecting Scirpus maritimus annual dynamics - Evidence from field data and greenhouse experiment. Aquatic Botany 77: 111–120.

Lissner, J. and Schierup, H.H. 1997. Effects of salinity on the growth of Phragmites australis. Aquatic Botany 55: 247-260.

Lickacz, J. and Penny, D. 2001. Soil organic matter. Government of Alberta. Agriculture and Rural Development. Plant Industry Division. Alberta.

Lijklema, L. 1994. Nutrient dynamics in shallow lakes: effects of changes in loading and role of sediment–water interactions. Hydrobiologia 275/276: 335–348.

Lillie, J.A. and Barko, J.W. 1990. Influence of sediment and groundwater on the distribution and biomass of Myriophyllum spicatum L. in Devil's lake, Wisconsin. Journal of Freshwater Ecology 55: 417-426.

Linthurst, R.A. and Seneca, E.L. 1980. The effects of standing water and drainage potential on the Spartina alterniflora-substrate complex in a north Carolina salt marsh. Estuarine and Coastal Marine Science 11: 41- 52.

Lombardi, T., Fochetti, T., Bertacchi, A. and Onnis, A. 1997. Germination requirements in a population of Typha latifolia. Aquatic Botany 56: 1–10.

Lacroix, C. and Mosher, C. 1995. Early development and viability testing of embryos of Scirpus acutus Muhl. Aquatic Botany 50: 117-125.

Lombardi, T., Fochetti, T., Bertacchi, A. and Onnis ,A. 1997. Germination requirements in a population of Typha latifolia. Aquatic Botany 56: 1 – 10.

Long, S.P. and Mason, C.F. 1983. Salt Marsh Ecology. Blackie, Glasgow, UK.

Looman, J. 1983. Distribution of plant species and vegetation types in relation to climate. Vegetatio 54 (1): 17- 25.

Lovett-Doust, J. 1989. Plant reproductive strategies and resource allocation. Tree 4: 230–233.

Lubke, R.A. 1988a. Description of the coastal region. In: Lubke, R.A., Bruton, M.N. and Gess, F.W. Field Guide to the Eastern Cape Coast. Wildlife Society of Southern Africa, Grahamstown.

183 Lubke, R.A. and van Wijk, Y. 1988b. Estuarine plants. In: (R.A. Lubke, F.W. Grass and M.N. Bruton, eds.). A Field Guide to the Eastern Cape Coast. The Grahamstown Centre of the Wildlife Society of Southern Africa, Grahamstown. pp. 133-145.

Lucht, W., Prentice, I.C. and Myneni, R.B. 2002. Climatic control of the high-latitude vegetation greening trend and Pinatubo influence. Science 296: 1687-1689.

Mahall, B.E. and Park, R.B. 1976a. The ecotone between Spartina foliosa and Salicornia virginica in salt marshes of northern San Francisco Bay. II. Soil water and salinity. Journal of Ecology 64: 793–809.

Mahall, B.E. and Park, R.B. 1976b. The ecotone between Spartina foliosa and Salicornia virginica in salt marshes of northern San Francisco Bay. III. Soil aeration and tidal immersion. Journal of Ecology 64: 811– 819.

Maheu-Giroux, M. and de Blois, S. 2007. Landscape ecology of Phragmites australis invasion in networks of linear wetlands. Landscape Ecology. 22(2): 285-301.

Marcum, K.B. and Murdoch, C.L. 1992. Salt tolerance of the coastal salt marsh grass, Sporobolus virginicus (L.) Kunth. New Phytologist 120(2): 281-288.

Martinez-Ghersa, M.A., Ghersa, C.M., Benech-Arnold, R.L., Donough, R.M. and Sanchez, R.A. 2000. Adaptive traits regulating dormancy and germination of invasive species. Plant Species Biology 15: 127-137.

Mateos-Naranjo, E., Redondo-Gomez, E.S., Luque, C.J., Castellanos, E.M., Davy, A.J. and Figueroa, M.E. 2008. Environmental limitations on recruitment from seed in invasive Spartina densiflora on a southern European salt marsh. Estuarine, Coastal and Shelf Science 79: 727–732.

Martin, A.R.H. 1960. The Ecology of Groenvlei, a South African Fen: Part I. The Primary communities. Journal of Ecology 48(1): 55-71.

Matheson, F.E., de Winton, M.D., Clayton, J.S., Edwards, T.M. and Mathieson, T.J. 2005. Responses of vascular (Egeria densa) and non-vascular (Chara globularis) submerged plants and oospores to contrasting sediment types. Aquatic Botany 83: 141–153.

McCorry, M.J. and Renou, F. 2003. Ecology and management of Juncus effusus (sort rush) on cutaway peatlands. Forest Ecosystem Research Group Report Number 69. Department of Environmental Resource Management, University College Dublin. 66 pp.

McKee, J. and Richards, A.J. 1996. Variation in seed production and germinability in common reed (Phragmites australis) in Britain and France with respect to climate. New Phytologist 133: 233–243.

Menendez, M., Sanchez, A. 1998. Seasonal variations in F-I responses of Chara hispida L. and Potamogeton pectinatus L. from stream to Mediterranean ponds. Aquatic Botany 61: 1–15.

184 Menendez, M. 2008. Leaf growth, senescence and decomposition of Lam. in a coastal Mediterranean marsh. Aquatic Botany 89: 365–371.

Middleboe, A.L. and Markager, S. 1997. Depth limits and minimum light requirements of freshwater macrophytes. Freshwater Biology 37: 553–568.

Miller, G.T. 1992. Living in the Environment (7th Edition): An Introduction to Environmental Science. Wadsworth Publishing. California. pp 310-332.

Minchinton, T.E. 2002. Disturbance by wrack facilitates spread of Phragmites australis in a coastal marsh. Journal of Experimental Marine Biology and Ecology 281: 89–107.

Minello, T.J. 2000. Temporal development of salt marsh value for nekton and epifauna: utilization of dredged material marshes in Galveston Bay, Texas, USA. Wetlands Ecology and Management 8: 327-342.

Mitsch W.J. and Gosselink, J.G. 1993. Wetlands. New York, NY, USA: Van Nostrand Reinhold. 154 pp.

Mitsch W.J. and Gosselink, J.G. 2000. The value of wetlands: The importance of landscape setting and scale. Ecological Economics 35: 25-33.

Montagna, P.A., Alber, M., Doering, P. and Connor, M.S. 2002. Freshwater Inflow: Science, Policy, Management. Estuaries 25(6): 1243-1245.

Mony, C., Mercier, E., Bonis, A. and Bouzille, J.B. 2010. Reproductive strategies may explain plant tolerance to inundation: A mesocosm experiment using six marsh species. Aquatic Botany 92: 99–104.

Mooring, M.T., Cooper, A.W. and Seneca, E.D. 1971. Seed germination response and evidence for height ecophenes in Spartina alterniflora from North Carolina. American Journal of Botany 58: 48–55.

Morton, R.M., Pollock, B.R. and Beumer, J.P. 1987. The occurrence and diet of fishes in a tidal inlet to a salt marsh in southern Morton Bay, Queensland. Australian Journal of Ecology 12: 217-237.

Mullins, P.H. and Marks, T.C. 1987. Flowering phenology and seed production of Spartina anglica. Journal of Ecology 74: 1037–1048.

Muir, D. 2000. Habitat requirements and propagation of supratidal salt marsh species. MSc University of Port Elizabeth. 164 pp.

Mulder, C.P.H., Rues, R.W. and Singer, J.S. 1996. Effects of environmental manipulations on Tnglochin palustris: Implications for the role of goose herbivory in controlling its distribution. Journal of Ecology 84:267-278.

Naidoo, G. and Mundree S.G. 1993. Relationship between morphological and physiological-responses to waterlogging and salinity in Sporobolus virginicus (L.) Kunth. Oecologia 93: 360–366.

185 Naidoo, G. and Naidoo, Y. 1992. Waterlogging responses of Sporobolus virginicus (L.) Kunth. Oecologia 90: 445-450.

Naidoo, G. and Naidoo, Y. 1998. Salt tolerance in Sporobulus virginicus: The importance of ion relations and salt secretion. Flora 193 337-344.

Naidoo, G. and Naidoo, Y. 2000. Morphological and physiological responses of Sporobolus virginicus to flooding. South African Journal of Aquatic Sciences 25: 295-297.

Naidoo G. and Kift, J. 2006. Responses of the saltmarsh rush Juncus kraussii to salinity and waterlogging. Aquatic Botany 84: 217–225.

Nelson, P. and Adam, P. 1995. Plant community organization in New South Wales sahmarshes: Species mosaics and potential causes. Wetlands (Australia) 14:1-18.

Noe, G.B. 2002. Temporal variability matters: Effects of constant vs. varying moisture and salinity on germination. Ecological Monographs 72(3): 427-443.

Noe, G.B. and Zedler, J.B. 2000. Differential affects of four abiotic features on the germination of salt marsh annuals. American Journal of Botany 87(11): 1679–1692.

Noe, G.B. and Zedler, J.B. 2001. Spatio-temporal variation of salt marsh seedling establishment in relation to the abiotic and biotic environment. Journal of Vegetation Science 12: 61-74.

Ni, J., Harrison, S.P., Prentice, C.I., Kutzbach, J.E. and Sitch, S. 2006. Impact of climate variability on present and Holocene vegetation: A model-based study. Ecological Modelling 191: 469–486.

Nygaard, G. and Sand-Jensen, K. 1981. Light climate and metabolism of Nitella flexilis (L.) Ag. in the bottom waters of oligotrophic lake Grane Langso, Denmark. Hydrobiologia 66: 685–699.

Obrador, B. and Pretus, J.L. 2010. Spatio-temporal dynamics of submerged macrophytes in a Mediterranean coastal lagoon. Estuarine, Coastal and Shelf Science 87:145–155.

O‘Callaghan, M. 1987. Salt marshes along the Cape coast. In: (R.O. Walmsley and M.L. Bolten, eds.). Ecology and conservation of wetlands in South Africa. CSIR Report No. 28.

O‘Callaghan, M. 1990. Salt marshes along the Cape Coast. In: Ecology and Conservation of Wetlands in South Africa. CSIR Occasional Report No. 28.

O‘Callaghan, M. 1992. The ecology and identification of the southern African Salicornieae (Chenopodiaceae). National Botanical Institute. Stellenbosch.

Ohlsson, T. 1979. Redox reactions in soils sequence of redox reactions in a waterlogged soil. Vordic Hydrology: 89-98.

186 Olff, H., Bakker, J.P. and Fresco, L.F.M. 1988. The effect of fluctuations in tidal inundation frequency on a salt marsh vegetation. Vegetatio 78: 13-19.

Onaindia, M., Albizu, I. and Amezaga, I. 2001. Effect of time on the natural regeneration of salt marsh. Applied Vegetation Science 4 (2): 247-256.

O'Neill, E. J. 1972. Alkali bulrush seed germination and culture. Journal of Wildlife Management 36: 649-652.

Onuf, C.P. 2006. Aspects of the biology of Salicornia bigelovii Torr. in relation to a proposed restoration of a wind-tidal flat system on the south Texas, USA coast. Wetlands 26(3): 649-666.

Ostendorp, W. 1991. Damage by episodic flooding to Phragmites reeds in a pre-alpine lake: Proposal of a model. Oecologia 86: 119-124.

Ostendorp, W. 1999. Susceptibility of lakeside Phragmites reeds to environmental stresses: Examples from Lake Constance-Untersee (SW Germany). Limnologica 29(1): 21-27.

Packham, J.R. and Willis, A.J. 1997. Ecology of Dunes, Salt marsh and Shingle. London, UK: Chapman and Hall. 352 pp.

Palomo, L. and Niell, F.X. 2009. Primary production and nutrient budgets of Sarcocornia perennis ssp. alpini (Lag.) Castroviejo in the salt marsh of the Palmones River estuary (Southern Spain). Aquatic Botany 91: 130–136.

Pan, D., Bouchard, A., Legendre, P. and Domon, G. 1998. Influence of edaphic factors on the spatial structure of inland halophytic communities: A case study in China. Journal of Vegetation Science 9: 797–804.

Papastergiadou, E. and Babalonas, D. 1992. Ecological Studies on aquatic macrophytes of a dam. Lake Kerkini, Greece. Hydrobiologia 90(2): 187-206.

Pearcy, R.W. and Ustin, S.L. 1984. Effects of salinity on growth and photosynthesis of 3 California (USA) tidal marsh species. Oecologia 62: 68–73.

Penhale, P.A. and Wetzel, R.G. 1983. Structural and functional adaptations of eelgrass (Zostera marina L.) to the anaerobic sediment environment. Canadian Journal of Botany 61(5): 1421-1428.

Pennings, S.C. and R. M. Callaway. 1992. Salt marsh plant zonation: The relative importance of competition and physical factors. Ecology 73:681–690.

Pennings, S.C., Stanton, L.E. and Brewer, J.S. 2002. Nutrient effects on the composition of salt marsh plant communities along the southern Atlantic and gulf coasts of the United States. Estuaries 25 6A: 1164–1173.

Pennings, S.C., Grant, M. and Bertness, M.B. 2005. Plant zonation in low-latitude salt marshes: Disentangling the roles of flooding, salinity and competition. Journal of Ecology 93(1): 159-167.

187 Perez, M., Carlos M., Duarte, C.M., Romero, J., Sand-Jensen, K. and Alcoverro, T. 1994. Growth plasticity in Cymodocea nodosa stands: The importance of nutrient supply. Aquatic Botany 47: 249-264.

Perissinotto, R., Walker, D.R., Webb, P., Wooldridge, T.H. and Bally, R. 2000. Relationships between zoo- and phytoplankton in a warm temperate, semi-permanently closed estuary, South Africa. Estuarine, Coastal and Shelf Science 51: 1-11.

Perissinotto R., Blair, A., Connell, A., Demetriades, N.T., Forbes, A.T., Harrison, T.H., Lyer, K., Joubert, M., Kibirige, I., Mundree, S., Simpson, H., Stretch, D., Thomas, C., Thwala, X. and Zietsman, I. 2004. Contribution to information required for the implementation of Resource Directed Measures for estuaries (Volume 2). Responses of the biological communities to flow variation and mouth state in two KwaZulu- Natal temporarily open/closed estuaries. Water Research Commission Report No. 1247/2/04.

PERL (Pacific Estuarine Research Laboratory). 1990. A manual for assessing restored and natural coastal wetlands with examples from Southern California. California Sea Grant Report No. T-CSGCP-021. La Jolla, California. 105 pp.

Petzer, G. 2010. Ndlambe drought worsens. Talk of the Town. Port Alfred. pp 1.

Pezeshki, S.R., de Laune, R.D. and Patrick, Jr. W.H. 1989. Effect of fluctuating rhizosphere redox potential on carbon assimilation of Spartina alterniflora. Oecologia 80(1): 132-135.

Pezeshki, S.R., de Laune, R.D., Kludze, H.K. and Choi, H.S. 1996. Photosynthetic and growth responses of cattail (Typha domingensis) and sawgrass (Cladium jamaicense) to soil redox conditions. Aquatic Botany 54: 25-35.

Pezeshki, S.R. 2001. Wetland plant responses to soil flooding. Environmental and Experimental Botany 46: 299–312.

Phillips, G.L., Eminson, D. and Moss, B. 1978. A mechanism to account for the macrophyte decline in progressively eutrophicated freshwaters. Aquatic Botany 4: 103-126.

Philipupillai, J. and Ungar, I.A. 1984. The effect of seed dimorphism on the germination and survival of Salicornia europaea L. populations. American Journal of Botany 71(4): 542–549.

PieIou, E.C. and Routledge, R.D. 1976. Salt marsh vegetation: Latitudinal gradients in the zonation patterns. Oecologia (Berlin) 24: 311-321.

Pierce, S.M. 1983. Estimation of the non-seasonal production of Spartina maritima (Curtis) Fernald in a South African Estuary. Estuarine, Coastal and Shelf Science 16: 241-254.

Pierce, S.M. 1984. A synthesis of plant phenology in the Fynbos Biome. South African National Scientific Programmes Report No. 88. Pretoria. 57 pp.

188 Pip, E. 1979. Survey of the ecology of submerged aquatic macrophytes in central Canada. Aquatic Botany 7: 339-357.

Pollard, D.A. 1994. A comparison of fish assemblages and fisheries in intermittently open and permanently open coastal lagoons on the south coast of New South Wales. Estuaries 17(3): 631–646.

Ponnamperuma, F.N. 1972. The chemistry of submerged soils. Advances in Agronomy 24:29–96.

Pooley, E. 1998. A field guide to wild flowers of KwaZulu-Natal and the Eastern Region. Natal Flora Publications Trust, Durban. 630 pp.

Proctor, V.W. 1960. Dormancy and germination of Chara oospores. New Bulletin of the Phycological Society of America. 13(40): 64.

Pujol, J.A., Calvo, J.F. and Ramirez-Diaz, L. 2001. Seed germination, growth, and osmotic adjustment in response to NaCl in a rare succulent halophyte from southeastern Spain. Wetlands 21: 256-264.

Qu, X.X. and Huang, Z.Y. 2005. The adaptive strategies of halophyte seed germination. Acta Ecologica Sinica 25: 2389-2398.

Radke, L.C., Howard, K.W.F. and Gell, P.A. 2002. Chemical diversity in southeastern Australian saline lakes I. Geochemical causes. Marine and Freshwater Research 53: 1-19.

Rai, U.N., Sinha, S., Tripathi, R.D. and Chandra, P. 1995. Wastewater treatability potential of some aquatic macrophytes: Removal of heavy metals. Ecological Engineering 5: 5-12.

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. Australian Journal of Marine and Freshwater Research 45:1469–1479.

Reddering, J.V. 1988. Coastal and catchment basin controls on estuary morphology of the south-eastern Cape coast. South African Journal of Science 84: 154–157.

Redondo, S., Rubio-Casal, A.E., Castillo, J.M., Luque, C.J., Alvarez, A.A., Luque, T. and Figueroa, M.E. 2004. Influences of salinity and light on germination of three Sarcocornia taxa with contrasted habitats. Aquatic Botany 78: 255–264.

Redondo-Gómez, S., Castillo J.M., Luque C.J., Luque, T., Figueroa M.E. and Davy A.J. 2007. Fundamental niche differentiation in subspecies of Sarcocornia perennis on a salt marsh elevational gradient. Marine Ecology Programme Services 347:15-20.

Richards, J.H. and Caldwell, M.M. 1987. Hydraulic lift: Substantial nocturnal water transport between soil layers by Artemisia tridentata roots. Oecologia 73: 486-489.

189 Riddin, T. and Adams, J.B. 2008a. Influence of mouth condition and water level on the macrophytes in a small temporarily open/closed estuary. Estuarine, Coastal and Shelf Science 79: 86-92.

Riddin, T. and Adams, J.B. 2008b. Appendix G. Specialist Report: Macrophytes. In: van Niekerk, L., Bate, G.C., Whitfield, A.K. (Eds.), An Intermediate Ecological Reserve Determination Study of the East Kleinemonde Estuary, Water Research Commission Report 1581/2/08, Pretoria. 204 pp.

Riddin, T. and Adams, J.B. 2009. The seed banks of two temporarily open/closed estuaries in South Africa. Aquatic Botany 90: 328-332.

Riddin, T. and Adams, J.B. 2010. The effect of a storm surge event on the macrophytes of a temporarily open/closed estuary, South Africa. Estuarine, Coastal and Shelf Science 89: 119-123.

Riehl, E. and Ungar, I.A. 1982. Growth and ion accumulation in Salicornia europaea under saline field conditions. Oecologia 54: 193-199.

Riis, T., Sand-Jensen, K. and Vestergaard, O. 2000. Plant communities in lowland Danish streams: Species composition and environmental factors. Aquatic Botany 66: 255-272.

Riis, T. and Biggs, B.J.F. 2001. Distribution of macrophytes in New Zealand streams and lakes in relation to disturbance frequency and resource supply: A synthesis and conceptual model. New Zealand Journal of Marine and Freshwater Research 35: 255-267.

Ritter, A.F., Wasson, K., Lonhart, S.I., Preisler, R.K., Woolfolk, A., Griffith, K.A., Connors, S. and Heiman, K.W. 2008. Ecological signatures of anthropogenically altered tidal exchange in estuarine ecosystems. Estuaries and Coast 31: 554-557.

Rivers, W.G. and Weber, D.J. 1971. The influence of salinity and temperature on seed germination in Salicornia bigelovii. Physiologia Plantarum 24: 73–75.

Roberts, J., Young, B. and Marston, F. 2000. Estimating the water requirements for plants of floodplain wetlands: A guide. Occasional Paper 04/00. Land and Water Resources Research and Development Corporation. Canberra.

Rodiyati, A., Arisoesilaningsih, E., Isag, Y. and Nakagoshi, N. 2005. Responses of Cyperus brevifolius (Rottb.) Hassk. and Cyperus kyllingia Endl. to varying soil water availability. Environmental and Experimental Botany 53: 259–269.

Roelofs, J.G.M., Schuurkes, J.A.A.R. and Smits, A.J.M. 1984. Impact of acidification and eutrophication on macrophyte communities in soft waters. ii. Experimental studies. Aquatic Botany 18: 389-411.

Rogel, J.A., Ariza, F.A. and Silla, R.O. 2000. Soil salinity and moisture gradients and plant zonation in Mediterranean salt marshes of southeast Spain. Wetlands 20: 357-372.

190 Rogel, J.A., Silla, R.O. and Ariza, F.A. 2001. Edaphic characterization and sediment ionic composition influencing plant zonation in a semiarid Mediterranean salt marsh. Geoderma 99: 81-98.

Roozen A. J. M. and Westhoff. V. 1985. A study on long-term salt-marsh succession using permanent plots. Vegetatio 61(1/3): 23-32.

Rørslett, B. 1991. Principal determinants of aquatic macrophyte richness in northern European lakes. Aquatic Botany 39: 173–193.

Rozema, J., Luppes, E. and Broekman, R. 1985. Differential response of salt marsh species to variation of iron and manganese. Vegetatio 62: 293–301.

Rubio-Casal, A.E., Castillo, J.M., Luque, C.J. and Figueroa. M.E. 2001. Nucleation and facilitation in salt pans in Mediterranean salt marshes. Journal of Vegetation Science 12(6): 761-770.

Rubio-Casal, A.E., Castillo, J. M., Luque, C.J. and Figueroa M.E. 2003. Influence of salinity on germination and seed viability of two primary colonizers of Mediterranean salt pans. Journal of Arid Environments 53: 145– 154.

Ryan, D.A., Heap, A.D., Radke, L. and Heggie, D.T. 2003. Conceptual models of Australia‘s estuaries and coastal waterways: Applications for coastal resource management. Geoscience Australia Record 2003/09, 136 pp.

Samson, D.A. and Werk, K.S. 1986. Size-dependent effects in the analysis of reproductive effort in plants. American Naturalist 127: 667–680.

Sanchez, J.M., San Leon, D.G. and Izco, J. 2001. Primary colonisation of mudflat estuaries by Spartina maritima (Curtis) Fernald in Northwest Spain: Vegetation structure and sediment accretion. Aquatic Botany 69: 15–25.

Sand-Jensen, K. and Sondergaard, M. 1979. Distribution and quantitative development of aquatic macrophytes in relation to sediment characteristics in oligotrophic Lake Kalgaard, Denmark. Freshwater Biology 9: 1-11.

Sand-Jensen, K., Prahl, C. and Stokholm, H. 1982. Oxygen release from roots of submerged aquatic macrophytes. Oikos 38: 349-354.

Sand-Jensen, K. and Borum, J. 1983. Regulation of growth of eelgrass (Zostera marina) in Danish coastal waters. Marine Technical Society 17: 15-21.

Sand-Jensen, K. and Madsen, T.V. 1992. Patch dynamics of the stream macrophyte, Callitriche cophocarpa. Freshwater Biology 27(2): 277-282.

Santamaria, L. and Hootsmans, M.J.M. 1998. The effect of temperature on the photosynthesis, growth and reproduction of a Mediterranean submerged macrophyte, Ruppia drepanensis. Aquatic Botany 60: 169- 188.

191 Santamaria, L. and van Vierssen, W. 1997. Photosynthetic temperature responses of fresh- and brackish-water macrophytes: A review. Aquatic Botany 58:135-150.

Santamarıa-Gallegos, N.A., Sánchez-Lizaso, J.L. and Félix-Pico, E.F. 2000. Phenology and growth cycle of annual subtidal eelgrass in a subtropical locality. Aquatic Botany 66: 329–339.

Sartor, C.E. and Marone, L. 2010. A plurality of causal mechanisms explains the persistence or transience of soil seed banks. Journal of Arid Environments 74: 303–306.

Scarton, F., Day, J.W. and Rismondo, A. 2002. Primary production and decomposition of Sarcocornia fruticosa (L.) Scott and Phragmites australis Trin. Ex Steudel in the Po Delta, Italy. Estuaries 25(3): 325–336.

Schlacher, T.A. and Wooldridge, T.H. 1996. Ecological responses to reductions in freshwater supply and quality in South Africa's estuaries: Lessons for management and conservation. Journal of Coastal Conservation 2: 115-130.

Schulze, B.R. 1984. Climate of South Africa. Part 8. General survey. Weather Bureau, Department of Transport, Republic of South Africa, 300 pp.

Schumann, E., Largier, J. and Slinger, J. 1999. Estuarine hydrodynamics. In Allanson, B.R. and Baird, D. (Eds.) Estuaries of South Africa, Cambridge University Press. 289-321 pp.

Schwartz, M.D. 2003. Phenology: An Integrative Environmental Science. Tasks for Vegetation Science 39. Kluwer Academic Publishers. 592 pp.

Schwarz, A.M., Hawes, I. and Howard-Williams, C. 1996. The role of photosynthesis/light relationship in determining lower depth limits of Characeae in South Islands. New Zealand lakes. Freshwater Biology 35: 69–80.

Schwarz, A.M. and Hawes, I. 1997. Effects of changing water clarity on characean biomass and species composition in a large oligotrophic lake. Aquatic Botany 56: 169–181.

Schwarz, A.M., de Winton, M. and Hawes, I. 2002. Species-specific depth zonation in New Zealand charophytes as a function of light availability. Aquatic Botany 72: 209–217.

Seabloom, E.W., van der Valk, A.G. and Moloney, K.A. 1998. The role of water depth and soil temperature in determining initial composition of prairie wetland coenoclines. Plant Ecology 138, 203–216.

Sederias, J. and Colman, B. 2007. The interaction of light and low temperature on breaking the dormancy of Chara vulgaris oospores. Aquatic Botany 87: 229–234.

Semenya, B.L. 2010. Flowering characteristics of salt marsh plants at the Kowie Estuary. Honours Project. Nelson Mandela Metropolitan University. 47 pp.

192 Seneca, E.D. 1974. Germination and seedling responses of Atlantic and Gulf Coast populations of Spartina alterniflora. American Journal of Botany 61: 947-956.

Setchell, W.A. 1924. Ruppia and its environmental factors. Proceedings of the National Academy of Sciences of the United States of America 10(6): 286-288.

Sfriso, A. and Ghetti, A.P. 1998. Seasonal variation in biomass, morphometric parameters and production of seagrasses in the lagoon of Venice. Aquatic Botany 61(3): 207-223.

Shaw, A.L. 2007. Rehabilitation of the Orange River Mouth Salt Marsh: Seed, Wind and Sediment Characteristics. MSc Thesis. Nelson Mandela Metropolitan University. 164 pp.

Shaw, G.A., Adams, J.B. and Bornman, T.G. 2008. Sediment characteristics and vegetation dynamics as indicators for the potential rehabilitation of an estuary salt marsh on the arid west coast of South Africa. Journal of Arid Environments 72: 1097–1109.

Shepherd, K.A., Macfarlane, T.D. and Colmer, T.D. 2005. Morphology, anatomy and histochemistry of Salicornioideae (Chenopodiaceae) fruits and seeds. Annals of Botany 95: 917–933.

Sheppard, J.N. 2010. Structure and functioning of fish assemblages in two South African estuaries, with emphasis on the presence and absence of aquatic macrophyte beds. MSc. Rhodes University. Grahamstown. 178 pp.

Shumway, S.W. and Bertness, M.D. 1992. Salt stress limitation of seedling recruitment in a salt marsh plant community. Oecologia 92: 490-497.

Silva, H., Caldeira, G. and Freitas, H. 2006. Salicornia ramosissima population dynamics and tolerance of salinity. Ecological Research 22: 125-134.

Silvertown, J.W. 1988. The demographic and evolutionary consequences of seed dormancy. In: A.J. Davy, M.J. Hutchings and A.R. Watkinson (Editors), Plant Population Ecology. Blackwell Scientific, Oxford, pp. 205- 220.

Smith, C.J. and de Laune, R.D. 1984. Effect of sediment moisture on carbon dioxide exchange in Spartina alterniflora. Plant and Soil 79: 291-293.

Snow, A. and Vince, S.W. 1984. Plant zonation in an Alaskan salt marsh. II. An experimental study of the role of edaphic conditions. Journal of Ecology 72: 669–684.

Snow, G.C. and Adams, J.B.A. 2005. Response of micro-algae in the Kromme Estuary to managed freshwater inputs. Water SA 32: 71-80.

Snow, G.C. and Taljaard, S. 2007. Water quality in South African temporarily open/closed estuaries: a conceptual model. African Journal of Aquatic Science 32(2): 99-111.

193 Soulié-Marsch, I. 2008. Charophytes, indicators for low salinity phases in North African sebkhet. Journal of African Earth Sciences 51: 69–76.

Spence, D.H.N. 1982. The zonation of plants in freshwater lakes. Advances in Ecological Research 12: 37-125.

Spencer, D.F. and Ksander, G.G. 2002. Sedimentation disrupts natural regeneration of Zannichellia palustris in Fall River, California. Aquatic Botany 73: 137–147.

Srivastava, D.S., Staicer, C. and Freedman, B. 1995. Aquatic vegetation of Nova Scotian lakes differing in acidity and trophic status. Aquatic Botany 51: 181–196.

Stanton, L.E. 2005. The establishment, expansion and ecosystem effects of Phragmites australis, an invasive species in Coastal Louisiana. PhD Thesis. Louisiana State University and Agricultural and Mechanical College. 166 pp.

Stearns, S.C. 1976. Life history tactics: A review of the ideas. Quarterly Review of Biology 51: 3-47.

Steffen, S., Mucina, L. and Kadereit, G. Unpublished. Revision of Sarcocornia (Chenopodiaceae) in South Africa, Namibia and Mozambique.

Steffen, S., Mucina, L. and Kadereit, G. 2009. Three new species of Sarcocornia (Chenopodiaceae) from South Africa. Kew Bulletin, Volume 64: 447–459.

Steinman, A.D., Meeker, R.H., Rodusky, A.J., Davis, W.P. and Hwang, S.J. 1997. Ecological properties of charophytes in a large subtropical lake. Journal of the North American Benthological Society 16: 781–793.

Steinman, A.D., Havens, K.E., Rodusky, A.J., Sharfstein, B., James, R.T. and Harwell, M.C. 2001. The influence of environmental variables and a managed water recession on the growth of charophytes in a large, subtropical lake. Aquatic Botany 72: 297–313.

Stockey, A. and Hunt, R. 1992. Fluctuating water conditions identify niches for germination in Alisma plantago- aquatica. Acta Oecologia 113: 227-229.

Streever, W.J., Wiseman, I., Turner P. and Nelson, P. 1996. Short term changes in flushing of tidal creeks following culvert removal. Wetlands (Australia) 15: 21-29.

Streever, W.J. and Genders, A.J. 1997. Effect of improved tidal flushing and competitive interactions at the boundary between salt marsh and pasture. Estuaries 20(4): 807-818.

Stross, R.G. 1979. Density and boundary regulations of the Nitella meadows in Lake George, New York. Aquatic Botany 6: 285–300.

Struyf, E., van Damme, S., Gribsholt, B., Bal, K., Beauchard, O., Middelburg, J.J. and Meire, P. 2007. Phragmites australis and silica cycling in tidal wetlands. Aquatic Botany 87: 134–140.

194 Strydom, N.A. 2002. Dynamics of early stage fishes associated with selected warm temperate estuaries in South Africa. PhD Thesis. Rhodes University. Grahamstown. 166 pp.

Sun, S., Cai, Y. and Tian, X. 2003. Salt marsh vegetation change after a short-term tidal restriction in the Changjiang Estuary. Wetlands 23(2): 257–266.

Svedäng, M.A. 1990.The growth dynamics of Juncus bulbosus L. - A strategy to avoid competition? Aquatic Botany 37(2): 123-138.

Taljaard, S., van Niekerk, K. and Joubert, W. 2009. Extension of a qualitative model on nutrient cycling and transformation to include microtidal estuaries on wave dominated coasts: Southern Hemisphere perspective. Estuarine, Coastal and Shelf Science 33: 325-338.

Taylor, D.I. 1983. The effects of a major macrophytes regression upon the primary production in the littoral of Swartvlei. Archiv für Hydrobiologia 96(3): 345-353.

Taylor, R., Adams, J.B. and Haldorsen, S. 2006. Primary habitats of the St Lucia Estuarine System, South Africa, and their responses to mouth management. African Journal of Aquatic Science 31(1): 31–41.

Ter Braak, C.J.F. and Šmilauer, P. 2002. CANOCO Reference Manual and CanoDraw for Windows User's Guide: Software for Canonical Community Ordination (version 4.5).

Thompson, K. and Grime, J.P. 1983. A comparative study of germination responses to diurnally-fluctuating temperatures. Journal of Applied Ecology 20: 141-156.

Teske, P.R. and Wooldridge, T.H. 2003. What limits the distribution of subtidal macrobenthos in permanently open and temporarily open/closed South African estuaries? Salinity vs sediment particle size. Estuarine, Coastal and Shelf Science 57: 225-238.

Tinley, K.L., 1985. Coastal Dunes of South Africa. South African National Scientific Programmes Report 109: 1- 300.

Titus, J.E., Feldman, R.S. and Grisé, D. 1990. Submersed macrophyte growth at low pH. CO² enrichment effects with fertile sediment. Oecologia 84(3): 307-313

Titus, J.E. and Hoover, D.T. 1991. Toward predicting reproductive success in submersed freshwater angiosperms. Aquatic Botany 41: 11-136.

Titus, J.E., Grisé, D., Sullivan, G. and Stephens, M.D. 2004. Monitoring submersed vegetation in a mesotrophic lake: correlation of two spatio-temporal scales of change. Aquatic Botany 79: 33–50.

Tölken, H.R. 1967. The species Arthrocnemum and Salicornia (Chenopodiaceae) in southern Africa. Bothalia 9: 255–307.

195 Torn, K., Martin, G. and Paalme, T. 2006. Seasonal changes in biomass, elongation growth and primary production rate of Chara tomentosa in the NE Baltic Sea. Annales Botanici Fennici 43(4): 276-283.

Troyo-Dieguez, E., Ortega-Rubio, A., Maya, Y. and Leon, J.L. 1994. The effect of environmental conditions on the growth and development of the oilseed halophyte Salicornia bigelorii Torr. in arid Baja California Sur, Mexico. Journal of Arid Environments 28: 207-213.

Touchette, B.W. 2007. The biology and ecology of seagrasses: Preface. Journal of Experimental Marine Biology and Ecology 350: 1-2.

Tubbs, C.R. and Tubbs, J.M. 1983. The distribution of Zostera and its exploitation by wildfowl in the Solent, southern England. Aquatic Botany 15: 223-239.

Turpie, J.K., Adams, JB., Joubert, A., Harrison, TD., Colloty. BM., Maree, EC., Whitfield, AK., Wooldridge, TH., Lamberth, SJ., Taljaard, S. and Van Niekerk, L. 2002. Assessment of the conservation priority of status of South African estuaries for use in management and water allocation. Water SA Vol. 28 NO. 2.

Tyler-Walters, H. 2001. Ruppia maritima, Beaked tasselweed. Marine Life Information Network: Biology and Sensitivity Key Information Sub-programme (on-line), Marine Biological Association of the United Kingdom, Plymouth. http://www.marlin.ac.uk/species/Ruppiamaritima.htm.

Uncles, R.J., Stephens, J.A. and Smith, R.E. 2002. The dependence of estuarine turbidity on tidal intrusion length, tidal range and residence time. Continental Shelf Research 22: 1835-1856.

Ungar, I.A. 1962. Influence of salinity on seed germination of halophyte plants. Ecology 23: 763-764.

Ungar, I.A. 1977. Salinity, temperature, and growth regulator effects on seed germination of Salicornia europaea L. Aquatic Botany 3: 329-335.

Ungar, I.A. 1978. Halophyte seed germination. The Botanical Review 44: 233–264.

Ungar, I.A., Kenner, D.K. and McGraw, D.C. 1979. The distribution and growth of Salicornia europaea on an inland salt plan. Ecology 60: 329-336.

Ungar, I.A. 1987a. Population ecology of halophyte seeds. The Botanical Review 53: 301-334.

Ungar, I.A. 1987b. Population characteristics, growth, and survival of the halophyte Salicornia europaea. Ecology 68(3): 569-575.

Ungar, I.A. 1995. Seed germination and seed bank ecology in halophytes, pp. 599–628, In J. Kigel and G. Galili (eds.) Seed Development and Germination. Marcel Dekker, Inc., New York, NY, USA.

Ungar, I.A. 2001. Seed banks and seed population dynamics of halophytes. Wetlands Ecology and Management 9: 499–510.

196 Ursino, N., Silvestri, S. and Marani, M. 2004. Subsurface flow and vegetation patterns in tidal environments. Water Resources Research 40(5): 1-11.

USDA, NRCS. 2010. The Plants Database. National Plant Data Center, Baton Rouge, LA 70874-4490, USA. van Donk, E., and Van de Bund, W.J. 2001. Impact of submerged macrophytes including charophytes on phyto-and zooplankton communities: Allelopathy versus other mechanisms. Aquatic Botany 72: 261-274. van den Berg, M.S., Coops, H., Simons, J. and de Keizer, A. 1997. Competition between Chara aspera and Potamogeton pectinatus as a function of temperature and light. Aquatic Botany 60: 241–250. van den Berg, M.S., Scheffer, M., Nes, E.V. and Coops, H. 1999. Dynamics and stability of Chara sp. and Potamogeton pectinatus in a shallow lake changing in eutrophication level. Hydrobiologia 408/409: 335– 342. van den Berg, M.S., Coops, H. and Simons, J. 2001. Propagule bank buildup of Chara aspera and its significant colonization of a shallow lake. Hydrobiologia 462: 9–17. van den Brink, F.W.B., Van der Velde, G., Bosman, W.W. and Coops, H. 1995. Effects of substrate parameters on growth responses of eight helophyte species in relation to flooding. Aquatic Botany 50: 79-97. van der Sman, A.J.M., Joosten, N.N. and Blom, C.W.P.M. 1993. Flooding regimes and life history characteristics: short-lived species in river forelands. Journal of Ecology 81: 121–130. van der Valk, A.G. 1996. The effects of prolonged flooding on the distribution and biomass of emergent species along a freshwater wetland coenocline. Vegetatio 110: 185–196. van der Valk, A.G. 2007. Development of post-disturbance vegetation in prairie wetlands. Plant Disturbance Ecology: 341-370. van Eck, W.H.J.M., Lenssen J.P.M., van de Steeg H.M., Blom C.W.P.M. and de Kroon, H. 2006. Seasonal dependent effects of flooding on species survival and zonation. A comparative study of 10 terrestrial plant species. Hydrobiologia 565: 59-69. van Nes, E.H., van den Berg, M.S., Clayton, J.S., Coops, H. and Scheffer, M. 1999. A simple model for evaluating the costs and benefits of aquatic macrophytes. Hydrobiology 415: 335–339. van Niekerk, L., Huizinga, P. and Theron, A. 2002. Semi-closed mouth states in estuaries along the South African coastline. In: Environmental Flows For River Systems Proceedings. Fourth International Ecohydraulics Symposium 31: 1 (ISSN 0378-4738). van Niekerk, L., Cowley, P.D., Bornman, T.G. 2008a. Appendix C. Specialist Report: Physical processes. In: van Niekerk, L., Bate, G.C., Whitfield, A.K. (Eds.), An Intermediate Ecological Reserve Determination Study of the East Kleinemonde Estuary Water Research Commission Report 1581/2/08, Pretoria.

197 van Niekerk, L., Bate, G.C. and Whitfield, A.K. (Eds). 2008b. An Intermediate Ecological Reserve Determination Study of the East Kleinemonde Estuary Water Research Commission. WRC Report: 1581/11/08. van Vierssen, W., van Kessel, C.M. and van der Zee, J.R. 1984. On the germination of Ruppia taxa in western Europe. Aquatic Botany 19 (3-4): 381-393. van Vierssen, W. 1993. Relationships between survival strategies of aquatic weeds and control measures. In: Pieterse, A.H., Murphy, K.J. (Eds.), Aquatic Weeds: The Ecology and Management of Nuisance Aquatic Vegetation. Oxford University Press, New York, pp. 238–253. van Wijck, C., Grillas, P., de Groot, C.J. and Ham, T.L. 1994. A comparison between the biomass production of Potamogeton pectinatus L. and Myriophyllum spicatum L. in the Camargue (southern France) in relation to salinity and sediment characteristics. Vegetatio 113: 171-180. van Wijck, C., de Groot, C.J. and Grillas, P. 1992. The effect of anaerobic sediment on the growth of Potamogeton pectinatus: The role of organic matter, sulphide, and ferrous iron. Aquatic Botany 44, 31–49. van Wijck, R.J. 1989. Ecological studies on Potamogeton pectinatus L.L V. Nutritional ecology, in vitro uptake of nutrients and growth limitation. Aquatic Botany 35: 319-335.

Varty, A.K. and Zedler, J.B. 2008. How Waterlogged Microsites Help an Annual Plant Persist Among Salt Marsh Perennials. Estuaries and Coasts: 31:300–312.

Vaughan, K.L., Rabenhorst, M.C. and Needelman, B.A. 1996. Saturation and temperature effects on the development of reducing conditions in soils. Wetland Soils 73(2): 663-667.

Vieira Jr., J. and Necchi Jr., O. 2003. Photosynthesis characteristics of charophytes from tropical lotic ecosystems. Phycological Research 51: 51–60.

Verhoeven, J.T.A. 1979. The ecology of Ruppia-dominated communities in Western Europe. I. Distribution of Ruppia representatives in relation to their autecology. Aquatic Botany 6: 197-267.

Verhoeven, J.T.A. 1980. The ecology of Ruppia-dominated communities in Western Europe. III. Aspects of production, consumption and decomposition. Aquatic Botany 6: 197-267.

Vestergaard, O. and Sand-Jensen, K. 2000. Alkalinity and trophic state regulate aquatic plant distribution in Danish lakes. Aquatic Botany 67: 85–107.

Vicente, M.J., Conesa, E., Alvarez-Roge, J., Franco, J.A. and Martinez-Sanchez, J.J. 2007. Effects of various salts on the germination of three perennial salt marsh species. Aquatic Botany 87: 167–170.

Vince, S.W. and Snow, A.A. 1984. Plant Zonation in an Alaskan Salt Marsh: I. Distribution, Abundance and Environmental Factors. Journal of Ecology 72(2): 651-667.

198 Vleeshouwers, L.M., Bouwmeester, H.J. and Karssen, C.M. 1995 Redefining seed dormancy: An attempt to integrate physiology and ecology. Journal of Ecology 83(6): 1031-1037.

Vorwerk, P.D., Whitfield, A.K., Cowley, P.D. and Paterson, A.W. 2001. A survey of selected Eastern Cape estuaries with particular reference to the ichthyofauna. Ichthyological Bulletin 72: 1-52.

Vretare, V., Weisner, S.E.B., Strand, J.A. and Granéli, W. 2001. Phenotypic plasticity in Phragmites australis as a functional response to water depth. Aquatic Botany 69: 127–145.

Vretare, V. 2002. The influence of ventilation systems on water depth penetration of emergent macrophytes. Freshwater Biology 47: 1097-1105.

Waisel, Y. 1972. Biology of Halophytes. Academic Press, NY, USA. 395 pp.

Walker, D.R. 2004. Plant and algal distribution in response to environmental variables in selected Eastern Cape estuaries. PhD Thesis. University of Port Elizabeth, Port Elizabeth. 176 pp.

Wang, H., Yu, D. and Xiao, K. 2008. The interactive effects of irradiance and photoperiod on Chara vulgaris L.: Concerted responses in morphology, physiology, and reproduction. Hydrobiologia 6(10): 33–41.

Wang, H., Yu, D. and Xiao, K. 2009. Study on the phenology of Chara vulgaris in Xin‘an Spring, north China. Frontiers of Biology in China 4(2): 207-213.

Ward, T., Butler, E. and Hill, B. 1998. Environmental Indicators for National State of the Environment Reporting, Estuaries and the Sea, Commonwealth of Australia. 81 pp.

Warwick, N.W.M. and Brock, M.A. 2003. Plant reproduction in temporary wetlands: The effects of seasonal timing, depth, and duration of flooding. Aquatic Botany 77: 153–167.

Weisner, S.E.B. and Graneli, W. 1989. Influence of substrate conditions on the growth of Phragmites australis after a reduction in oxygen transport to below-ground parts. Aquatic Botany 35: 71-80.

Weisner, S.E.B., Eriksson, P.G., Graneli, W. and Leonardson, L. 1994. Influence of macrophytes on nitrate removal in wetlands. Ambio 23: 363–366.

Weisner, S.E.B. 1996. Effects of organic sediment on performance of young Phragmites australis clones at different water depth treatments. Hydrobiologia 330: 189-194.

Weisser, P.J., Whitfield, A.K. and Hall, C.M. 1992. The recovery and dynamics of submerged aquatic macrophyte vegetation in the Wilderness lakes, southern Cape. Bothalia 22: 283-288.

Welling, C.H., Pederson, R.L. and van der Valk, A.G. 1988. Temporal patterns in recruitment from the seed bank during drawdowns in a prairie wetland. Journal of Applied Ecology 25: 999–1007.

199 West, R.J. and King, R.J. 1996. Marine, brackish and freshwater fish communitities in the vegetated and bare shallows of an Australian coastal river. Estuaries 19(1): 31-41.

Westlake, D.F. 1963. Comparisons of plant productivity. Biological Reviews 38: 385–425.

Wetzel, R.G., Brammer, E.S., Lindström, L. and Forsberg, C. 1985. Photosynthesis of submersed macrophytes

in acidified lakes II. Carbon limitation and utilization of benthic CO2 sources. Aquatic Botany 22: 107-120.

Whitfield, A.K. 1980. A quantitative study of the trophic relationships within the fish community of the Mhlanga Estuary, South Africa. Estuarine and Coastal Marine Science 10: 417-435.

Whitfield, A.K. 1984. The effects of prolonged aquatic macrophyte senescence on the biology of the dominant fish species in a southern African coastal lake. Estuarine, Coastal and Shelf Science 18: 315-329.

Whitfield, A.K. 1989. The benthic invertebrate community of a southern Cape estuary: structure and possible food sources. Transactions of the Royal Society of Southern Africa 47: 159-179.

Whitfield, A.K. 1992. A characterization of southern African estuarine systems. South African Journal of Aquatic Science 18: 89-103.

Whitfield, A.K. 1998. Biology and ecology of fishes in southern african Estuaries. Ichthyological Monographs of the J.L.B. Smith Institute of Ichthyology 2. 223 pp.

Whitfield, A.K. 2000. Available scientific information on individual South African estuarine systems. Water Research Commission Report No. 577/3/00.

Whitfield, A.K. 2005. Fishes and freshwater in southern African estuaries - A review. Aquatic Living Resources 18: 275-289.

Whitfield, A.K. and Bate, G. 2007. A Review of Information on Temporarily Open/Closed Estuaries in the Warm and Cool Temperate Biogeographic Regions of South Africa, with Particular Emphasis on the Influence of River Flow on These Systems. Interim report to the Water Research Commission on the Project ―The freshwater requirements of intermittently open Cape estuaries‖ WRC Report No 1581/1/07.

Whitfield, A.K., Adams, J.B., Bate, G.C., Bezuidenhout, K., Bornman, T.G., Cowley, P.D., Froneman, P.W., Gama, P.T., James, N.C., Mackenzie, B., Riddin, T., Snow, G.C., Strydom, N.A., Taljaard, S., Terörde, A.I., Theron, A.K., Turpie, J.K., van Niekerk, L., Vorwerk, P.D. and Wooldridge, T.H. 2008. A multidisciplinary study of a small, temporarily open/closed South African estuary, with particular emphasis on the influence of mouth state on the ecology of the system. African Journal of Marine Science 30(3): 453–473.

Wiegleb, G. 1988. Analysis of flora and vegetation in rivers: Concepts and applications. In: Symoens, J.J. (Ed.), Vegetation of inland waters. Handbook of Vegetation Science, vol. 15/1. Kluwer Academic Publishers, Dordrecht. pp. 311–340.

200 Wiegleb, G. and Brux, H. 1991. Comparison of life history characteristics of broad-leaved species of the genus Potamogeton L. I. General characterization of morphology and reproductive strategies. Aquatic Botany 39: 131–146.

Wijte, A.H.B.M. and Gallagher, J.L. 1996. Effect of oxygen availability and salinity on early life history stages of salt marsh plants. I. Different germination strategies of Spartina alterniflora and Phragmites australis (Poaceae). American Journal of Botany 83(10): 1337-1342.

Wilcox, K. and Petrie, S. 2000. Investigation and long term monitoring of Phragmites australis at Long Point, Lake Erie, Ontario. Longpoint Waterfowel and Wetlands Research Fund. Ontario. 6 pp.

Wilman, V. 2006. Bolboschoenus maritimus: Working for Wetlands. Peninsula Project. Cape Town. 5 pp.

Winter, U., Meyer, M.I.B. and Kirst, G.O. 1987. Seasonal changes of ionic concentrations in the vascuolar sap of Chara vulgaris L. growing in a brackish water lake. Oecologia (Berlin) 74:122-127.

Winter, U. and Kirst, G.O. 1990. Salinity response of a freshwater charophyte, Chara vulgaris. Plant, Cell and Environment 13: 123–134.

Wolaver, T.G., Zieman, J. and Kjerfve, B. 1986. Factors affecting short-term variability in sediment pH as a function of marsh elevation in a Virginia mesohaline marsh. Journal of Experimental Marine Biology and Ecology 101: 227-237.

Wolff, S.L. and Jefferies, R.L. 1987. Morphological and isozyme variation in Salicornia europaea (s.l.) (Chenopodiaceae) in northeastern North America. Canadian Journal of Botany 65: 1410–1419.

Wolters, M., Garbutt, A., Bakker, J.P. and Carey, P.D. 2008. Restoration of salt-marsh vegetation in relation to site suitability, species pool and dispersal traits. Journal of Applied Ecology 45: 904–912.

Wood, E.J.F. 1959. Some east Australian sea-grass communities. Proceedings of the Linnean Society of New South Wales 84: 218-226.

Wood, R.D. 1950. Stability and zonation of Characeae. Ecology 31: 642-647.

Woodell, S.J. 1985. Salinity and seed germination patterns in coastal plants. Plant Ecology 61(1-3): 223-229.

Yamasaki, S. and Tange, I. 1981. Growth responses of Zizania latifolia, Phragmites australis and Miscanthus sacchariflorus to varying inundation. Aquatic Botany 10: 229-239.

Yong, H., Wenjie, D., Xiaoyin, G. and Li, D. 2007. Terrestrial growth in China and its relationship with climate based on the MODIS data. Acta Ecologica Sinica 27(12): 5086−5092.

Young, D.R. and Nobel, P.S. 1986. Predictions of soil-water potentials in the north-western Sonoran Desert. Journal of Ecology 74: 143-154.

201 Young, G.C. and Potter, I.C. 2002. Influence of exceptionally high salinities, marked variations in freshwater discharge and opening of estuary mouth on the characteristics of the ichthyofauna of a normally-closed estuary. Estuarine, Coastal and Shelf Science 55: 223–246.

Zare, G. and Keshevarzi, M. 2007. Morphological study of Salicorniaea (Chenopodiaceae) native to Iran. Pakistan Journal of Biological Sciences 10(6): 852-860.

Zedler, J.B., Covin, J. and Norby, C. 1986. Catastrophic events reveal the dynamic nature of salt marsh vegetation in southern California. Estuaries 9(1): 75-80.

Zedler, J.B., Paling, E. and McComb, A. 1990. Differential responses to salinity help explain the replacement of native Juncus kraussii by Typha orientalis in western Australian salt marshes. Australian Journal of Ecology 15(1): 57-72.

Zedler, J.B., Nordby, C.S. and Kus, B.E. 1992. The ecology of Tijuana Estuary, California: A national estuarine research reserve. NOAA Office of Coastal Resource Management, Sanctuaries and Reserves Division, Washington, D.C., USA. 168 pp.

Zedler, J.B., Nelson, P. and Adam, P. 1995. Plant community organization in New South Wales salt marshes: Species mosaics and potential causes. Wetlands (Australia) 14:1-18.

Zedler, J.B. (Principal Author). 1996. Tidal Wetland Restoration: A Scientific Perspective and Southern California Focus. Published by the California Sea Grant College System, University of California, La Jolla, California. Report No. T-038. 56 pp.

Zedler, J.B., Callaway, J.C., Desmond, J.S., Vivian-Smith, G., Williams, G.D., Sullivan, G., Brewster, A.E. and Bradshaw, B.K. 1999. Californian salt marsh vegetation: An improved model of spatial pattern. Ecosystems 2: 19-35.

Zedler, J.B., Callaway, J.C. and Sullivan, G. 2001. Declining biodiversity: Why species matter and how their functions might be restored in Californian tidal marshes. Bioscience 51: 1005-1017.

Zedler, J.B., Morzaria-Luna, H. and Ward, K. 2003. The challenge of restoring vegetation on tidal, hypersaline substrates. Plant and Soil 253: 259–273.

Zedler, J.B., Bonin, C.L., Larkin, D.J. and Varty, A. 2008. Salt Marshes. Encyclopedia of Ecology. 3132-3141 pp.

Zedler, J.B., Harding, B. and Williams, G. Unknown. Inter-continental comparisons of marsh restoration challenges. Unknown.

202 8. APPENDIX

8.1 ABIOTIC CONDITIONS IN THE EAST KLEINEMONDE AND KOWIE ESTUARIES 8.1.1 Supratidal habitat

T-test for Dependent Samples (EK and KW Supratidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Redox EK -11.8814 195.4850

Redox KW 148.6565 133.3525 51 -160.538 202.3741 -5.66510 50 0.000001

T-test for Dependent Samples (EK and KW Supratidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

pH EK 6.757255 0.987933

pH KW 7.662157 0.376788 51 -0.904902 0.981551 -6.58376 50 0.000000

T-test for Dependent Samples (EK and KW Supratidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

EC EK 40.97471 12.22387

EC KW 28.64157 9.56085 51 12.33314 14.94523 5.893267 50 0.000000

T-test for Dependent Samples (EK and KW Supratidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Salinity EK 21.20882 6.162586

Salinity KW 15.69725 5.954082 51 5.511569 8.026373 4.903893 50 0.000010

8.1.2 Intertidal habitat

T-test for Dependent Samples (EK and KW intertidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Redox EK -79.0971 141.2904

Redox KW 107.9421 140.2773 34 -187.039 208.6499 -5.22702 33 0.000009

T-test for Dependent Samples (EK and KW intertidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

pH EK 7.353235 0.517322

pH KW 7.751176 0.315554 34 -0.397941 0.501143 -4.63017 33 0.000055

T-test for Dependent Samples (EK and KW intertidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

EC EK 46.46441 13.91963

EC KW 35.95147 10.55010 34 10.51294 17.10735 3.583282 33 0.001079

T-test for Dependent Samples (EK and KW intertidal variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Salinity EK 24.31676 7.529034

Salinity KW 20.27588 6.477323 34 4.040882 9.085760 2.593310 33 0.014063

203 8.1.3 Reed and sedge habitat

T-test for Dependent Samples (EK and KW Reeds&Sedges variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Redox EK -126.791 111.9747

Redox KW -42.443 220.6156 34 -84.3479 214.9315 -2.28831 33 0.028658

T-test for Dependent Samples (EK and KW Reeds&Sedges variables) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

pH EK 7.334412 0.466139

pH KW 7.766176 0.268114 34 -0.431765 0.513192 -4.90576 33 0.000024

8.2 MACROPHYTE PHENOLOGY OF THE SELECTED SPECIES IN THE EAST KLEINEMONDE AND KOWIE ESTUARIES 8.2.1 Juncus kraussii and Juncus acutus Juncus kraussi (Jk)

T-test for Dependent Samples (Jk) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Live Cover Mar9 46.15909 24.19821

Live Cover Apr9 30.63636 17.31631 11 15.52273 15.53482 3.314043 10 0.007826

T-test for Dependent Samples (Jk) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Live Cover Nov9 25.63636 26.46610

Live Cover Dec9 15.90909 20.63712 11 9.727273 8.753181 3.685713 10 0.004207

T-test for Dependent Samples (Jk) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Live Cover Jan10 16.81818 17.64550

Live Cover Feb10 25.18182 26.50214 11 -8.36364 11.33378 -2.44747 10 0.034407

Juncus acutus (Ja)

T-test for Dependent Samples (Kowie Seasons JA) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer 52.80000 2.520582

Winter 46.42500 1.738534 4 6.375000 2.912473 4.377723 3 0.022061

T-test for Dependent Samples (Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Height Nov9 91.85000 19.17903

Height Dec9 84.61000 15.47253 10 7.240000 9.140958 2.504649 9 0.033605

204 Inflorescence number

T-test for Dependent Samples (Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Sep 09 Inf 8.80000 8.41691

Oct 09 Inf 41.20000 35.96542 10 -32.4000 33.16692 -3.08916 9 0.012947

T-test for Dependent Samples (Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov09 Inf 40.50000 31.44042

Dec09 Inf 29.30000 26.52064 10 11.20000 13.26482 2.670032 9 0.025621

Seed numbers

T-test for Dependent Samples (Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct09 Seed 0.0 0.0

Nov09 Seed 205897.1 271898.0 10 -205897 271898.0 -2.39466 9 0.040249

T-test for Dependent Samples (Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar-10 Seed 318748.5 243196.3

Apr-10 Seed 149720.2 149699.4 10 169028.3 109818.1 4.867270 9 0.000887

Juncus kraussi (EK) versus Juncus acutus (KW)

T-test for Independent Samples (EK vs KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

JK Cover vs. JA Cover 17.00000 46.00000 -3.48337 19 0.002487 11 10 16.90562 21.18700 1.570640 0.491328

Inflorescence number

T-test for Dependent Samples (Jk vs Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov2009 EK 0.50000 0.84984

Nov2009 KW 40.50000 31.44042 10 -40.0000 31.79797 -3.97796 9 0.003216

T-test for Dependent Samples (Jk vs Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb2009 EK 3.60000 6.05897

Feb2009 KW 10.60000 10.91584 10 -7.00000 6.548961 -3.38007 9 0.008126

Seed numbers

205 T-test for Dependent Samples (Jk vs Ja) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov2009 EK 0.0 0.0

Nov2009 KW 205897.2 271898.0 10 -205897 271898.0 -2.39466 9 0.040249

Seed viability Juncus kraussii T-test for Dependent Samples (JK Seed) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Feb10 EK 95.00000 5.00000

Jan10 EK 60.00000 20.31010 5 35.00000 16.58312 4.719399 4 0.009176

Juncus acutus

T-test for Dependent Samples (KW JA Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov-09 77.00000 4.47214

Dec-09 52.00000 15.24795 5 25.00000 16.95582 3.296902 4 0.030020

T-test for Dependent Samples (KW JA Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Dec-09 52.00000 15.24795

Jan-10 79.00000 6.51920 5 -27.0000 18.23458 -3.31095 4 0.029628

T-test for Dependent Samples (KW JA Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar-10 88.00000 10.36822

Apr-10 69.00000 11.93734 5 19.00000 13.87444 3.062127 4 0.037580

T-test for Dependent Samples (KW JA Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr-10 69.00000 11.93734

May-10 98.00000 4.47214 5 -29.0000 8.215838 -7.89280 4 0.001394

8.2.2 Sporobolus virginicus East Kleinemonde Estuary

T-test for Dependent Samples (Sporob Feb9-Feb10) Marked differences are significant at p < .05000

206 Mean Std.Dv. N Diff. Std.Dv. t df p

Aug9 Live Cover 37.27273 21.48996

Sep9 Live Cover 5.00000 10.00000 11 32.27273 30.36146 3.525408 10 0.005489

T-test for Dependent Samples (Sporob Feb9-Feb10) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Live Cover 37.27273 10.80825

Jan10 Live Cover 0.90909 3.01511 11 36.36364 13.43334 8.978003 10 0.000004

T-test for Dependent Samples (EK sporob Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Height Feb9 11.57778 2.732571

Height Mar9 14.39000 2.941624 9 -2.81222 2.896307 -2.91290 8 0.019502

Kowie Estuary

T-test for Dependent Samples (Sporob Kowie Heights) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar9 Height 11.36500 6.559643

Aug9 Height 16.06971 7.639392 10 -4.70471 4.154431 -3.58114 9 0.005920

T-test for Dependent Samples (Sporob KW Growth) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct9 Height 14.28131 8.82174

Nov9 Height 32.80303 33.29076 99 -18.5217 33.96488 -5.42586 98 0.000000

T-test for Dependent Samples (Sporob KW Growth) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar10 Height 15.92308 9.04788

Jun10 Height 28.32051 25.84192 78 -12.3974 27.37775 -3.99928 77 0.000145

T-test for Dependent Samples (Sporob Seed) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

EK % 54.66667 18.00926

KW % 83.26667 6.04676 3 -28.6000 20.62111 -2.40223 2 0.138244

T-test for Dependent Samples (Sporob t-test) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

KWApr9 54.00000 19.81161

EKJan10 90.00000 11.72604 5 -36.0000 19.49359 -4.12948 4 0.014500

207

T-test for Dependent Samples (Sporob t-test) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

KWApr9 63.33333 15.27525

EKFeb10 78.33333 18.92969 3 -15.0000 5.000000 -5.19615 2 0.035099

T-test for Dependent Samples (Sporob t-test) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

KWJun9 37.00000 24.13504

EKJan10 90.00000 11.72604 5 -53.0000 16.04681 -7.38537 4 0.001792

T-test for Independent Samples (Sporob Seed EKvsKW) Note: Variables were treated as independent samples Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

EK Seed 2009 vs. 54.66667 85.16667 -3.53537 22 0.001859 15 9 23.41143 13.86092 2.852805 0.140006 KW seed 2010

8.2.3 Sarcocornia decumbens East Kleinemonde Estuary

T-test for Dependent Samples (EK Sdec Cover) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

09-Jul Cover 24.00000 32.30067

09-Aug Cover 0.00000 0.00000 10 24.00000 32.30067 2.349631 9 0.043331

T-test for Dependent Samples (EK Sdec Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

09-Feb Height 27.02500 4.003967

09-Jul Height 39.28333 4.209711 6 -12.2583 4.844937 -6.19753 5 0.001596

T-test for Independent Samples (EK Sdec) Note: Variables were treated as independent samples Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Jul-09 cover vs. 28.00000 5.705882 2.993009 25 0.006140 10 17 30.11091 5.987119 25.29368 0.000000 Jun-10 cover

Kowie Estuary

T-test for Dependent Samples (Sdec KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May9 Cover 55.10000 18.62764

Jul9 Cover 62.00000 21.28641 10 -6.90000 8.198238 -2.66151 9 0.025981

T-test for Dependent Samples (Sdec 2 KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov09 Live Cover 60.80000 20.78354

Apr09 Live Cover 72.00000 18.28782 10 -11.2000 9.886017 -3.58259 9 0.005907

T-test for Dependent Samples (Sdec 2 KW) Marked differences are significant at p < .05000

208 Mean Std.Dv. N Diff. Std.Dv. t df p

Jun09 Height 21.45778 3.250012

Jul09 Height 24.72500 3.316059 10 -3.26722 1.952302 -5.29214 9 0.000499

T-test for Dependent Samples (Sdec KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jul09 Height 24.72500 3.316059

Aug09 Height 21.91500 3.488079 10 2.810000 2.747605 3.234089 9 0.010255

T-test for Dependent Samples (SdecKW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Height 25.76000 3.734732

Dec9 Height 29.51000 3.848939 10 -3.75000 3.174289 -3.73581 9 0.004656

T-test for Dependent Samples (SdecKW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May10 Height 22.90000 1.577621

Jun9 Height 26.09000 3.897136 10 -3.19000 3.444625 -2.92852 9 0.016797

T-test for Independent Samples (Sdec KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Summer vs. Autumn 28.13333 23.65000 3.523587 5 0.016855 3 4 2.030599 1.369915 2.197158 0.516850

T-test for Independent Samples (Sdec 2 KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Summer vs. Winter 28.13333 23.55000 2.798180 5 0.038075 3 4 2.030599 2.217356 1.192401 0.972645

East Kleinemonde Estuary compared to the Kowie Estuary

T-test for Dependent Samples (EK KW Sdec) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

EK Cover Jun9 36.30000 25.12657

KW Cover Apr10 72.00000 18.28782 10 -35.7000 27.65683 -4.08193 9 0.002750

T-test for Independent Samples (EK KW Sdec) Note: Variables were treated as independent samples

Valid Valid Std.Dev Std.Dev Mean Mean t-value df p F-ratio p N N . .

EK Cover Jun10 vs. KW Cover 6.06250 72.0000 - 2 0.00000 5.99409 18.2878 9.30843 0.00021 16 10 Apr10 0 0 13.4512 4 0 4 2 9 5

209

East Kleinemonde Estuary Seed number

T-test for Dependent Samples (Sdec) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar-09 0.00 0.00

Apr-09 21351.86 26569.53 10 -21351.9 26569.53 -2.54128 9 0.031644

T-test for Dependent Samples (Sdec) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb-09 102.88 325.32

Apr-09 21351.86 26569.53 10 -21249.0 26577.03 -2.52832 9 0.032324

Kowie Estuary Inflorescence number

T-test for Dependent Samples (Sdec2) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jul-09 241.6000 177.6884

Aug-09 66.3333 46.7206 10 175.2667 189.0738 2.931352 9 0.016720

T-test for Dependent Samples (Sdec2 #) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun2009 Sdec2 346.609 209.732

May2009 Sdec1 2188.129 2194.176 10 -1841.52 2183.839 -2.66659 9 0.025766

T-test for Dependent Samples (Sdec1&2 #) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun2009 Sdec2 346.609 209.732

June2009 Sdec1 1583.591 1672.043 10 -1236.98 1620.611 -2.41371 9 0.039011

Seed number

T-test for Dependent Samples (Shybrid and Sdec KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

April 2009 KW SHybrid 220365.2 230901.6

June 2009 KW Sdec 48576.3 28913.9 10 171788.9 227254.1 2.390471 9 0.040526

East Kleinemonde Estuary compared to the Kowie Estuary

T-test for Independent Samples (EK vs KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Sdec EK Cover vs. Sdec KW 6.062500 65.30000 -12.1564 24 0.000000 16 10 5.994094 18.16009 9.178863 0.000233

210

Seed viability T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

April 09EK 99.00000 2.23607

May 09EK 92.00000 17.88854 5 7.000000 18.57418 0.842701 4 0.446836

T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

May 09EK 92.00000 17.88854

June 09EK 98.00000 2.73861 5 -6.00000 19.17029 -0.699854 4 0.522582

T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

June 09EK 98.00000 2.738613

July 09EK 89.00000 9.617692 5 9.000000 9.617692 2.092457 4 0.104540

T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

July 09EK 89.00000 9.61769

Aug 09EK 94.00000 10.83974 5 -5.00000 12.74755 -0.877058 4 0.429973

T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Aug 09EK 94.00000 10.83974

Sep 09EK 87.00000 9.74679 5 7.000000 14.40486 1.086611 4 0.338307

T-test for Dependent Samples (Sdec seeds EK t-tests) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Sep 09EK 87.00000 9.746794

Oct 09EK 89.00000 8.944272 5 -2.00000 7.582875 -0.589768 4 0.587050

T-test for Dependent Samples (Sdec seed viability) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

April10 KW 9.00000 6.51920

May10KW 47.00000 10.95445 5 -38.0000 7.582875 -11.2056 4 0.000361

T-test for Dependent Samples (Sdec seed viability) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May10KW 47.00000 10.95445

June10KW 21.00000 12.44990 5 26.00000 13.87444 4.190279 4 0.013803

Mean monthly germination of S. decumbens seeds in East Kleinemonde Estuary (2009) compared to the Kowie Estuary (2010) T-test for Independent Samples (Sdec mean monthly germination) Note: Variables were treated as independent samples d Valid Valid Mean Mean t-value p Std.Dev. Std.Dev. F-ratio p f N N Sdec EK 2009 vs. Sdec KW 92.5714 25.6666 9.22129 0.00001 4.64962 19.4250 17.4537 0.00631 8 7 3 2010 3 7 6 5 9 7 4 1

211 T-test for Dependent Samples (Sdec mean monthly germination) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Sdec EK 2009 96.33333 3.78594

Sdec KW 2010 25.66667 19.42507 3 70.66667 23.15887 5.285156 2 0.033986

8.2.4 Sarcocornia hybrid Kowie Estuary

T-test for Dependent Samples (Kowie Shybrid) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb-09 37.80000 27.67992

Aug-09 15.40000 15.79170 10 22.40000 20.36446 3.478365 9 0.006956

T-test for Dependent Samples (Kowie Shybrid – Live cover) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

09-May 24.10000 20.70132

09-Jul 14.20000 13.21447 10 9.900000 10.27889 3.045713 9 0.013889

T-test for Dependent Samples (Kowie Shybrid – Live cover) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

09-May 24.10000 20.70132

09-Aug 15.40000 15.79170 10 8.700000 10.82230 2.542141 9 0.031600

T-test for Dependent Samples (Kowie Shybrid Live cover) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

09-Aug 15.40000 15.79170

09-Sep 21.70000 22.97849 10 -6.30000 8.743887 -2.27843 9 0.048687

T-test for Dependent Samples (Kowie Shybrid) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov2009 Live Cover 60.80000 20.78354

June2010 Live Cover 68.30000 18.67887 10 -7.50000 10.43765 -2.27226 9 0.049181

T-test for Dependent Samples (Kowie Shybrid Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Height 11.17500 3.896598

Mar9 Height 8.28000 1.251710 10 2.895000 3.720174 2.460851 9 0.036108

T-test for Dependent Samples (Kowie Shybrid Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Height 15.40123 6.103493

212 Mar10 Height 13.39506 4.372296 81 2.006173 5.868610 3.076632 80 0.002866

T-test for Dependent Samples (Kowie Shybrid Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Height 15.27326 5.950609

5.01162 6.05026 Apr10 Height 10.26163 3.609405 86 7.681631 85 0.000000 8 8

T-test for Dependent Samples (Kowie Shybrid Seedlings Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Height 9.06250 5.175170

7.79380 0.00001 Feb10 Height 14.14286 5.616303 56 -5.08036 -4.87796 55 7 0

Kowie Estuary: S.decumbens compared to the Sarcocornia hybrid

T-test for Independent Samples (EK vs KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Sdec KW vs. Shybrid KW 65.30000 11.18421 10.44800 27 0.000000 10 19 18.16009 9.937598 3.339437 0.028231

8.2.5 Salicornia meyeriana East Kleinemonde

T-test for Dependent Samples (Sali EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr9 Cover 5.954545 5.685308

May9 Cover 1.636364 2.975659 11 4.318182 5.273864 2.715616 10 0.021723

T-test for Dependent Samples (Sali EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Height 17.09302 6.366071

Mar9 Height 22.09302 5.502441 43 -5.00000 6.930780 -4.73066 42 0.000025

Kowie Estuary

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May9 Cover 0.609091 0.470010

Jun9 Cover 2.909091 2.416797 11 -2.30000 1.959592 -3.89277 10 0.002996

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun9 Cover 2.909091 2.416797

213 Jul9 Cover 9.363636 7.579878 11 -6.45455 5.968478 -3.58673 10 0.004956

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Cover 9.81818 6.73525

Dec9 Cover 15.90909 11.36182 11 -6.09091 7.354652 -2.74673 10 0.020593

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May10 Cover 7.81818 7.56066

Jun10 Cover 13.36364 11.47408 11 -5.54545 5.698485 -3.22756 10 0.009058

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May9Height 0.700000 0.421637

Jun9 Height 2.020000 1.130192 10 -1.32000 0.854790 -4.88331 9 0.000867

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jul9 Height 2.333278 1.742543

Aug9 Height 3.884000 2.811102 10 -1.55072 1.201633 -4.08096 9 0.002754

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Sep9Height 5.800000 3.983996

Oct9 Height 7.746111 4.892041 10 -1.94611 1.837830 -3.34859 9 0.008543

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct9 Height 7.682828 4.645741

Nov9 Height 9.640909 5.522989 11 -1.95808 1.304691 -4.97759 10 0.000555

T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Height 9.64091 5.522989

Dec9 Height 12.73636 6.976428 11 -3.09545 1.828449 -5.61485 10 0.000223

214 T-test for Dependent Samples (Sali KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May10 Height 1.725455 1.244720

Jun10 Height 4.531818 2.745293 11 -2.80636 1.715647 -5.42516 10 0.000291

East Kleinemonde Estuary compared to the Kowie Estuary

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

EK Cover Mar9 vs. KW Cover 7.00000 22.3000 - 2 0.00665 7.07106 15.7624 4.96911 0.01529 12 10 Mar9 0 0 3.02752 0 0 8 7 1 4

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Std.Dev Std.Dev Mean Mean t-value df p F-ratio p N N . .

EK Cover May10 vs. KW Cover 5.68181 22.3000 - 3 0.00049 8.45935 15.7624 3.47196 0.01794 22 10 Mar10 8 0 3.90310 0 8 0 7 0 1

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

EK Inf Apr9 vs. KW Inf Mar9 449.4426 2508.089 -2.13948 21 0.044299 12 11 1514.011 2938.910 3.768029 0.039641

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

EK Inf Apr9 vs. KW Inf May9 449.4426 4007.798 -4.35440 21 0.000278 12 11 1514.011 2350.939 2.411152 0.164852

T-test for Dependent Samples (Sali EKvsKW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t

Sali KW May09 Inf 3682.250 2509.392

Sali EK Apr09 Inf 449.500 1513.993 12 3232.750 3351.925 3.340939

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

EK Inf May10 vs. KW Inf 433.250 6275.45 - 3 0.00004 841.475 5743.04 46.5801 0.00000 22 11 Mar10 0 5 4.74450 1 5 8 4 7 0

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

EK Inf Jun10 vs. KW Inf 552.521 6275.45 - 0.00004 898.567 5743.04 40.8491 0.00000 32 23 11 Mar10 7 5 4.73676 3 9 4 0 0

215 T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

EK Seed Apr9 vs. KW Seed 24049.6 264223. - 2 0.01745 80998.1 311787. 14.8172 0.00010 12 11 Mar9 4 9 2.58018 1 7 7 4 2 7

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Std.Dev Std.Dev Mean Mean t-value df p F-ratio p N N . .

EK Seed May10 vs. KW Seed 12565.2 640292. - 3 0.00003 50728.2 615286. 147.114 0.00000 22 11 Mar10 0 3 4.83000 1 5 1 5 6 0

T-test for Dependent Samples (Sali EKvsKW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Sali KW Mar10 Seed 640292.3 615286.5

Sali EK Jun10 Seed 1155.3 1005.2 11 639137.0 615324.8 3.444973 10 0.006279

T-test for Independent Samples (Sali EKvsKW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

EK Seed Jun10 vs. KW Seed 3276.94 640292. - 3 0.00001 5721.44 615286. 11564.9 0.0 23 11 Mar10 5 3 5.05158 2 7 4 5 4 0

Seed viability

T-test for Dependent Samples (Sali Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May09 70.00000 22.36068

Jun09 27.00000 24.64752 5 43.00000 32.51923 2.956741 4 0.041689

T-test for Dependent Samples (Sali Seed) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

May-10 80.00000 11.18034

Jun-10 29.00000 14.74788 5 51.00000 9.617692 11.85726 4 0.000290

8.2.6 Sarcocornia tegetaria East Kleinemonde Estuary

T-test for Dependent Samples (Steg EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Cover 23.54545 18.62989

Dec10 Cover 2.27273 3.43776 11 21.27273 18.80474 3.751908 10 0.003772

216 T-test for Dependent Samples (Steg EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Height 8.66250 1.725802

Mar9 Height 14.01528 6.149871 8 -5.35278 6.078578 -2.49070 7 0.041554

T-test for Dependent Samples (Steg Height EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar9 Height 13.39865 6.374166

Apr9 Height 8.14189 2.687516 74 5.256757 7.706626 5.867721 73 0.000000

T-test for Dependent Samples (Steg EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb10 Height 1.101250 1.645823

Mar10 Height 6.378472 4.797293 24 -5.27722 4.241083 -6.09585 23 0.000003

Kowie Estuary

T-test for Dependent Samples (Steg KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr9 Cover 50.30000 28.35117

May9 Cover 58.40000 26.61328 10 -8.10000 8.279157 -3.09385 9 0.012849

T-test for Dependent Samples (Steg KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Height 9.02889 3.606563

Jul9 Height 16.27000 7.135794 10 -7.24111 7.953254 -2.87912 9 0.018204

T-test for Dependent Samples (Steg KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun9 Height 12.27000 4.855364

Jul9 Height 16.27000 7.135794 10 -4.00000 4.580636 -2.76143 9 0.022060

East Kleinemonde Estuary versus the Kowie Estuary

T-test for Independent Samples (EK vs KW) Note: Variables were treated as independent samples

Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Steg EK Cover vs. Steg Kw 14.95455 46.50000 -4.12970 41 0.000174 33 10 15.88944 33.79760 4.524331 0.001362

T-test for Independent Samples (Steg EKvsKW) Note: Variables were treated as independent samples

Valid Valid Std.Dev Std.Dev Mean Mean t-value df p F-ratio p N N . .

217 Jun10 Seed EK vs. May10 Seed 12159.6 2388.05 2.20420 3 0.03398 18655.6 5037.15 13.7166 0.00000 19 19 KW 3 3 3 6 9 1 6 6 1

Seed viability East Kleinemonde T-test for Independent Samples (Steg seed2) Note: Variables were treated as independent samples Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

May-09 vs. Jun-09 66.00000 37.50000 3.417040 9 0.007664 5 6 14.74788 12.94218 1.298507 0.766198

8.2.7 Phragmites australis East Kleinemonde Estuary

T-test for Dependent Samples (EK Phrag) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Cover 17.07000 10.76332

Aug9 Cover 3.60000 6.20394 10 13.47000 13.36696 3.186654 9 0.011066

T-test for Dependent Samples (Phrag EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr9 Cover 15.40000 11.99259

May9 Cover 10.50000 8.75912 10 4.900000 6.136412 2.525117 9 0.032495

T-test for Dependent Samples (Phrag EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar9 Height 72.08750 22.46905

Apr9 Height 90.98125 23.47167 8 -18.8938 16.08285 -3.32277 7 0.012717

T-test for Dependent Samples (Phrag EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun9 Height 93.9188 20.04587

Jul9 Height 117.7177 23.30908 8 -23.7990 12.99750 -5.17896 7 0.001282

T-test for Dependent Samples (Phrag EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar10 Height 88.1300 47.88790

Jun10 Height 108.2700 61.14424 10 -20.1400 27.02054 -2.35703 9 0.042809

Kowie Estuary

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

218 Oct9 Cover 8.80000 9.93087

Nov9 Cover 14.10000 14.34070 10 -5.30000 6.037844 -2.77584 9 0.021547

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Cover 14.10000 14.34070

Dec9 Cover 19.40000 16.48029 10 -5.30000 4.989990 -3.35874 9 0.008406

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr10 Cover 20.10000 18.12580

May9 Cover 11.60000 10.99697 10 8.500000 8.618456 3.118814 9 0.012342

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer Cover 19.15000 4.547160

Winter Cover 12.57500 3.705289 4 6.575000 4.031026 3.262197 3 0.047053

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Autumn Cover 14.12500 3.326034

Winter Cover 12.57500 3.705289 4 1.550000 0.785281 3.947630 3 0.028990

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct9 Height Ph 79.06111 41.97202

Nov9 Height Ph 99.58889 40.85543 90 -20.5278 48.40259 -4.02341 89 0.000120

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Height Ph 99.5889 40.85543

Dec9 Height Ph 114.4111 42.24431 90 -14.8222 52.43285 -2.68183 89 0.008727

T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Height Ph 115.7222 43.22271

Feb10 Height Ph 137.0889 41.94797 90 -21.3667 52.40217 -3.86820 89 0.000208

219 T-test for Dependent Samples (Phrag Kw) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar10 Height Ph 142.8750 47.63665

Apr10 Height Ph 153.5375 39.42304 80 -10.6625 47.74217 -1.99757 79 0.049206

8.2.8 Bolboschoenus maritimus East Kleinemonde Estuary

T-test for Dependent Samples (Bolbo EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jul9 Cover 19.00000 16.79947

Aug9 Cover 0.00000 0.00000 10 19.00000 16.79947 3.576498 9 0.005963

T-test for Dependent Samples (Bolbo EK) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jun9 Height 60.04222 21.83149

Jul9 Height 85.98500 4.18742 10 -25.9428 23.69973 -3.46157 9 0.007143

Kowie Estuary

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Feb9 Cover 38.50000 21.99116

Mar9 Cover 10.00000 6.30696 10 28.50000 19.91789 4.524823 9 0.001437

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar9 Cover 10.00000 6.306963

Apr9 Cover 3.50000 1.080123 10 6.500000 5.359312 3.835344 9 0.003995

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct9 Cover 7.40000 3.56526

Nov9 Cover 26.30000 13.21657 10 -18.9000 11.87387 -5.03349 9 0.000706

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Nov9 Cover 26.30000 13.21657

Dec9 Cover 40.00000 24.26703 10 -13.7000 13.21657 -3.27795 9 0.009561

220

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar10 Cover 15.60000 13.04011

Apr10 Cover 6.20000 4.07704 10 9.400000 9.868018 3.012298 9 0.014661

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer Cover 32.77500 8.828127

Autumn Cover 9.20000 5.044469 4 23.57500 13.37744 3.524591 3 0.038789

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer Cover 32.77500 8.828127

Winter Cover 7.00000 2.437212 4 25.77500 7.879245 6.542505 3 0.007259

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Aug9 Height 56.22900 9.221441

Sep9 Height 65.82639 5.665947 10 -9.59739 10.65009 -2.84971 9 0.019098

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer Height 80.22500 5.786406

Autumn Height 74.82500 6.027368 4 5.400000 1.838478 5.874426 3 0.009841

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Summer Height 80.22500 5.786406

Winter Height 64.72500 6.142407 4 15.50000 4.084932 7.588866 3 0.004747

T-test for Dependent Samples (Bolbo KW) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Mar9 KW Seed 3294.632 3086.904

Jan10 KW Seed 199.600 258.668 10 3095.032 2874.854 3.404469 9 0.007818

East Kleinemonde Estuary compared to the Kowie Estuary

221 T-test for Independent Samples (EK vs KW) Note: Variables were treated as independent samples

Valid Valid Mean Mean t-value df p Std.Dev. Std.Dev. F-ratio p N N

Bolbo EK Cover vs. Bolbo 0.10000 4.40000 - 0.00096 0.31622 3.43834 118.222 0.00000 18 10 10 KW 0 0 3.93813 4 8 6 2 0

Seed viability T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Mar EK 0.000000 0.000000

April EK 4.000000 8.944272 5 -4.00000 8.944272 -1.00000 4 0.373901

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

April EK 4.000000 8.944272

May09 EK 2.000000 4.472136 5 2.000000 10.95445 0.408248 4 0.704000

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

May09 EK 2.00000 4.47214

Jun09 EK 11.00000 10.83974 5 -9.00000 10.83974 -1.85656 4 0.136945

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Jun09 EK 11.00000 10.83974

Jul09 EK 0.00000 0.00000 5 11.00000 10.83974 2.269127 4 0.085810

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Jul09 EK 0.000000 0.000000

Aug09 EK 3.000000 6.708204 5 -3.00000 6.708204 -1.00000 4 0.373901

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Aug09 EK 3.00000 6.70820

Sep09 EK 36.00000 28.15138 5 -33.0000 28.41654 -2.59674 4 0.060255

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Sep09 EK 36.00000 28.15138

Oct09 EK 15.00000 21.21320 5 21.00000 25.83602 1.817518 4 0.143290

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Oct09 EK 15.00000 21.21320

Nov09 EK 13.00000 19.87461 5 2.000000 28.85308 0.154997 4 0.884331

222 T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Nov09 EK 13.00000 19.87461

Jan10KW 46.00000 17.81853 5 -33.0000 30.53686 -2.41643 4 0.073044

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

May10 EK 5.00000 5.000000

Jun10 EK 11.00000 5.477226 5 -6.00000 5.477226 -2.44949 4 0.070484

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Mar09 EK 31.00000 11.93734

Jan10KW 46.00000 17.81853 5 -15.0000 14.57738 -2.30089 4 0.082857

T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10KW 57.50000 3.535534

May10KW 26.40000 1.979899 2 31.10000 1.555635 28.27273 1 0.022508

Seed viability of East Kleinemonde seeds in 2010 compared to Kowie Estuary seeds in 2010 T-test for Dependent Samples (Bolbo t-tests all months) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10KW 46.00000 17.81853

May10 EK 5.00000 5.00000 5 41.00000 21.62175 4.240119 4 0.013262

T-test for Independent Samples (Bolbo t-tests all months) Note: Variables were treated as independent samples Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

Jun10 EK vs. May10KW 11.00000 26.40000 -3.69731 5 0.014039 5 2 5.477226 1.979899 7.653061 0.527946

T-test for Independent Samples (Bolbo t-tests all months) Note: Variables were treated as independent samples Mean Mean t-value df p Valid N Valid N Std.Dev. Std.Dev. F-ratio p

May10 EK vs. May10KW 5.000000 26.40000 -5.61048 5 0.002488 5 2 5.000000 1.979899 6.377551 0.575333

8.2.9 Ruppia cirrhosa and Chara vulgaris Ruppia cirhhosa

T-test for Dependent Samples (Ruppia) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Apr9 Biomass 0.020100 0.008373

May9 Biomass 0.315200 0.241889 10 -0.295100 0.242642 -3.84595 9 0.003931

T-test for Dependent Samples (Ruppia) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Aug9 Biomass 0.805200 0.559750

223 Sep9 Biomass 6.051900 2.826015 10 -5.24670 3.187268 -5.20556 9 0.000560

T-test for Dependent Samples (Ruppia) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Sep9 Biomass 6.05190 2.826015

Oct9 Biomass 11.09600 6.346206 10 -5.04410 4.699696 -3.39402 9 0.007948

T-test for Dependent Samples (Ruppia) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Oct9 Biomass 11.09600 6.346206

Nov9 Biomass 18.48450 3.546660 10 -7.38850 6.855734 -3.40802 9 0.007774

T-test for Dependent Samples (Ruppia) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Biomass 11.35690 4.770248

Feb10 Biomass 4.45300 2.779545 10 6.903900 4.618457 4.727130 9 0.001078

Reproductive output T-test for Dependent Samples (Ruppia fl and seeds) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Sep9 Fl/m2 277.20 508.433

Oct9 Fl/m2 10880.10 8814.028 10 -10602.9 8546.965 -3.92295 9 0.003495

T-test for Dependent Samples (Ruppia fl and seeds) Marked differences are significant at p < .05000 Mean Std.Dv. N Diff. Std.Dv. t df p

Sep9 Seed 69.300 219.146

Oct9 Seed 6237.300 3866.440 10 -6168.00 3744.204 -5.20937 9 0.000557

Chara vulgaris

T-test for Dependent Samples (Chara) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Sep9 Biomass 0.199900 0.152494

Oct9 Biomass 0.764200 0.343933 10 -0.564300 0.423325 -4.21537 9 0.002255

T-test for Dependent Samples (Ruppia Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Aug9 Height 22.96700 8.04624

Sep9 Height 65.38974 11.75965 10 -42.4227 19.05981 -7.03850 9 0.000061

T-test for Dependent Samples (Ruppia Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

224 Sep9 Height 65.38974 11.75965

Oct9 Height 74.61289 11.97525 10 -9.22316 6.841312 -4.26324 9 0.002101

T-test for Dependent Samples (Ruppia Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Dec9 Height 82.94544 9.74081

Jan10 Height 62.58000 25.38266 10 20.36544 21.93958 2.935388 9 0.016611

T-test for Dependent Samples (Ruppia Height) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t df p

Jan10 Height 62.58000 25.38266

Feb10 Height 32.55750 12.05560 10 30.02250 18.01085 5.271239 9 0.000513

T-test for Dependent Samples (Chara) Marked differences are significant at p < .05000

Mean Std.Dv. N Diff. Std.Dv. t

Jan10 Biomass 0.579000 0.184723

Feb10 Biomass 0.364900 0.181573 10 0.214100 0.286654 2.361887

225