The Botanical Importance and Health of the Bushmans Estuary, Eastern Cape, South Africa

THE BOTANICAL IMPORTANCE AND HEALTH OF THE BUSHMANS ESTUARY, EASTERN CAPE, SOUTH AFRICA

By Nolusindiso Jafta

Submitted in fulfilment of the requirements for the degree of

MAGISTER SCIENTIAE

In the Faculty of Science at Nelson Mandela Metropolitan University

January 2010

Supervisor: Prof. J.B. Adams

Abstract

The Bushmans Estuary is one of the few permanently open estuaries in the Eastern Cape that is characterized by large intertidal salt marshes. Freshwater inflow to the estuary has decreased as a result of abstraction by more than 30 weirs and farm dams in the catchment. The mean annual run-off is naturally low (38 x 106 m3 y-1) and thus abstraction and reduction of freshwater inflow to the estuary is expected to cause a number of changes. The aims of this study were to determine the current health/status of the estuary based on the macrophytes and microalgae and identify monitoring indicators for the East London Department of Water Affairs, River Health Programme. Changes in the estuary over time were determined from available historical data which were compared with present data. This analysis showed that under normal average conditions freshwater inflow to the estuary is very low, less that 0.02 m3 s-1 most of the time. Under these conditions the estuary is in a homogenous marine state. Vertical and horizontal salinity gradients only form when high rainfall and run-off occurs (> 5 m3.s-1). Salinity gradients from 30.1 PSU at the mouth to 2.2 PSU in the upper reaches were measured in 2006 after a high flow event. However the estuary quickly reverted back to its homogenous condition within weeks after this flood.

This study showed that freshwater inflow increased nutrient input to the estuary. Total oxidised nitrogen (TOxN) and soluble reactive phosphorus (SRP) concentrations were higher in August 2006, after the flood, than during the other low flow sampling sessions. TOxN decreased from a mean concentration of 21.6 µM in 2006 to 1.93 µM in February 2009. SRP decreased from 55.3 µM to 0.2 µM respectively. With the increased nutrient availability, the response in the estuary was an increase in phytoplankton biomass. After the 2006 floods the average water column chlorophyll-a was 9.0 µg l-1, while in the low freshwater inflow years it ranged from 2.1 to 4.8 µg l-1. The composition of the phytoplankton community was always dominated by flagellates and then diatoms, with higher cell numbers in the nutrient- enriched 2006 period.

Although the water column nutrient data indicated that the estuary was oligotrophic, benthic microalgal biomass (11.9-16.1 µg.g-1) in the intertidal zone was comparable with nutrient rich estuaries. Benthic species indicative of polluted conditions were found (Nitzschia frustulum, Navicula gregaria, Navicula cryptotenelloides). These benthic species were found at the sites where wastewater / sewage seepage had occurred. Benthic diatom species also indicated freshwater inflow. During the high flow period in 2006 the dominant diatoms were fresh to brackish species that were strongly associated with the high concentrations of TOxN and SRP (Tryblionella constricta, Diploneis smithii, Hippodonta cf. gremainii, and Navicula species). During the freshwater limited period of 2008 and 2009 the benthic diatom species shifted to a group responding to the high salinity, ammonium and silicate concentrations. The species in this group were Nitzschia flexa, Navicula tenneloides, Diploneis elliptica, Amphora subacutiuscula and Nitzschia coarctata.

Ordination results showed that the epiphytic diatom species responded to different environmental variables in the different years. Most of the species in 2008/2009 were associated with high salinity, temperature, dissolved oxygen, ammonium and silicate concentrations while the response was towards TOxN and SRP in 2006. The dominant species were Cocconeis placentula v euglyphyta in 2006; Nitzschia frustulum in 2008; and Synedra spp in 2009. The average biomass of the epiphytes was significantly lower in May 2008 than in both August 2006 and February 2009; 88.0 + 17.7 mg.m-2, 1.7 + 0.8 mg.m-2, and 61.8 + 14.4 mg.m-2 respectively.

GIS mapping of past and present aerial photographs showed that submerged macrophyte (Zostera capensis) cover in 1966 and 1973 was less than that mapped for 2004. Salt marsh also increased its cover over time, from 86.9 ha in 1966 to 126 ha in 2004, colonizing what were bare sandy areas. Long-term monitoring of the health of the Bushmans Estuary should focus on salinity (as an indicator of inflow or deprivation of freshwater), benthic diatom identification and macrophyte distribution and composition (for the detection of pollution input), and bathymetric surveys (for shallowing of the estuary due to sedimentation).

Acknowledgements

I would like to extend my greatest and sincerest gratitude to my supervisor, Professor Janine Adams. I thank her very much for all her much needed pushes, inspiration and abundant help she has given to me to be able to reach my goal, which seemed impossible to achieve at times.

I would also like to thank: Dr. Gavin Snow for letting me disrupt his schedule at times with my sampling trips and input to my project; Ms. Shireen Prinsloo for all the help she provided during sampling trips; and Mrs Patricia Smailes, who contributed a significant amount of time assisting with the microalgal counts and diatom identification, and a whole lot of motivation.

Thank you to Mr. Mandilakhe Mdodana for providing data for the 2006 sampling trip and Mary-Anne Crocker for the May 2008 epiphyte data. And, thank you to all the postgraduate students and staff in the Botany Department at NMMU and my friends for their input and support during this journey.

I would also like to express my gratitude to the National Research Foundation and the River Health Programme of the East London Department of Water Affairs for funding this project.

My sincerest gratitude also goes to my family and Lukhanyo Mdyesha for their continuous love and support.

III

CONTENTS

Abstract ...... I Acknowledgements ...... III List of Figures ...... VI List of Tables ...... IX Chapter 1: Background ...... 1 1.1 General Introduction ...... 1 1.2 Available information on the Bushmans River catchment ...... 3 1.2.1 Climate ...... 3 1.2.2 Geology and geomorphology ...... 5 1.2.3 Uses and anthropogenic impacts ...... 5 1.2.4 Physical characteristics of the Estuary ...... 7 1.2.5 Water quality of the estuary ...... 9 1.2.6 The biota of the estuary...... 11 1.2.7 The importance of the estuary ...... 13

Chapter 2: Literature Review ...... 16 2.1 Botanical communities in estuaries and their importance ...... 16 2.1.1 Salt marsh...... 16 2.1.2 Submerged macrophytes ...... 20 2.1.3 Reeds and Sedges ...... 27 2.1.4 Phytoplankton ...... 30 2.1.5 Benthic microalgae ...... 34 2.1.6 Epiphytic microalgae ...... 41 2.2 The botanical importance of estuaries ...... 45 2.3 The health of estuaries ...... 46 2.2.1 The South African Estuarine Health Index ...... 47 2.2.2 Monitoring ...... 49 2.2.3 The South African monitoring protocols ...... 55

Chapter 3: The botanical importance and present status of the Bushmans Estuary ...... 62 3.1 Introduction ...... 62 3.2 Materials and Methods ...... 63 3.2.1 Assessment of the present status of the estuary ...... 63 3.2.1.1 Environmental factors ...... 64 3.2.1.2 Phytoplankton ...... 65 3.2.1.3 Benthic microalgae ...... 66 3.2.1.4 Epiphytic microalgae ...... 68 3.2.1.5 Salt marsh cover and distribution ...... 70 3.2.1.6 Zostera capensis biomass ...... 70 3.2.1.7 Human activities ...... 71 3.2.1.8 Statistical Analysis ...... 72 3.2.2 Change in the botanical importance of the estuary ...... 72 3.2.2.1 Assessment of habitat areas from aerial photographs ...... 72 3.2.2.2 The Botanical Importance Rating Index ...... 73 3.3 Results ...... 76 3.3.1 Assessment of the present status of the estuary ...... 76 3.3.1.1 Environmental variables ...... 76

3.3.1.2 Phytoplankton ...... 84 3.3.1.3 Benthic microalgae ...... 88 3.3.1.4 Epiphytic microalgae ...... 93 3.3.1.5 Current salt marsh cover and distribution ...... 99 3.3.2.6 Zostera capensis biomass ...... 103 3.3.1.8 Human activities ...... 103 3.3.2 Change in the botanical importance of the estuary ...... 107 3.3.2.1 Assessment of habitat areas from aerial photographs and changes over time ...... 107 3.3.2.2 The Botanical Importance Rating index ...... 115 3.4 Discussion ...... 116 3.4.1 Assessment of the present status of the estuary ...... 116 3.4.1.1 Phytoplankton ...... 119 3.4.1.2 Benthic microalgae ...... 121 3.4.1.3 Epiphytic microalgae ...... 125 3.4.1.4 Salt marsh cover and distribution ...... 126 3.4.1.5 Zostera capensis biomass ...... 128 3.4.1.6 Human activities ...... 130 3.4.2 Change in the botanical importance of the estuary ...... 130 3.5 Conclusion ...... 132

Chapter 4: Application of the Estuarine Health Index and identification of monitoring indicators for the Bushmans Estuary ...... 134 4.1 Introduction ...... 134 4.2 Materials and Methods ...... 135 4.2.1 Data Availability ...... 135 4.2.2 Estuarine Health Index ...... 135 4.3 Results ...... 135 4.3.1 Data Availability ...... 135 4.3.2 Estuarine Health Index ...... 139 4.4 Discussion ...... 144 4.4.1 Change in the health of the estuary ...... 144 4.4.2 Monitoring indicators for the Bushmans Estuary ...... 145 4.5 Conclusion ...... 146

Chapter 5: References ...... 147

APPENDICES ...... 165 Appendix A: List of diatom species found in Bushmans Estuary ...... 166 Appendix B: Images of the identified diatom species in the Bushmans Estuary ...... 172 Appendix C: Full names of the salt marsh species identified in the Bushmans Estuary ... 183

List of Figures

Figure 1.1: Location map and aerial photograph of the Bushmans Estuary ...... 3

Figure 2.1: The proposed monitoring protocol for South African estuaries incorporating an objectives hierarchy to identify operational goals and measure their achievement through the use of indicators (from Adams and McGwynne, 2004)...... 57

Figure 2.2: An estuary management protocol that incorporates the monitoring protocol proposed for South African estuaries (taken from Adams and McGywnne, 2004, modified from Rogers and Biggs, 1999)...... 58

Figure 2.3: Conceptual framework of the anticipated abiotic and biotic processes and interactions relevant to estuaries (from Taljaard et al., 2003) ...... 60

Figure 3.1: The location of the sampling sites along the estuary for: microalgae and physico- chemical (environmental) variables (M&E1-7; in red); Zostera capensis (Z1-9; in black); and salt marsh (Lower, Middle and Upper Estuary; in grey)...... 63

Figure 3.2: Regression comparison of the leaf length and biomass of Zostera capensis in 2009...... 71

Figure 3.3: a) Mean monthly flow from 2006 to 2008, with the arrows indicating the months in which sampling was done; b) Historical freshwater flow since 1957 recorded at Station P1H003 in Alicedale at the confluence of the Bushmans with the New Year‟s River...... 78

Figure 3.4: Clarity of the water column indicated by Secchi depth in August 2006, October 2008 and February 2009...... 79

Figure 3.5: Top and bottom water column salinity along the length of the estuary in August 2006, May, October 2008 and February 2009...... 79

Figure 3.6: Top and bottom water column temperature along the estuary in August 2006, May, October 2008 and February 2009...... 80

Figure 3.7: Top and bottom water column dissolved oxygen concentrations along the estuary in August 2006, May and October 2008. No data were available in 2009 as the meter was faulty...... 80

Figure 3.8: The distribution and composition of sediment particle types along the intertidal and subtidal areas of the estuary in August 2006, October 2008 and February 2009...... 81

Figure 3.9: The average nutrient concentrations (+ standard error) along the estuary in August 2006, May and October 2008, and February 2009 for ammonium (a), total oxidised nitrogen (b), soluble reactive phosphorus (c) and silicate (d)...... 83

Figure 3.10: The top and bottom phytoplankton chlorophyll-a concentrations (average + standard error) along the estuary in August 2006, May and October 2008, and February 2009...... 86

Figure 3.11: The composition of the phytoplankton community, with a – d being the relative abundance of the phytoplankton community groups for August 2006, May and October 2008, and February 2009, and e – h are the cell numbers of the phytoplankton groups in the estuary in 2006, 2008 and 2009...... 87

Figure 3.12: The average intertidal and subtidal benthic chlorophyll-a along the estuary in August 2006, May, October 2008 and February 2009 (average + S.E)...... 90

Figure 3.13: Ordination (Canonical Correspondence Analysis) of the benthic diatom species found in the Bushmans Estuary in August 2006, October 2008 and February 2009 (Example of abbreviated names: 061S = Subtidal Site 1 sampled in 2006). The full names for the epiphytic diatom species are in Appendix A. The following codes for the environmental variables stand for: MPB_Chla = benthic microalgal chlorophyll- a (biomass); Temp = temperature; NH4 = ammonium; Cond = electrical conductivity; Si2+ = silicate; D.O. = dissolved oxygen; TOxN = total oxidised nitrogen; SRP = soluble reactive phosphorus...... 92

Figure 3.14: The chlorophyll-a (biomass) of the epiphytic microalgae along the estuary in August 2006, May 2008 and February 2009 (Average + S.E. bars)...... 94

Figure 3.15: The organic content of the epiphytic microalgae along the estuary in August 2006, May 2008 and February 2009 (Average + S.E. bars)...... 95

Figure 3.16: Ordination (Canonical Correspondence Analysis) of the epiphytic diatom species and water column environmental variables in August 2006, May 2008 and February 2009 (Example of abbreviated names: S108 = Site 1 sampled in 2008). The full names for the epiphytic diatom species are in Appendix A. The following codes for the environmental variables stand for: Epi_chla = epiphytic microalgal chlorophyll-a (biomass); Org_cont = epiphytic organic content; Secchi = secchi depth; Temp = temperature; NH4 = ammonium; Si2+ = silicate; D.O. = dissolved oxygen; TOxN = total oxidised nitrogen; SRP = soluble reactive phosphorus...... 96

Figure 3.17: Ordination (Canonical Correspondence Annalysis) of the benthic and epiphytic diatom species found in the estuary in: a) 2006; b) 2008; and c) 2009 (Example of abbreviated names: 06E1 = Epiphytes Site 1 in 2006; 08B1I = Benthic Site 1 Intertidal in 2008). The full names for the epiphytic diatom species are in Appendix A. The following codes for the environmental variables stand for: MPB_Chla = benthic microalgal chlorophyll-a (biomass); Epi_chla = epiphytic microalgal chlorophyll-a (biomass); Org_cont = epiphytic organic content; Secchi = secchi depth; Cond = electrical conductivity; Temp = temperature; NH4 = ammonium; Si2+ = silicate; D.O. = dissolved oxygen; TOxN = total oxidised nitrogen; SRP = soluble reactive phosphorus...... 98

Figure 3.18: Dominant salt marsh species in the lower, middle and upper reaches of the Bushmans Estuary in 1995 and 2008. Full species names can be found in Appendix C...... 100

VII

Figure 3.19: Areas in the lower reaches of the estuary that had stands of freshwater reeds and sedges...... 101

Figure 3.20: The reeds and sedges communities along the estuary in the upper reaches. a) The unhealthy pure stands of P. australis; b) Mixed stands of the reed and sedge, B. maritimus ...... 102

Figure 3.21: Z. capensis biomass along the estuary in October 2008 and February 2009 measured using the leaf length method...... 105

Figure 3.22: Z. capensis biomass along the estuary in February 2009 comparing the leaf length method and the dry weight biomass method...... 105

Figure 3.23: Anthropogenic impacts in the Bushmans Estuary (Pictures taken by J.B. Adams). a) Jetty, b) The green circle highlights a reed community in the mouth area that has formed due to a leaking drain, c) A restaurant floating on the estuary disturbing the salt marsh area, d) Bait collection, e) Road on supratidal salt marsh area, f) Housing development below the 5 m contour line...... 106

Figure 3.24a: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (1942 and 1966)...... 109

Figure 3.24b: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (1973 and 1990s)...... 110

Figure 3.24c: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (2004)...... 111

Figure 3.25a: Map illustrating the changes along the length of the estuary in 1966...... 112

Figure 3.25b: Map illustrating the changes along the length of the estuary in 1973 ...... 113

Figure 3.25c: Map illustrating the changes along the length of the estuary in 2004...... 114

Figure 3.26: Evidence of agricultural activity near the estuary and abandoned land that has become degraded...... 116

Figure 3.27: Gully/valley formations towards the banks of the estuary...... 119

VIII

List of Tables

Table 1.1: Registered farm dams in the Bushmans River Catchment indicating surface area, capacity and year construction completed (Albany Coast Situation Assessment report – DWAF, 2004)...... 9

Table 1.2: List of the common birds counted along the Bushmans Estuary (that would be birds spotted more than 10 times during the 17 counts that were conducted) (retrieved from the CWAC website)...... 14

Table 1.3: List of birds spotted in the vicinity of the Bushmans Estuary ...... 15

Table 2.1: The comparison of phytoplankton and microphytobenthos biomass (as kg chlorphyll-a/estuary) for selected South African systems (adapted from Rodriguez, 1993) ...... 35

Table 2.2: Calculation of the Estuarine Health Score ...... 48

Table 2.3: Ecological Management Categories ...... 49

Table 3.1: The productivity values for the different plant community types from Colloty et al. (1999) ...... 73

Table 3.2: The scoring system for species richness based on the number of species present . 74

Table 3.3: The community richness scores based on the number of communities present ..... 74

Table 3.4: The scoring system for plant community type rarity ...... 75

Table 3.5: Results of the correlation analysis of phytoplankton biomass with different environmental variables showing R-values (The bold R-values show significant correlations, i.e. the p value was less than 0.05) ...... 85

Table 3.6: Benthic diatom species list showing the common species for the different sampling trips and the dominant species (species with > 20% relative abundance)...... 91

Table 3.7: Summary of the CCA for the benthic diatom species against environmental variables found in the estuary in 2006, 2008 and 2009...... 92

Table 3.8: Epiphytic diatom species list showing the common species for the different sampling trips and the dominant species (species with > 5% relative abundance). The relative abundance is an average value for the different estuary sites in one year...... 95

Table 3.9: Summary of the CCA for the epiphytic diatom species against environmental variables found in the estuary in 2006, 2008 and 2009...... 96

Table 3.10: Anthropogenic activities recorded in the estuary during the sampling trips and the total number of the individuals that were participating in these activities, as well as

the number of jetties on the intertidal area as an estimate of the extent of development...... 104

Table 3.11: Change in the cover of the different habitats in the lower reaches of the Bushmans Estuary...... 108

Table 3.12: Overall change in the cover of the different habitat types from 1966 to 2004 along the entire length of the estuary...... 108

Table 3.14: Comparison of the measured benthic microalgal biomass in the Bushmans Estuary with that of other estuaries as measured by Snow (2007)...... 123

Table 4.1 Data availability on sediment dynamics, hydrodynamics and water quality ...... 136

Table 4.2 Data availability on microalgae ...... 137

Table 4.3 Data availability on macrophytes ...... 137

Table 4.4 Data availability on invertebrates ...... 138

Table 4.5 Data availability on fish ...... 139

Table 4.6 Data availability on birds ...... 139

Table 4.7: Hydrological health score ...... 139

Table 4.8: Mouth condition score ...... 140

Table 4.9: Water quality health score ...... 140

Table 4.10: Physical habitat health score ...... 141

Table 4.11: Biotic health score for microalgae ...... 141

Table 4.12: Biotic health score for macrophytes ...... 142

Table 4.13: Biotic health score for invertebrates ...... 142

Table 4.14: Biotic health score for fish ...... 143

Table 4.15: Biotic health score for birds ...... 143

Chapter 1: Background

1.1 General Introduction

The Bushmans Estuary is situated in the Eastern Cape Province and opens into the Indian Ocean at 33o42‟S; 26o40‟E (Figure 1.1), between the villages of Bushmans River mouth and Kenton-on-Sea. The head of the estuary is situated at Harvest Vale, some 33 km from the mouth. The catchment area has been estimated to be around 2700 km2 by Day (1981), 2670 km2 by Reddering and Esterhuysen (1981) while Bornman and Klages (2004) estimated an area of 2678 km2.

The estuary is classified as a large, permanently open system and is dominated by marine water with a tidal effect reaching as far as 40 km upstream. It is navigable for approximately 30 km (Bornman and Klages, 2004). It is one of the few large permanently open estuaries in the Eastern Cape that is characterized by large intertidal salt marshes. In terms of the conservation importance, the estuary has been rated as 41 out of 250 ranked estuaries in South Africa (Turpie and Clark, 2007). A preliminary study identified the ecological integrity of the Bushmans Estuary as fair, which means there is a noticeable degree of degradation in the catchment and/or estuary (Whitfield, 2000).

Mean annual runoff of the estuary catchment is low; 38 x 106 m3.a-1 (Reddering and Esterhuysen, 1981). Due to low freshwater input caused by several dams and weirs (approximately 30), salinity is usually high. Nutrient input to the estuary is probably low because of the reduced freshwater input. However herbicides, pesticides and fertilizers from the surrounding agricultural areas may be transported into the estuary during floods which would impact on the water quality of the estuary. Little information is available on the botanical characteristics of the estuary. The increase in salinity may have changed the distribution of organisms, for example a decrease in phytoplankton biomass (Hilmer and Bate, 1990) and increase in the spread of the submerged macrophyte Zostera capensis to the middle and upper reaches of the estuary (Adams and Talbot, 1992).

The estuary is dominated by flood tides, causing an accumulation of marine sand in the lower reaches (Reddering, 1988a). Reddering and Esterhuysen (1981) suggested that the sand should be dredged out of the estuary but there were a lot of questions that were raised about the feasibility of this at the time. Marine sedimentation in the lower reaches of the estuary has become an increasing problem and the local residents of Bushmans River Mouth and Kenton- on-Sea would like to have the sediment dredged out of the estuary. However, before this can be done an assessment of the present health of the estuary and the formulation of a long term monitoring plan is necessary.

The objectives of this study were to: Assess the present status of the Bushmans Estuary. o Determine changes over time in the species richness, abundance and community composition of the macrophytes and microalgae of the estuary; . Determine the biomass, distribution and species composition of the microalgae. The estuary was sampled during both low and high flow conditions; . Map the present and past distribution of the macrophytes using available aerial photographs, and assess the change in cover over time; Collate available information on the estuary and assemble according to the requirements of the Estuarine Health Index (as described in the DWA RDM Methodology for the Determination of the Ecological Water Requirements for Estuaries; Version 2, 2008a); Provide input to a long-term monitoring and sampling programme for the Department of Water Affairs, East London River Health programme, by identifying monitoring indicators.

The research tested the following hypotheses: Freshwater abstraction and an increase in salinity have increased the distribution and cover of the seagrass, Zostera capensis, which is usually found in the saline lower reaches. Lack of floods and increased freshwater abstraction has increased sediment stability, which has allowed the seagrass to colonise the entire length of the estuary Marine sedimentation in the lower reaches of the estuary has increased available habitat for salt marsh, which has increased in cover over time. This has compensated for the loss of habitat due to jetties and retaining walls. Low freshwater input and associated nutrients has decreased the species richness and biomass of microalgae.

Figure 1.1: Location map and aerial photograph of the Bushmans Estuary

1.2 Available information on the Bushmans River catchment

1.2.1 Climate

Temperature

Temperature follows a seasonal pattern. The highest mean maximum temperature is 26.3 °C during January and February. The lowest mean minimum temperature is 10.2 °C during July. The difference between mean monthly maximum and mean monthly minimum temperature is smaller for winter months than for summer months. Mean monthly temperatures are between 21.7 °C in January and 15.4 °C in July. The temperatures in the area are mild due to the moderating influence of the sea (Bornman and Klages, 2004).

Precipitation

Rainfall

According to the South African Weather Service, the study area has a bi-modal rainfall pattern with peaks in precipitation during March and October. The mean annual rainfall, over

3 a 30-year period (1961-1990), is 717 mm (Bornman and Klages, 2004). The Alexandria Forest Station measured an average annual rainfall of 74.5 mm for the period 1956 to 1981 (Jolly, 1983). Reddering and Esterhuysen (1981) recorded that the catchment area of the Bushmans River receives annual precipitation ranging from 300 - 400 mm in the upper reaches to 801 – 900 mm in the lower reaches of the estuary. The peak number of rainy days occurs during the summer season with approximately 10 – 11 days rain per month (Bornman and Klages, 2004). The highest rainfall intensity is experienced during the summer months. The most rain that fell in the Port Alfred area during the period (1961 – 1990) measured 175 mm in 24 hours.

Thunderstorms and other forms of precipitation

The Port Alfred and Bathurst region have few incidences of thunderstorms, with about 19 thunderstorms per year on average. Most of these occur in the months from spring through summer into early autumn, with the highest number occurring in March. Fog is uncommon in the area and seems to be the result of moist air from the sea being blown inland and up valleys. The highest incidence of fog in the Port Alfred/Bathurst area is two days per month recorded during the months of March and November. Hail and snow are extremely rare in the region with less than one day with either snow or hail recorded per year (Bornman and Klages, 2004).

Wind

Westerly winds are prevalent in the region for both summer and winter. The wind regime in the Port Alfred area is influenced by seasonal circulation systems. During the summer months, onshore easterly winds are common. Westerly, and especially north-westerly, winds dominate the region during the winter months. Winds with a velocity of more than 30 m.s-1 occur most frequently during the summer months, from September to December (Bornman and Klages, 2004).

Evaporation rates and humidity

The average monthly evaporation rate for the region ranges seasonally from 105 mm in winter to 211 mm in summer and follows a similar trend as the mean temperatures for Port Alfred and Bathurst. Runoff volumes can be estimated from the rainfall and evaporation data, and indicate that the maximum runoff should occur during the summer months. The annual relative humidity in the area shows seasonal fluctuations and ranges from a maximum of 80% to a minimum of 40% for summer and winter, respectively. The mean relative humidity of the air is 72 % (Bornman and Klages, 2004).

1.2.2 Geology and geomorphology

The major part of the lower reaches of the Bushmans River flows through the black shale, compact siltstone and olive-grey sandstone of the Bokkeveld Group (Reddering and Esterhuysen, 1981). Bokkeveld siltstone and shale weather away relatively rapidly, forming valleys and low rolling hills. Along the coast, overlying and resting on the shale, a succession of thin marine sediments, referred to as the Alexandria Formation (Miocene to Pliocene in age), are found. This is mostly of marine origin, deposited during a complex series of regressions of sea level (Rust 1998). Being limestone, the Alexandria Formation is a good aquifer and is responsible for some karst topography in the coastal zone. Karstic landforms may accumulate rainwater for a short while after heavy rain. Some emerge from the base of the limestone just above high tide level at Bushmans River. Unconsolidated beach deposits and bare dunes occur at the coast. Floodtide delta sediment is supplied by a mobile dunefield updrift of the estuary mouth (Harrison et al. 1996). Mudstone, which is extremely susceptible to erosion, constitutes approximately 70% of this formation (Palmer, 1980).

During the 1940s, wind-transported sand caused the inlet channel to migrate northwards, creating a navigable channel for small boats on the Kenton side of the estuary. When due to reduced sand input as a direct result of sand stabilisation (from the construction of the R72 bridge and town development), the inlet migrated back to its original position the channels at the Kenton-on-Sea and Bushmans riverfronts naturally filled with sediment (Reddering and Esterhuysen, 1981).

The fundamental distribution of sediment along the river is mainly marine sand in the lower estuary, progressing gradually into muddy sediment upstream. The banks of the estuary are mostly composed of soft sediments (Forbes, 1998). This could be the result of the river being subjected to frequent flooding resulting in extensive mud deposition on the intertidal sand banks (Robertson, 1984). However, there are a number of rocky areas that run down to the water‟s edge (Forbes, 1998).

1.2.3 Uses and anthropogenic impacts

According to DEAT (2001), agriculture accounted for about 6% of the catchment land-cover in the Bushmans system and approximately 4% of the catchment appeared to be degraded and this mainly comprised of degraded bushland. A high proportion (90%) of the Bushmans catchment was found to be natural and comprised mainly forest and woodland, shrub-, and grassland. Urban development accounted for less than 1% of the land-cover and this was mostly residential and industrial development associated with the coastal resort of Kenton- on-Sea and the towns of Paterson and Alicedale in the upper catchment (DEAT, 2001).

The only major impoundment in the catchment affecting the river is the New Years River Dam situated on the New Years River tributary, which joins the Bushmans River some 130 km from the mouth. The reason for dam construction is a lack of water in the region as it is documented that during droughts water is scarce.

According to Bornman and Klages (2004), there have been long-term reductions in freshwater and this has changed the Bushmans from an estuary characterised by salinity gradients to an essentially homogeneous marine system. This has caused a significant change in the distribution of organisms. For example, there is a likely decrease in phytoplankton biomass and an increase in submerged macrophytes (like Zostera capensis, Adams and Talbot, 1992), an extinction of the river pipefish (Syngnathus watermeyerii) (Whitfield and Ter Morshuizen, 1882) and reduced food resources for suspension feeders (Grange and Allanson, 1995).

In 1942 wind blown sand from the beach to the southwest of the inlet had migrated across the previously existing inlet, forcing the inlet channel to migrate north-eastward from its original position. This formed a navigable channel to the Kenton side of the estuary. The 1942 aerial photograph also shows that Kenton-on-Sea had not yet been developed to any extent. By 1955 the inlet channel had returned to its original position and the town development was in progress on the banks. Between 1955 and 1980 the estuary remained relatively unchanged. The national road bridge had been constructed (1959), and the control of wind blown sand from the beach stabilised the inlet channel (Reddering and Esterhuysen, 1981). Continued south-westerly migration of the channel in the lower reaches during the 1980s led to the construction of the Deacon Cement bag wall on the western bank. This structure consists of rocks packed into wire mesh cages. This wall essentially canalized part of the lower reaches of the Bushmans Estuary. A gabion jetty was also installed on the western river bank upstream from the Deacon Sand bag wall to deflect the water towards the east. The presence of the sandbag wall and gabion jetty may be of dubious benefit for the river, as they violate the principle of unimpeded river flow. However, the ills of the river lie in its upper reaches where it is starved from water by excessive abstraction (Bornman and Klages, 2004).

The estuary serves as a disposal site for a reverse osmosis plant that is located at the mouth. The Albany Coast Water Board (ACWB) supplies bulk water to the community of Kenton- on-Sea and Bushmans River Mouth through the process of desalinizing water from the Indian Ocean (Bornman and Klages, 2004). The salt and chemical products are released into the Bushmans Estuary at the mouth. A specialist study investigated the impacts of this disposal on the estuary and showed complete mixing within 10 m2 of the discharge pipe into the estuary. A lens of dense saline water did not form near the bottom as was expected (Bornman and Klages, 2004). The amount of brine discharged is small relative to the total volume of water in the estuary and there is also turbulent tidal mixing that facilitates the dispersion (as long as the estuary mouth stays open). Therefore the environmental effect is small. However, according to Nel et al. (2006) additional water needs are expected in the Bushmans

6 catchment due to the expected increase in standard of living and tourism opportunities in the region and this poses future threats to the estuary.

The land adjacent to the river and estuary has been mostly utilised for farming activity, which in some areas has disturbed the intertidal areas (Forbes, 1998). The river flows through the districts of Alexandria and Albany, which support extensive pastoral agriculture. Sheep, cattle and goats are the most important stock on land and the stocking rate was calculated as one large animal unit to 5.6 ha. There is no industrial input into the system, though phosphorus and nitrogen may be washed into the river in run-off from fertilised agricultural lands (as well as herbicides and pesticides) (Palmer, 1980). There might also be septic tank input into the lower reaches of the estuary as the surrounding towns, Bushmans River Mouth and Kenton-on-Sea, use septic tanks (Ndlambe, 2007).

Development in the Bushmans Estuary is concentrated around the mouth. Kenton-on-Sea lies on the east bank and the smaller Bushmans River Mouth on the west bank. Up river there are holiday and residential homes (Forbes, 1998). The estuary is primarily used for recreational activities because the Kenton-on-Sea and Bushmans River Mouth are also popular holiday resorts. The activities are fishing, bait collecting, boating, water-skiing, and jet skiing. The facilities that have been provided are a municipal jetty and slipway, 23 private jetties and a small boat marina near the bridge (Forbes, 1998). Forbes (1998) also raised the issue of over- exploitation of some of these activities in the estuary, especially during peak seasons, which could lead to the degradation of this natural resource. The concerns are high diversity of recreational activities and leisure cruising exceeding required limits, especially in the lower reaches and close to the mouth.

1.2.4 Physical characteristics of the Estuary

The mouth of the Bushmans Estuary is constricted by a rocky outcrop on the east bank and a sand spit on the west bank (Robertson, 1984). The constricted, but permanently open, tidal inlet is 50 – 60 m wide and 2 – 3 m deep (Harrison et al. 1996). As a result, the estuary is dominated by flood tides, causing an accumulation of sand on flood-tide deltas in the lower estuary (Reddering, 1988b).

A coastal bridge crosses the estuary approximately 2 km from the mouth; at which point it is 381 m wide. The bridge was built in 1958 (Baird et al., 1983). Water flows into two channels under this bridge – the east channel is 37 m wide and the west channel 56 m. A rubble embankment (288 m) or „island‟ separates the two channels (Palmer, 1980). The main channel depth of the estuary ranges from 2 to 5 m in the navigable portion of the system (Bornman and Klages, 2004). The sediments 3.5 km up the estuary consist almost entirely of mud with a small component of sand (of terrestrial origin), while down river the sediment is mainly quartz–rich sand of marine origin (Gerber, 1992).

Baird et al. (1983) found that the road bridge would cause localised sediment increase because before it was built spring tides and freshwater spates were able to overtop the „island‟ and thus flush away fine sediment. Now, the central part of the bridge sits on this island and prevents flushing. Thus a build-up of fine material immediately up and downstream of the bridge is expected. The mud banks on either side of the bridge have increased from 130 mm in 1957 to 810 m when Baird et al. (1983) did their research, measured along the length axis of the estuary.

The Bushmans Estuary is dominated by sandy sediment and as a result has wide intertidal and supratidal flats and comparatively shallow channels (Reddering and Rust, 1990). The Bushmans River drains through mud-depleted source rocks (Reddering, 1988b). Sediment moves as bed-load into the estuary in response to asymmetric flood and ebb tidal currents (Baird et al. 1981). The presence of marine sand in the lower reaches of the Bushmans Estuary is due to marine sand moving up and down the channel as bed-load during flood and ebb tides. An average of 20 m3 of sand is transported into the estuary over a single spring tidal cycle (Baird et al. 1981).

On top of the naturally low MAR of the catchment (38 x 106 m3.a-1, Reddering and Esterhuysen, 1981 or 57.94 106 m3.a-1, Doudenski, 2004), the occurrence of a number of impoundments have further reduced river flow. The New Year‟s Dam is the major impoundment with a capacity of 4.7 x 106 m3. There are also a number of registered farm dams in the catchment and their capacities are shown in Table 1.1. It is also believed that there are a number of unregistered dams but the capacities of these are unknown.

The tidal currents have their highest value of about 1 m.s-1 at the tidal inlet during spring tides and have low values (< 0.3 m s-1) at distances greater than 6 km from the tidal inlet (Baird et al. 1981). The volume of water exchanged over a tidal cycle, called the „tidal prism‟, is estimated to be in the order of 1 x 106 m3 (Reddering, 1988a, 1988b; Reddering and Rust, 1990). The Bushmans River estuary has a semi-diurnal tidal regime. The constricted mouth causes dampening of the tidal amplitude in the water level from the sea to the estuary (Gerber, 1992).

Table 1.1: Registered farm dams in the Bushmans River Catchment indicating surface area, capacity and year construction completed (Albany Coast Situation Assessment report – DWAF, 2004).

NAME OF DAM SURFACE AREA (HA) CAPACITY (1000 M3) YEAR BUILT New Years River Dam 96 4700 1959 Jameson 15 575 1906 Milner 7.7 255 1898 Concorde 1 80 Springvale 6 163 1983 Kureeleegte 7 64 1962 Blackburn 0 150 1970 Teafontein 3 204 1906 Table Hill Big 8 364 1951 Oakwell 2 82 1969 Brakkloof River 4 75 1960 Strowan 6 139 1950 Hilton 3 60 1954 Shenfield Camp 2 60 1952 Mountain View 1 250 1988 Arnhem 2 80 TOTAL 163.7 7301

1.2.5 Water quality of the estuary

In the Bushmans River, good water quality (Class I) can be expected in the upper Bushmans Rivers to Alicedale. South of Alicedale (this includes the estuary part of the river), water quality deteriorates rapidly due to significant salt loads originating from the Nanaga and Weltevred Formations. Runoff continues to become progressively more saline downstream (Doudenski, 2004).

Nutrients

Because the system is largely influenced by the marine environment, and there is decreased freshwater inflow, it is assumed that the nutrient availability will be largely determined by the sea. However, herbicides, pesticides and fertilizers from surrounding agricultural areas may be transported into the estuary during floods and these may increase the nutrient concentration in the water column of an estuary. Geomorphology also plays a role in nutrient availability; the mudstone and siltstone have relatively high nutrients. High sedimentation rates also mean that there is a great deal of nutrient transportation associated with sediment particles.

Trace metal concentrations were low in the late 70s and early 80s. It was assumed at the time that if there were no further industrial and town developments, these concentrations would remain very low (Watling and Watling, 1983).

Salinity and turbidity

Due to freshwater abstraction by several dams and weirs (approximately 30), salinity levels are usually high. Salinity higher than that of seawater (> 35 PSU) is often recorded during drought years. In a recent study by Bornman and Klages (2004), salinity measurements were 36.2 PSU at high tide and 35.7 PSU at low tide and no salinity stratification was documented. This shows that the estuary has lost the physical characteristics of an estuary and functions as a marine embayment, or as an arm of the sea (Bornman and Klages, 2004). The long-term reductions in freshwater have changed the Bushmans from an estuary characterised by salinity gradients to an essentially homogeneous marine system (Bornman and Klages, 2004). Robertson (1984) also documented that during low rainfall periods salinity as high as 32 PSU was evident as far as 21 km from the mouth, with slight salinity differences within the water column. However, the vertical and horizontal gradients recorded after a flood in July 1983 indicated that the estuary can be reset by small floods (equivalent to 9.4 m3.s-1 of flow). After four weeks the salinity distribution did, however, return to the homogenous state (Robertson, 1984).

The system is well mixed with almost no stratification of the water column at any stage of the tidal cycle and has a low turbidity (< 10 NTU) (Bornman and Klages, 2004). However, according to the Mineralogical Classification of DWAF, both the Bushmans and the Kariega Rivers are classified as “completely unacceptable” because the maximum total dissolved salts (TDS) in these rivers are often greater than 3400 mg.l-1. In the Bushmans River mean TDS has been measured at 2200 mg.l-1 and the upper level at 4000 mg.l-1 during flood events (Doudenski, 2004). This is very high compared to what the TDS of the river is in the headwaters, before it reaches Alicedale, which is generally less than 200 mg.l-1 except during rainfall events (when 500 mg.l-1 TDS can be expected) (Doudenski, 2004). Additionally, Hooker (1996) found that boating activities resuspend sediment and cause an increase in the turbidity of the water.

Temperature

Temperature recordings during surveys of the Bushmans Estuary showed that, mostly, the surface and bottom temperaures were almost identical except those recorded on 13/08/1983, which was after the flood, when very slight differences were apparent (Robertson, 1984). The highest temperatures were recorded during the hotter January and February months (ranging from 20 to 24 oC) and the lowest during winter, in July and August (13.5 to 17oC). Robertson (1984) also recognised that there was a temperature difference between high tide and low tide, with the incoming tide accompanied by a sudden drop in temperature. Bornman and Klages (2004) also recorded this variation with tide, with their results showing the mean temperature ranging from 18.9 oC during low tide to 16.8 oC during high tide because of the influx of cool marine water and no vertical or horizontal stratification evident.

1.2.6 The biota of the estuary

Microalgae

Increased salinity may have caused a significant change in the distribution of organisms, e.g. a decrease in phytoplankton biomass (Hilmer and Bate, 1990). The phytoplankton near the mouth is expected to be of marine origin, whereas at the head species are probably of freshwater origin. The middle reaches of the estuary will be expected to have more salinity tolerant species (Bornman and Klages, 2004).

Macrophytes

Submerged Macrophytes

Hodgson (1986) recorded that Zostera capensis only occurred 5 km up the estuary from the mouth. More recently, eelgrass has been recorded in the lower, middle, and upper reaches of the estuary. An increase in salinity and stable sediment conditions would have encouraged expansion upstream. Forbes (1998) and Hooker (1996) found that bait digging and people walking on Zostera beds disturbs the plants and this needs to be examined because recreation is a major activity in the Bushmans Estuary.

Salt marshes

Spartina maritima dominates the salt marshes in the lower 4 km of the estuary, and in total it covers almost 31% of the total salt marsh area of the Bushmans Estuary. In the middle reaches Sarcocornia species occurred as a co-dominant (Adams, 1995).

In 1995 the upper reaches of the Bushmans Estuary was dominated by Sarcocornia spp. which covered 48% of the upper salt marshes where Spartina maritima only covered 5%. Limonium scabrum consistently covered 12-14 % of all salt marsh areas measured. Bassia diffusa occurred in 16 of the 20 marsh areas sampled but always in low numbers. It only covered about 6% of the total marsh area (Adams 1995). Hodgson (1986) measured a mean biomass of 319 g.m-2 dry weight for Z. capensis, 1570 g.m-2 for S. maritima and 400 g.m-2 for Sarcocornia spp.

Reeds and Sedges

Phragmites australis (Cav.) Trin ex Steud was found in one marsh, 5.2 km from the mouth, where it covered 3 % of the total marsh area. This localised area, in which Phragmites was found, is an area where freshwater enters the Bushman River from a minor tributary, thus making it possible for the Phragmites to survive (Adams, 1995).

The salt marshes in the Bushmans Estuary are well zoned with Zostera capensis at the low water mark, followed by Spartina maritima in the intertidal zone. Above the Spartina zone Sarcocornia spp. occur followed by Triglochin maritima. Where present, Atriplex sp. and Bassia diffusa indicated the upper limits of the marsh area (Adams, 1995). This pattern is in accordance with the zonation pattern, which usually occurs in estuaries where tidal exchange predominates (Adams, 1991).

Invertebrates

Increased salinity will cause decreased phytoplankton availability eventually affecting the food resource for suspension feeders (Grange and Allanson, 1995). The secondary production and availability of the invertebrates is dependent on primary productivity.

According to some unpublished data by Hodgson, found in Hodgson (1986), the Bushmans Estuary had 10 – 20 Solen cylindraceus per m2 and 120 mudprawns (Upogebia africana) per m2 at that time. Forbes (1998) recorded 140 Upogebia africana, 50 sandprawns (Callianassa kraussi) and 10 pencil bait (Solen spp.) per m2 and found that these species were more abundant in the lower reaches than the rest of the estuary.

In the adjacent Kariega Estuary a great diversity of invertebrate species has been recorded and this was largely attributed to the fact that Zostera capensis extended for the length of the estuary and provided microhabitats for many species (Hodgson, 1986). The estuary also experiences reduced freshwater input and thus less siltation, which also acts in favour of the invertebrates. The same can be expected in the Bushmans; Zostera now occurs along the entire length of the estuary and is thus expected to increase invertebrate numbers.

Fish

It is now believed that the river pipefish Syngnathus watermeyerii might be extinct as this species has not been caught in this system for about 40 years. The river pipefish is endemic and used to occur in Bushmans and Kasouga estuaries. It is believed that, among other factors, salinity increase might be the cause of the extinction of this species (Whitfield and Ter Morshuizen, 1992).

In a preliminary study by the CSIR team, Harrison et al. (2000), as part of the State-of-the Environment study for South Africa‟s estuaries for the Department of Environmental Affairs and Tourism found that 84% of the fish assemblages in the Bushmans Estuary were similar to the reference assemblages. The species relative abundance was similar to the reference conditions by 54%.

Birds

The coordinated waterbird counts (CWAC) group has been counting birds along the Bushmans Estuary, from the mouth to Ghio Bridge (15 km from the mouth) since 20 January 2000. The last count reported on the CWAC website was done on the 7th of February 2008. The CWAC project does summer and winter counts of the birds, and nine summer and eight winter counts have been conducted for the Bushmans Estuary. Table 1.2 shows the common birds that have been counted in the estuary by the CWAC programme. During an outing to the Amakhala Game Reserve on the 18th of April 2004 (in the vicinity of the Bushmans Estuary), Bo Bonnevie reported that quite a large number of birds was also observed (Table 1.3); from the News Magazine of the Diaz Cross Bird Club (Bonnevie, 2004).

An Osprey (Pandion haliaetus) and three African Skimmers (Rynchops flovirostris) have visited the estuary for a short period in March 1971 (Jubb, 1972). This was unsual as these two species are uncommon in the area and have been frequently observed in the Zambezi River (Jubb, 1972). The Osprey is widely distributed, though not always common, and prefers coastal reaches and large inland waters of temperate and tropical regions and the African Skimmer favours large inland tropical waters and some coastal waters (Jubb, 1972).

Estuarine biota in the Bushmans Estuary are typical of permanently open estuaries in the Eastern Cape and are reported in Jubb (1972); Palmer (1980), Heydorn and Grindley (1982), Robertson (1984), Whitfield and Ter Morshuizen (1992), Harrison et al. (1996), Robertson (1996) and Lubke and de Moore (1998).

1.2.7 The importance of the estuary

The conservation importance of the Bushmans Estuary has been rated as 41 out of the 250 ranked estuaries in South Africa and the importance score is 78.1 out of 100 (Turpie and Clark, 2007). This importance score is the combination of the scores of the size of the estuary, which is 100, habitat importance (60), zonal type rarity (20), and biodiversity importance (84.5). The ecological integrity is fair, which means there is a noticeable degree of degradation in the catchment and/or estuary (Whitfield, 2000).

Turpie et al. (2002) ranked the estuary as 44th in terms of conservation importance and 18th as a desired protected area, with no part that was protected at the time. Harrison et al. (1996) scored the estuary 8 out of 9 on the overall estuarine health.

According to Turpie and Clark (2007) the estuary needs to be rehabilitated immediately. The priority of the rehabilitation is water quantity, to fix inappropriate bank stabilisation and to clear alien species.

Table 1.2: List of the common birds counted along the Bushmans Estuary (that would be birds spotted more than 10 times during the 17 counts that were conducted) (retrieved from the CWAC website).

Summer Winter All Species Name Min Avg Max Min Avg Max Min Avg Max Cape Cormorant: Phalacrocorax capensis 1 0.6 3 2 14.3 29 1 7.0 29 Reed Cormorant: Phalacrocorax africanus 2 3.4 12 6 21.4 42 2 11.9 42 Grey Heron: Ardea cinerea 6 13.1 25 5 9.9 14 5 11.6 25 Black-headed Heron: Ardea melanocephala 1 11.3 50 1 1.9 4 1 6.9 50 Great Egret: Egretta alba 1 6.6 19 3 7.5 12 1 7.0 19 Little Egret: Egretta garzetta 2 7.6 24 6 16.6 24 2 11.8 24 African Sacred Ibis: Threskiornis aethiopicus 1 3.8 15 5 14.5 23 1 8.8 23 Hadeda Ibis: Bostrychia hagedash 2 0.2 2 1 6.9 34 1 3.4 34 Egyptian Goose: Alopochen aegyptiacus 2 3.7 17 2 5.6 13 2 4.6 17 Yellow-billed Duck: Anas undulata 1 3.8 16 1 9.8 47 1 6.6 47 Ruddy Turnstone: Arenaria interpres 1 8.0 14 2 0.3 2 1 4.4 14 Common Ringed Plover: Charadrius hiaticula 3 14.1 30 0.0 3 7.5 30 White-fronted Plover: Charadrius marginatus 2 9.7 17 8 21.5 43 2 15.2 43 Grey Plover: Pluvialis squatarola 15 31.7 50 4 10.9 32 4 21.9 50 Blacksmith Lapwing: Vanellus armatus 6 8.7 32 2 9.5 20 2 9.1 32 Curlew Sandpiper: Calidris ferruginea 5 22.4 60 0.0 5 11.9 60 Little Stint: Calidris minuta 19 2.1 19 0.0 19 1.1 19 Sanderling Sanderling: Calidris alba 6 56.2 110 7 0.9 7 6 30.2 110 Ruff Ruff: Philomachus pugnax 120 28.3 135 0.0 120 15.0 135 Common Sandpiper: Actitis hypoleucos 1 4.1 8 1 0.9 2 1 2.6 8 Common Greenshank: Tringa nebularia 9 25.9 51 1 8.8 19 1 17.8 51 Common Whimbrel: Numenius phaeopus 18 30.2 41 3 7.0 14 3 19.3 41 Black-winged Stilt: Himantopus himantopus 2 1.0 7 2 2.9 14 2 1.9 14 Kelp Gull: Larus dominicanus 7 15.0 33 12 20.1 47 7 17.4 47 Common Tern: Sterna hirundo 5 20.2 59 1 0.1 1 1 10.8 59 Swift Tern: Sterna bergii 1 5.0 18 5 12.1 44 1 8.4 44 Pied Kingfisher: Ceryle rudis 2 6.2 14 4 8.9 12 2 7.5 14 Cape Wagtail: Motacilla capensis 2 4.9 15 2 5.8 10 2 5.3 15 Mountain Wagtail: Motacilla clara 11 1.2 11 0.0 11 0.6 11

Table 1.3: List of birds spotted in the vicinity of the Bushmans Estuary

Ostrich Three-banded Plover Bar-throated Apalis Reed Cormorant Crowned Plover Yellow-breasted Apalis Darter Temminck's Courser Bleating Warbler Grey Heron Rock Pigeon Cloud Cisticola Black-headed Heron Red-eyed Dove Grey-backed Cisticola Black-crowned Night Heron Cape Turtle Dove Neddicky Hadeda Ibis Laughing Dove Karoo Prinia South African Shelduck Green-spotted Dove Dusky Flycatcher African Fish Eagle Knysna Lourie Fiscal Flycatcher Jackal Buzzard Little Swift Cape Batis Rock Kestrel Speckled Mousebird Cape Wagtail Helmeted Guineafowl Red-faced Mousebird Grassveld Pipit Stanley's Bustard Malachite Kingfisher Orange-throated Longclaw Olive Woodpecker Brown-hooded Kingfisher Fiscal Shrike Clapper Lark Hoopoe Southern Boubou Spike-heeled Lark Red-billed Woodhoopoe Bokmakierie Red-capped Lark Black-eyed Bulbul Olive Bush Shrike European Swallow Terrestrial Bulbul Glossy Starling Greater Striped Swallow Sombre Bulbul Red-winged Starling Rock Martin Cape Rock Thrush Red-billed Oxpecker Black Saw-wing Swallow Familiar Chat Lesser Double-collared Sunbird Fork-tailed Drongo Mocking Chat Greater Double-collared Sunbird Black-headed Oriole Stonechat Black Sunbird Black Crow Cape Robin Collared Sunbird White-necked Raven White-browed Robin Cape White-eye Crowned Hornbill Karoo Robin Cape Sparrow Red-fronted Tinker Barbet Titbabbler Forest Weaver Knysna Woodpecker Quail Finch Spectacled Weaver Cardinal Woodpecker Bully Canary Spotted-backed Weaver Cape Weaver Streaky-headed Canary

Chapter 2: Literature Review

2.1 Botanical communities in estuaries and their importance

In deep clear estuaries, phytoplankton appears to be the more important primary producers, whilst in broad shallow estuaries the important contributors are submerged macrophytes, benthic microalgae growing on the mud flats and either salt marsh vegetation or mangroves (Day, 1981). However, besides the physical structure, a number of other factors collectively play a role in the botanical structure of estuaries. Based on Colloty‟s BIR (botanical importance rating) index (Colloty et al., 1998) and the guidelines for the determination of the present ecological state by DWAF (2008a), the botanical importance score will be higher if most of the nine identified communities are present and if they each cover larger areas. Therefore, the factors that govern the presence or absence of these communities have to be understood (Riddin,1999). Of the nine communities, only the salt marsh, submerged macrophytes, reeds and sedges and microalgae (phytoplankton and benthic) will be discussed as these occur in the Bushmans Estuary.

2.1.1 Salt marsh

Occurrence and distribution

Salt marsh vegetation forms a transition between land and salty or brackish water (e.g., sloughs, bays, estuaries). It comprises the low growing halophytic grasses, sedges and succulents, which cover the muddy banks of estuaries at and above mid-tide levels in temperate regions (Day, 1981; Lubke and de Moor, 1998; Riddin, 1999). Halophytes are species that have the ability to complete their life cycle in salt concentrations from 100 – 200 mM NaCl (Ranwell, 1972). Along the South African coast, only a few estuaries support the growth of salt marshes. More than 75% of the 17 000 ha of the salt marsh area in the country is confined to only five systems. These are the Langebaan Lagoon (containing more than 30% of South Africa‟s salt marsh), Knysna Lagoon, the Swartkops Estuary, the Berg Estuary, and the Olifants Estuary (O‟Callaghan, 1994).

The environmental characteristics that affect the occurrence of certain species in a salt marsh community are mostly patterns of tidal inundation and salinity. Salt marsh plants show distinct zonation patterns along tidal inundation and salinity gradients. This zonation is more distinct in estuaries with a large tidal range (Adams et al., 1999); where there is a small tidal range mosaic patterns occur rather than distinct zonal bands. Due to these environmental features salt marshes in South Africa have been divided into subtidal, intertidal and supratidal marsh areas and are often referred to in the literature as the low marsh, middle marsh and high marsh (Bornman, 2002). At the low water mark salt marsh growth is limited by

16 inundation period, while growth at the upper level is limited by competition with terrestrial plants (Carter, 1988).

In estuaries where distinct zonation is evident species composition along these zones also differs. In the Langebaan Lagoon, with predominating tidal exchanges, Adams et al. (1999) described that the salt marsh species composition in the mean sea level to mean high water neap zone (low marsh) only consisted of Spartina maritima. The next zone, which is in the mean high water neap to mean high water spring zone (middle marsh), consists of Sarcocornia perrenis, Triglochin bulbosa, Triglochin striata, Salicornia meyeriana, Cotula coronopifolia, Limonium sp., Bassia diffusa and Sueada inflata. This zone was lastly followed by a community consisting of Sarcocornia pillansii, Puccinella angusta, Disphyma crassifolium and Plantago crassifolia in the upper marsh zone above mean high water spring (Adams et al., 1999). Other facts important in shaping salt marsh plant communities, additional to zonation are nutrient availability, perturbation by wrack, ice damage and herbivory, as well as positive interactions (e.g. plant cover alleviating saline stress) (César et al., 2003).

The characteristics of the sediment that salt marshes grow in are also very important for their survival. The most crucial are the salinity of the sediment, moisture and organic content, sediment texture and sediment compaction shear strength. Sediment salinity determines the vertical and horizontal zonation of salt marsh species (Wolters et al., 2005). The depth of the water table, rainfall and evaporation on the marsh, groundwater seepage from adjacent land and the salinity of the tidal water that inundates the marsh control sediment salinity (Cisneros et al., 1999).

Sediment moisture content is variable both spatially and temporally and is influenced by a number of factors, such as topography, vegetation cover, depth to water table and rainfall (Gomez-Plaza et al., 2001). Moisture induces seed germination, which makes it a requirement for plant growth. The organic matter is important as it increases the water holding capacity of the sediment and also provides the sediment with nutrients (Bai et al., 2005). Most moisture and organic content can be found in clay sediment; however the moisture is often not available to plants due to the small particle size of clay (Gomez-Plaza et al., 2001). The texture of the soil also plays an important role in seed germination, seedling establishment, root growth and distribution of the salt marsh vegetation (Bornman et al., 2004). The distribution of the two main sediment sources in estuaries, fluvial and marine sediments, determines the distribution of the plants. The deposition of these sediments is controlled by the size of the particle and the speed of the water current (Shaw, 2007).

Bornman et al., (2002) highlighted the importance of freshwater flooding in reducing the salinity content of floodplain and high marsh soils thus promoting healthier conditions for salt marsh plants. Species distribution, germination, and growth in a salt marsh are primarily determined by soil salinity and moisture. A single large pulse of freshwater during rain is

17 essential for the survival of the vegetation on the floodplain salt marsh. The rainfall event, if large enough, will trigger seed germination, seedling growth, and allow the mature plants to access the fresher water accumulating on the soil surface (Bornman, 2002). In the Olifants Estuary, the highest percentage vegetation cover was recorded in low-lying areas on the floodplain and was related to the depth to the water table. Most plants extend their roots to the low salinity water found there as the high salinity of the surface soil makes it unsuitable for plant growth. The plants appear to use this resource only as a source of moisture to maintain essential metabolic processes to survive the drier months of the year (Bornman et al., 2002). Crain et al. (2004) added that salt marsh plants can thrive in freshwater environments but due to the high competition with other plants they are displaced from these environments. They grow best in the absence of competition when in freshwater and rather persist in saline environments because of their ability to tolerate salt and can dominate alone without having to compete for resources with a large variety of plants (Crain et al., 2004).

Ecological importance

In many parts of the world the value of salt marshes is increasingly recognised and more widely appreciated, and salt marsh restoration is a growing practice in science. Salt marshes are being recognised for their support for biodiversity, aesthetics and economy (Spurgeon, 1998; Shaw, 2007,). To highlight the importance of salt marsh for conservation, a species of gecko was re-discovered in the salt marshes of the Kromme Estuary (Branch, 2001), after being thought to be extinct for a long time. It is the only local lizard restricted to saltmarsh habitats, and is known only from the Kromme Estuary and a few sites near Port Elizabeth.

Salt marshes have a number of important functions. Among these are sediment stabilisation and bank protection, filtering of sediments and pollutants, providing shelter and food for marine and estuarine organisms (Adams et al., 1999). Spartina, particularly, acts as a sediment stabilizer during times of erosion (Brown et al., 1998), and their stems and epiphytic algae and meiofauna provide important resources in salt marsh food webs (Gregg and Fleeger, 1997). Salt marshes support waterbirds, and because people value recreational bird-watching, various governments support the conservation and restoration of coastal wetlands and endorse international agreements such as the Ramsar and Bonn Conventions (Shaw, 2007).

Furthermore, salt marshes are important inorganic and organic nutrient sources because they serve as zones of nutrient production and retention. The estuarine water level and degree of tidal flushing determine the amount of nutrient release into the water column. When plant biomass decays on the marsh surface, its energy enters the estuarine food chain as detritus and this depends on tidal flushing (Adams et al., 1999). Salt marshes may make up a large contribution to the primary production of an estuary. In the Kromme, marsh plants were found to only cover 38% of the estuary but contributed over 78% of the total primary production. Less than 10% of this production reaches the grazing food chain and the

18 remainder gets broken into sediment detritus (Heymans and Baird, 1995). Also, in Kariega, a supratidal Sarcocornia-Bassia salt marsh only exported 6% of production to the estuary (Taylor and Allanson, 1995).

Supratidal marshes can function as a sink for organic carbon. During high spring tides detrital plant material originating from low salt marsh areas is deposited onto high salt marsh areas. After long periods of exposure, which are common in the high marsh area, the heterotrophic breakdown and decomposition of the detrital matter occurs. Carbon is then released into the atmosphere as carbon dioxide produced by respiration (Taylor and Allanson, 1995).

Disturbances and threats

Estuarine salt marsh vegetation is under threat on a global scale through the potential impact of global sea-level rise, and on a local scale due to losses from natural and human-induced disturbances (Svensson et al., 2007). Where estuarine food webs are reliant on significant inputs from fringing salt marsh and macrophytes, there is the potential for flow-on effects to higher trophic levels, such as large invertebrates and fish.

Prolonged Inundation

Salt marsh inundation studies have shown that an increase in the submergence of marshes causes changes in soil physico-chemical properties including oxygen depletion (decline in soil redox potential), accumulation of toxic compounds and a change in nutrient cycling (Adams and Bate, 1994a; 1995). When they investigated the effect of a rise in sea level would have on higher salt marsh plants, Miller et al. (2001) found that increased inundation rates reduced respiration more than production. The result was a net loss in organic carbon from the high marsh. These alterations adversely affect plants by altering metabolic functioning, causing limitations to gas exchange and increasing anaerobic root respiration and eventually inhibiting growth. Inevitably, after prolonged periods of total submergence, the inundated organs die, and possibly the whole plant (Jackson, 1990). Mouth closure in the Great Brak Estuary and subsequent submergence of Sarcocornia natalensis communities for two to three months led to the dieback of this community (CSIR, 1992). Sarcocornia perennis also got stressed after an inundation of just two weeks (Adams and Bate, 1994a). O‟Callaghan (1990) found that prolonged inundation of Sarcocornia spp. slows flowering and subsequent seed production of the plant.

Salinity and reduced freshwater inflow

In many semi-arid areas, the natural functioning of estuaries is threatened because of freshwater impoundment (Bornman et al., 2002). This matter gained prominence when modifications to the normal flow of the Orange River on the border between Namibia and South Africa resulted in the collapse of the salt marsh near the mouth and a decrease in

19 migrant bird numbers that threatened the Ramsar status of this unique wetland. The Council for Scientific and Industrial Research (CSIR 1991) identified the lack of back flooding by freshwater as the cause for the dieback of Sarcocornia pillansii (Moss) A.J. Scott, which was the dominant species in that salt marsh. Shaw et al. (2008) reported that the groundwater and surface sediments of the Orange River Mouth are hypersaline and above the tolerance range of the dominant plant Sarcocornia pillansii. Further findings were that the remaining small areas of S. pillansii have the potential to produce 40 billion seeds, which would be sufficient to revegetate the desertified marsh, but due to hypersaline sediment conditions in this estuary the successful establishment and survival of the seedlings remained low.

Although most halophytic plants grow optimally in the presence of salt, they are inhibited by high salt concentration, with reduced growth at 35 PSU (Ungar, 1991). Increases in salinity have affected marsh plants in a number of ways: increases in ash content due to the accumulation of Na+ and Cl- (Ungar, 1991), increases in the fresh weight, dry weight and succulence (up to the species‟ salinity tolerance limit after which it decreases again; Adams and Bate, 1994a), and also lowering of plant water potential. Plant water potential and tissue water content are decreased when salt concentrations increase. Consequently, growth inhibition occurs, which is attributed to dehydration at high salinity. The dehydration results from loss of cell turgor because of inadequate tissue osmotic adjustment (Ungar, 1991). Laboratory tests performed by Shaw et al. (2008) indicated that seeds of S. pillansii germinate best in freshwater, with 40% germination at 0 PSU compared to 5% at 35 PSU. Exposure to hypersaline conditions (70 PSU) decreased the viability of the seeds. This study concluded that the potential of the Lower Orange River marsh to rehabilitate naturally depends on sufficient freshwater entering the marsh (Shaw et al., 2008).

2.1.2 Submerged macrophytes

Submerged macrophytes are plants rooted in both soft subtidal and low intertidal substrata, whose leaves and stems are completely submersed for most states of the tide (Adams et al., 1999). There are about sixty such species worldwide. Touchette (2007) describes seagrasses as forming part of a critical, and fragile, ecosystem inhabiting shallow coastal embayments and estuaries throughout the world.

Occurrence and distribution

In South Africa the seagrass, Zostera capensis occupies the mudbanks of most permanently open Cape estuaries (Adams et al., 1999) and survives and grows best in the 15 to 35 PSU salinity range (Adams and Bate, 1994b). The leaf turnover of Z. capensis is in the order of days and shoot turnover is in the order of one to three months (Light and Woelkerling, 1992). During its life time, Z. capensis continually forms new leaves, and old leaves become senescent and detach, especially in late summer (Verhagen and Nienhuis, 1983). Halophila ovalis is an opportunistic species that seldom occurs on its own; it is usually associated with

Z. capensis (Day, 1981). It usually occupies estuaries immediately after flood events because it has the ability to rapidly colonize sandy substrata (Talbot et al., 1990). In temporarily open estuaries that are characterized by fluctuating salinities Ruppia cirrhosa occurs, however, it can also occur in the calm, brackish upper reaches of permanently open estuaries (Adams et al., 1992; Adams and Bate, 1994b).

Since these plants must photosynthesize, they are limited to growing submerged in the photic zone, and most occur in shallow water, but may grow in deeper areas where the water is particularly clear (Ohrel and Register, 2006). They are primarily limited to areas where they remain submerged, but some species can withstand exposure during low-water periods (e.g., low tide). However, prolonged exposure during low tide and inundation by deep water during high tide, especially when the water is cloudy, can make for undesirable habitat conditions.

Desiccation or exposure periods can drastically affect the usually submerged aquatic vegetation. The ability of a submerged macrophyte to survive desiccation largely depends on its ability to recover from exposure, and its rapid ability to not only recuperate the desiccated leaves but also produce new leaves. Zostera capensis can recover within a day after short- term desiccation, while Ruppia cirrhosa took four days to recover (Adams and Bate, 1994b). The ability of Z. capensis to recover quickly is because of its possession of a leaf sheath that protects the basal meristem during exposure. However, with long periods of desiccation, Z. capensis cannot recover, as well as R. cirrhosa (Adams and Bate, 1994b). The rate of desiccation depends on light, temperature, humidity, wind, and nature and topography of the substrate. Waterlogged conditions of creek sediments and water being trapped in small pools or trapped within the seagrass beds allow seagrasses to meet the evaporative demand of their large leaves during tidal emergence (Talbot and Bate, 1987). Z. capensis is dominant in tidal marine South African estuaries because of its stronger morphological structure and ability to survive daily periods of exposure compared with that of R. cirrhosa (Adams and Bate, 1994b).

Salinity, temperature, and sediments also determine, to a large extent, which species can survive (Ohrel and Register, 2006). Heavy turbidity, siltation and decreased light penetration have been responsible for population losses in Kwa-Zulu Natal and Eastern Cape estuaries (Day, 1981). Below ground biomass often dominates the total plant biomass of seagrass communities (Stevenson, 1988; Kuo and McComb, 1989) because the plant allocates a big portion of their production to the roots (Duarte et al., 1998).

Seagrasses are sensitive to the deposition of sediment directly on top of them when the sediment deposition is greater than their ability to grow through it (McKenzie, 2007). Furthermore, sediment stability influences macrophyte colonization. Submerged macrophytes do not occur in systems where the sediment is constantly modified by dynamic processes (Adams et al., 1999). This was the case in the Palmiet River Estuary, found in a study by Adams and Talbot (1992), which is influenced by strong water currents and frequent

21 flooding. On the other hand the Kromme Estuary experienced increased biomass and area distribution of Zostera capensis. This was because freshwater inflow had been reduced and consequently sediment stability increased (Adams and Talbot, 1992).

Sediment characteristics affect seagrass growth, germination, survival and distribution. Sediment texture, in particular, affects diffusion of oxygen, rhizome elongation and levels of nutrients and phytotoxins, such as sulfides. Sandy-textured sediments tend to diffuse oxygen more readily, obstruct rhizome elongation, and have lower fertility. On the contrary, finer textured sediments will tend to have higher fertility, allow rhizome elongation, and will tend to have greater levels of anoxia as pore water will have less interaction with the overlying water column. These anaerobic conditions may stimulate germination in some species, but also result in elevated sulfide levels, which inhibit leaf biomass production in mature plants and are toxic to seedlings of some species (McKenzie, 2007).

In the Tweed River Estuary in New South Wales, Australia, Hossain (2005) found increases in area covered by seagrasses in the period 1999 to 2001. The increase was attributed to water clarity and salinity, which may be associated with drought conditions. Geomorphic stability was also considered as a contributor to seagrass increase. The study also showed variation in seagrass biomass between sites and periods of time (seasons). At the two different sites where seagrass biomass was monitored, the highest biomass of Zostera capicorni was recorded in -2 spring. The mean biomass was 78 g.m in the marine dominated open embayment site at Towra Point while at a site in the Tweed River Estuary the average mean biomass was 208 g.m-2. Both these sites were sandy marine deltas but differed in wave action and degree of pollution. The differing biomass was particularly attributed to high rates of sediment delivery in the Towra Point in Wooloova Bay (Hossain, 2005).

According to Adams et al. (1999), “in estuaries subject to episodic flooding the related sedimentary disturbances appear to be the most important factor determining the state of seagrasses”. However, complete removal is also a factor. For example, in the Kwelera and Nahoon estuaries, a 15 year flood completely removed Zostera capensis beds (Talbot et al., 1990). Adams et al. (1999) have noted that it is the moderate (2-3 years) and light floods (1 year) that lead to fluvial deposits and smothering of macrophyte beds, resulting in impaired growth and shortened leaf lengths. Also, currents that are greater than 1 m.s-1 have been said to result in the removal of submerged macrophytes (Adams et al., 1999), while 0.5 m.s-1 results in mechanical damage and those less than 0.1 m.s-1 favour the growth and establishment of the macrophytes (Adams, 2008). Studies on South African estuaries have shown that changes in Z. capensis biomass are linked to flooding activities rather than seasonal influences (Edgecumbe, 1980; Talbot et al., 1990). Sediment loading and removal caused by flooding can disturb the submerged macrophyte community. This would lead to reductions in light, and increased concentrations of silt, organic matter and nutrients (Campbell and McKenzie, 2001, 2004). In the Sandy Strait, Queensland the loss and recovery of intertidal seagrass meadows were assessed following the flood related catastrophic loss of

22 seagrass meadows in February 1999. Mapping surveys showed that approximately 90% of intertidal seagrasses in the northern Great Sandy Strait disappeared after the February 1999 flooding of the Mary River. Full recovery of all seagrass meadows took 3 years. Reduced water quality that was characterised by 2-3 fold increases in turbidity and nutrient concentrations during the 6 months following the flood, was followed by a 95% loss of seagrass meadows in the region. Reduction in available light due to increased flood associated turbidity in February 1999 was the likely cause of seagrass loss (Campbell and McKenzie, 2004).

On the other hand, due to the construction of water storage dams, the frequency of flooding events has been reduced. This leads to an increase in the growth and expansion of submerged macrophytes (Adams et al., 1999). This has been the case in the freshwater deprived Kromme Estuary, where upriver dams reduce the effect of all floods smaller than 1-in-30 years (Bickerton and Pierce, 1988). The submerged macrophytes in this estuary have expanded due to increased sediment stability and improved water clarity that was related to a lack of freshwater input (Adams and Talbot, 1992). Adams (2005) noticed that, since the construction of the Mpofu Dam, Zostera capensis biomass in the Kromme Estuary increased from 217 to 273 g.m-2 (Adams and Talbot, 1992). The reduced flooding and stable salinity and sediment conditions, due to the dam, had promoted the growth of the macrophyte (Adams, 2005).

Ecological importance

Seagrass meadows are a major source of primary production, providing habitat and food for associated organisms (Austoni et al., 2007). Under optimal conditions, seagrasses are highly productive with biomass accumulation rates comparable to many agriculturally important plant species. Aboveground tissues also provide substrate for epiphytic organisms, which further enhances total productivity of the system by as much as 35% (Touchette, 2007). Submerged macrophyte beds are highly diverse and productive ecosystems, and can harbour hundreds of associated species from all phyla, for example juvenile and adult fish, epiphytic and free-living macroalgae and microalgae, mollusks, bristle worms, and nematodes (Ohrel and Register, 2006). In addition, juvenile and larval fish and crustaceans use the macrophyte beds as protective nurseries and to hide from predators (Whitfield, 1984; Ohrel and Register, 2006). Shedding crabs conceal themselves in the vegetation until their new shells have hardened (Ohrel and Register, 2006). Thus, as stated by Touchette (2007), apart from their ecological significance, seagrasses provide considerable economic value by contributing to recreational and commercial fisheries, and outdoor sporting activities (e.g. waterfowl hunting and tourism).

The macrophytes feed epifaunal and benthic invertebrates because detritus, diatoms and filamentous algae are trapped in them (Whitfield, 1989; Ohrel and Register, 2006). Indirectly they provide food for carnivorous fish species which feed on the diverse and abundant

23 invertebrates they harbour (Whitfield, 1984). Although only a few truly aquatic species consume the living plants (e.g., manatees, sea turtles, and some species of fish), several types of waterfowl and small mammals rely on them as a major portion of their diet (Ohrel and Register, 2006). Bait organisms are mainly found in the Zostera zone (Adams et al., 1999), for example, mudprawn (Upogebia africana), cracker shrimp (Alpheus crassimanus), bloodworm (Arenicola loveni) and pencil bait (Solen cylindraceus).

During the growing seasons of spring and summer, submerged macrophytes supply oxygen to the water through the process of photosynthesis, thereby helping to support the survival of aquatic organisms. They also play an important role in nutrient trapping and recycling (Adams et al., 1999). The plants take up large quantities of nutrients, which remain locked in the plant biomass throughout the warm weather seasons. As the plants die and decay in autumn, they slowly release the nutrients back into the ecosystem at a time when phytoplankton blooms pose less of a problem (Ohrel and Register, 2006). In Swartvlei, Potamogeton pectinatus acted as a nutrient pump by utilizing the sediment as a phosphorus source and releasing it into the water after decay (Howard-Williams and Allanson, 1987).

Since roots bind the sediments on the estuary bottom and retard water currents, plants minimize water movement and allow suspended sediments to settle, thus improving water clarity (Fonseca et al., 1982; Ohrel and Register, 2006; McKenzie, 2007). The submerged macrophyte community acts as a protection layer (buffer) between the coast and the catchment environment. The macrophyte beds buffer the shoreline and minimize erosion by dampening the energy of incoming waves (Ohrel and Register, 2006). They also affect coastal water quality by absorbing nutrients and trapping sediments acting as a buffer between catchment inputs and reef communities (McKenzie, 2007).

In the early to mid-1930s, the importance of seagrass became apparent after 90 % of Zostera marima was destroyed along the European and North American coasts. This catastrophic seagrass declines was called the “wasting disease” and while the primary cause of this episode has not been resolved, the geomorphological and biological consequences to seagrass loss became apparent. Often in the absence of seagrasses, it was noticeable that substrates became increasingly coarser, sandy beaches eroded to rocky slopes, and macroinvertebrate assemblages shifted from burrowing and deposit-feeding species to encrusted filter feeding organisms (Touchette, 2007). In North America, other organisms that rely on seagrasses as food or habitat also declined with eelgrass disappearance, including waterfowl (e.g. Branta bernicla) and shellfish (e.g. Argopecten irradians).

Disturbances and threats

According to the above important ecological facts, submerged macrophytes are an important component of marine ecosystems; however, their distribution is in decline due to anthropogenic disturbances and declining water quality (Austoni et al., 2007; Touchette,

2007). Such activities include cultural eutrophication, herbicide runoff, and increased turbidity related to coastal development, boat traffic, and dredging activities. Seagrasses are particularly vulnerable to environmental perturbations that result in diminished light availability, that is, eutrophication and suspended sediments (Touchette, 2007).

Recreational activities

Due to their diverse and protected nature, estuaries are popular recreational outlets with an ever-growing demand for leisure and recreational opportunities. Pressure on estuarine environments is thus an inevitable phenomenon, which then ends up affecting the biological inhabitants. Recreational boating can result in an increase in turbidity which then reduces light penetration into the water column (Forbes, 1999; Adams, 2005). This reduces photosynthetic activity, and has been thought to be one of the major factors responsible for reduced biomass of submerged macrophytes (Mason and Bryant, 1975).

It has been noted that boats, with two-stroke outboard motors, release about 40 % fuel into the water (Muratori, 1968). Substantial amounts of fuel enter water courses and rapidly become dispersed through the mixing action of the propellers, which detrimentally affects water quality (Forbes, 1998). A retardation of plant growth due to boat emissions was observed in the Bushmans Estuary by Hooker (1996) but the concentration of these emissions was not sufficiently toxic to kill the plants over the four week study period. The fuel compounds, which mostly consist of raw fuel, non-volatile oil, volatile oil, lead and phenols (Jackivics and Kuzminski, 1973), can also retard faunal growth by inhibiting respiratory appendages from functioning adequately (Forbes, 1998).

Hooker (1996) also highlighted the physical removal of the plants by boat propellers. The finding was that up to 58 % of the total biomass of Z. capensis beds can be lost at one time when a boat is driven over the beds.

Bait digging was another activity that was found to have a big impact on the Zostera capensis beds of the Bushmans Estuary (Hooker, 1996). In the areas where bait digging was occurring the submerged macrophytes were growing sparsely. Marginal vegetation may also be damaged by people walking parallel to the waters edge or seeking access to the water for swimming and fishing. Forbes (1999) found that low levels of human trampling decrease both the total plant cover and species diversity in a variety of habitats.

Increased nutrient input

While nitrogen and phosphorus may play an important role in the growth of aquatic macrophytes, an excess of these can have deleterious effects. With high nutrient concentrations, the macroscopic and microscopic algae grow in large amounts and become abundant as attached epiphytes or free floating forms, reducing light penetration in the water

25 column (McKenzie, 2007; Austoni et al., 2007; Touchette, 2007). Increased epiphytic growth can result in up to 65% shading of the macrophyte leaves, reducing photosynthetic rate and leaf densities of the seagrasses (McKenzie, 2007). As the problem progresses, internal equilibria undergo a short circuit and the imbalance of phosphorus to nitrogen ratio can favour cyanobacteria and/or picoplankton species (Austoni et al., 2007) as well as nutrient + −2 toxicity (e.g. NH4 and NO ; Touchette, 2007). Macroalgal blooms have been recognized as one of the most catastrophic symptoms of community degeneration. In a study by Campbell and McKenzie (2001), investigations on seagrass meadows have been undertaken in Queensland in order to provide an early warning mechanism for detecting potential threats and damage to seagrass resources. A persistant and frequent abundance of filamentous epiphytic algae on seagrass in Whitsundays was detected and was cause for concern as these algae place at risk important productive seagrass meadows used for feeding by dugong and turtle populations. The excessive abundance of filamentous algae in Pioneer Bay appears to be detrimentally impacting the density of seagrasses in the region. Sewage derived nitrogen has been implicated in the growth of filamentous algae. There are, however, many factors that can lead to filamentous algal blooms and these include nutrient enrichment, organic enrichment and favourable light and temperature conditions (Campbell and McKenzie, 2001).

Other anthropogenic disturbances

A case of complete destruction of Zostera marina in the Thau lagoon (French Mediterranean Sea) has occurred and recolonisation of this macrophyte was studied (Plus et al., 2003). For many years, the Thau lagoon (south Mediterranean coast of France) suffered from anoxic events. The triggering factor was the degradation of green algae and probably organic matter coming from aquaculture, accelerated by high temperatures. Proportional oxygen saturation decreased to 0% in bottom waters during the episode, and toxic sulphuric compounds were released by the sediment (Plus et al., 2003). Subsequently, benthic vegetation which had been previously described as dense within the shellfish cultivation structures, comprising of Zostera marina meadows mixed with Gracilaria spp., Alsidium corallinum and Codium fragile populations, totally disappeared. Then, Plus et al. (2003) studied the recolonization of the eelgrass and found that the recolonisation took place surprisingly rapidly as biomass similar to those from untouched areas were reached only nine months after seed germination. The recolonisation success was partly due to a high seedling survival rate as well as a rapid vegetative recruitment (ranging from 0.012 to 0.042 per day). Two phases of recovery could be observed: a rapid multiplication of shoots during the first 3 months was followed by an increase in biomass due to elongation of leaves.

2.1.3 Reeds and Sedges

Occurrence and distribution

Reeds, sedges and rushes are usually rooted in soft intertidal or shallow subtidal subtrata. They have photosynthetic aerial portions, which are partially and/or periodically submersed (Adams et al., 1999) in fresh to brackish water. The species that usually occur in freshwater zones are: Schoenoplectus littoralis (Schrad.) Palla; Typha capensis (Rohrb.) N.E. Br.; Cyperus laevigatus L. and Schoenus nigricans L. In the mid-reaches of an estuary, on parts of marsh that are occasionally flooded, Bolboschoenus maritimus (L.) Palla and Juncus kraussii Hochst are usually found. Phragmites australis (Cav.) Trin ex Steud and Schoenoplectus triqueter (L.) Palla grow in brackish zones with salinities of less than 15 PSU. S. triqueter can grow in deeper water and at slightly higher salinities than P. australis (Adams et al., 1999). P. australis forms dense beds in the brackish upper reaches (<15 PSU) of South African estuaries that have a gradient of decreasing salinity up the length of the estuary (Adams et al., 1992; Adams and Bate, 1999a). In marine dominated estuaries, salinity is uniform (35 PSU) from the mouth to the head of the estuary. In these systems, P. australis beds have been found but only at the confluence of small freshwater streams and seepage areas flowing into the estuary. The plants are, however, tidally inundated with saline water (Adams and Bate, 1999a). This has been recorded in the marine dominated Kromme Estuary, in which the reed is tidally inundated by 35 PSU salinity water. This has been possible because it was also found that the root system was immersed in fresh or brackish water (Adams and Bate, 1994b). Groundwater seepage plays an important role in marine-dominated estuaries like these as it creates points of biotic diversity where both brackish and saline species may occur. The same findings were recorded in the Goukou and Keurbooms estuaries, where pore water salinity was lower than surface water salinity within Phragmites stands and a decrease in the height of P. australis was associated with an increase in the pore water salinity towards the water‟s edge (Adams and Bate, 1999a).

P. australis is a rhizomatous grass that can form either pure monospecific stands in wet areas, or it can associate with Scirpus sp., Schoenoplectus sp., Typha sp., and Cladium sp. Environmental factors that affect the performance of P. australis are salinity, thickness of the litter layer, water level regimes, nutrient status, aeration, burning, grazing and trampling (Haslam, 1971). Engloner (2009) reviewed about 190 publications on P. australis and still could not get a definite answer to the frequently asked question of whether internal (genetic) determinacy or external constraints (and what sort of them) have stronger effects on reed structure and growth. From the contradictory results in the literature the conclusion was that, the significance of these factors may vary from habitat to habitat (Engloner, 2009). Nevertheless, it seems to be clear, that increase of salinity diminishes the density, height, basal diameter and biomass of shoots, whereas increasing nutrient (or only nitrogen) availability generally increases the values of these characters. Furthermore, with higher nitrogen, the proportion of schlerenchyma is lowered and therefore, the lower the bending

27 strength. Thus, due to mechanical effects, culms break down easily in eutrophicated stands (Engloner, 2009).

Ecological importance

Reeds, sedges and rushes (emergent macrophytes) form an important component of the brackish and freshwater regions of estuaries, providing habitat for many birds, invertebrates and fish (Adams et al., 1999). The fish spawn and the birds nest and feed in these habitats. They also provide food for detritivores and form a substrate for periphyton and bacteria. They are important contributors of detritus in estuaries. In the Mhlanga Estuary, detrital aggregates from Phragmites have been found to be the most important food source for 90% of the fish community (Whitfield, 1980). P. australis and Schoenoplectus triqueter die back in late summer and end up releasing particulate matter into the water. Then, their decomposing products fertilise and enrich the water (Adams et al., 1999). They also provide a bank stabilization function due to their effectiveness in trapping sediment and can protect banks from erosion (Adams et al., 1999).

Emergent macrophytes also play an important role commercially. Due to their ability to remove nutrients; they have been used in wasrewater treatment systems. They have also been used in rural areas for the construction of mats and thatching material and for craftwork (Begg, 1986; Adams et al., 1999). In areas where high nutrient inputs are experienced the very lush growth of these macrophytes can be used to indicate riparian eutrophication (Brix, 1993).

Disturbances and threats

High salinity

Increased salinity, mostly resulting from freshwater abstractions in the riverine reaches, results in overall reduced plant performance and shoot height. This is due to a diversion of energy away from active meristematic growth to the maintenance of osmotic balance (Hellings and Gallagher, 1992). The survival rate of P. australis when inundated with saline water has been examined. The reed dies completely when inundated for 94 days at 30 PSU water (Benfield, 1984). In 20 PSU salinity water for two weeks, however, it gets stressed but recovers when returned to favourable conditions (Adams and Bate, 1994b). Laboratory studies by Adams and Bate (1999a) tested whether P. australis could survive tidal inundation with saline water (35 PSU) if its roots and rhizomes were located in freshwater (0 PSU). Plants which were supplied with freshwater to the roots but were tidally inundated with saline water had greater stem elongation and less dead leaves than plants that were supplied with a salinity of 20 PSU to the roots. The study thus concluded that P. australis will probably only survive intertidal flushing with saltwater if its roots and rhizomes are located in brackish, less than 20 PSU water (Adams and Bate, 1999a). It took Juncus kraussi 150 days in freshwater

28 to recover from an inundation of 15 to 30 PSU water (Heinsohn and Cunningham, 1992). On the other hand, deterioration of Typha domingensis was found to be evident in the 7 to 10 PSU salinity range (Glenn et al., 1995).

Sedimentation

Reed encroachment has been recorded in a number of South African estuaries. According to Riddin (1999), this is probably due to catchment mismanagement leading to erosion and the resulting high sedimentation rates. Also, in North American salt marshes, Phragmites australis has been replacing other vegetation at a rate of between 1 and 6% of marsh surface area per year (Weinstein and Balletto, 1999). Large areas now have monotypic stands of this reed and this has led to difficult access to the marsh by nekton, fragmentation of existing stands of other salt marsh plants (Weinstein and Balletto, 1999; Riddin, 1999), and the reduction of bird species diversity (Benoit and Askins, 1999). Since emergent macrophytes are able to trap sediment, many of the smaller KwaZulu-Natal systems have been seriously affected by increasing rates of sedimentation (Begg, 1978). In a more recent study, Riddin (1999) also found that increased sedimentation rate in the Siyaya Estuary has resulted in complete colonization of the estuary by the highly productive reeds.

Sediment and associated nutrient characteristics determine species composition and distribution of emergent macrophytes (Riddin, 1999). Studies have shown that fine-textured sediments provide a proportionally greater aboveground biomass than coarser sediment (Barko and Smart, 1978) and this was possibly in response to the greater nutrient availability associated with the finer sediments. Conversely, plants in sandy sediment had increased below ground biomass thereby increasing the absorptive surface and/or storage mass.

Nutrients

Nutrients have been shown to have a great impact on the health of emergent macrophyte stands. A linear relationship has been found between shoot height and nutrient availability in P. australis (Haslam, 1971). Stands associated with nutrient deficient substrates usually exhibit shorter height, decreased shoot density and diameter, and lower competitive powers (Haslam, 1971). Nutrients trapped in fine textured sediment have contributed towards greater above ground biomass of Cyperus esculentus and Scirpus validus than coarser sediments have (Barko and Smart, 1978). This was because sandy sediments have a lesser ability to trap nutrients than fine sediment. Thus, caution should be taken towards increased anthropogenic nutrient increases, which end up being trapped in sediments and can contribute to monospecific stands of emergent macrophytes and reduced species diversity. Not only the macrophytic plants would be affected by this, but also the fauna that benefit from diverse habitats as was mentioned by Benoit and Askins (1999).

2.1.4 Phytoplankton

Research on South African estuaries has mostly focused on phytoplankton (suspended in the water column) and benthic microalgae attached to sand grains (episammic algae). The other microalgae are usually attached to the bottom sediments (epipelic), rocks (epilithic), and other plants below the water surface (epiphytic) (Adams and Bate, 1999b). For the purposes of the current study, the focus will be on the planktonic, benthic and epiphytic microalgae.

Phytoplankton are microscopic algae that are the primary food and oxygen producers within freshwater, estuarine, and marine habitats (Ohrel and Register, 2006). They lack the means to counteract transport by water currents. Phytoplankton are dominant in large channel-like estuaries that are characterized by large catchment areas and high freshwater inputs, like the Sundays and Gamtoos estuaries (Adams and Bate, 1999b).

Occurrence and distribution

Two important factors are recognized as controlling community structure of phytoplankton. The first is related to physical processes such as mixing of water masses, light, temperature, turbulence, and salinity, and the second is associated with nutrients. Phytoplankton community composition changes as a consequence of species succession, which occurs in response to new conditions encountered in the environment (Huisman et al., 2001).

According to Adams and Bate (1999b) phytoplankton are dominant in large channel-like estuaries with large catchments and high mean annual runoff and where freshwater introduces nutrients and creates stratified conditions. These microalgae are suspended horizontally and vertically in the water column with their spatial distribution controlled by water motion (Adams and Bate, 1994b; 1999b). Under calm conditions the phytoplankton may settle out at the bottom and take up the benthic habitat (Adams and Bate, 1999b).

The abundance of phytoplankton in the water column is based on seasonal patterns related to nutrients, light intensity, temperature, and grazing (predation) pressures. Monitoring the types and relative abundances of plankton populations, in conjunction with nutrients and other environmental parameters can provide significant insight into the health of an aquatic ecosystem (Ohrel and Register, 2006). The main environmental factors recognized as controlling community structure of phytoplankton are physical, (mixing of water masses, light, temperature, turbulence and salinity) and chemical (nutrients) (Brogueira, 2007).

Common phytoplankton groups are diatoms, dinoflagellates, coccolithophorids, cryptomonads, green algae, blue-green algae, red algae and brown algae. Day (1981) stated that the distribution of these phytoplankton groups in estuaries differs according to salinity and temperature, and also based on nutrients (Ohrel and Register, 2006). In the lower reaches diatoms dominate the phytoplankton community and dinoflagellates are only

30 important at certain seasons. Small nanoflagellates are usually abundant in the upper reaches (Day, 1981). Additionally, according to Ohrel and Register (2006), waters having relatively low nutrient levels are dominated by diatoms, which are a highly desirable source of food. In water with higher nutrient concentrations, cyanobacteria and dinoflagellates become more abundant. These phytoplankton species are less desirable as a food source to animals (Ohrel and Register, 2006). Furthermore, and contrary to Ohrel and Register‟s findings, Margalef (1978) states that diatoms are favoured in nutrient-rich, well-mixed waters during spring tides. And similarly, the dinoflagellates were also found in nutrient-rich waters, but when there is a stable stratified water column. The flagellates would then increase in abundance when the stratification has set in and the water has become depleted of nutrients (Margalef, 1978). In a study by Adams and Bate (1994b), in all the estuaries that were studied (Berg, Palmiet, Goukou, Gourits, Great Brak, Keurbooms, Gamtoos and Sundays) the phytoplankton community was dominated by flagellates. In the study the concentration of nitrate, which was found to be the favoured nutrient when it comes to phytoplankton biomass, varied greatly among these estuaries and so the nutrient levels could have not determined the species composition. However, all the estuaries displayed horizontal and vertical salinity gradients, which must be an indicator that the estuaries provided the stratified condition that flagellates relate to. So, from this study it was concluded that phytoplankton biomass relates better to water column nitrate concentration and the salinity gradient determines the species distribution (Adams and Bate, 1994b).

South African studies have shown that phytoplankton biomass in estuaries was positively correlated with freshwater input; these are studies that have been done on the Kariega, Keiskamma, Great Fish (Allanson and Read, 1995), Swartkops, and Sundays estuaries (Hilmer, 1984; Hilmer, 1990; Hilmer and Bate, 1990). From these studies it was concluded that the phytoplankton communities are supported by the freshwater input because it supplied nutrients and maintained stable stratified conditions. However, Adams and Bate (1999b) mention that “light reduction from heavy silt load, prevention of nutrient addition to the surface layer by strong vertical stability, or rapid advection such that the flushing time exceeds the time scale of phytoplankton growth, will act to reduce phytoplankton production”. At the same time, “in regions where there are sufficient surface nutrients, vertical stability imposed by runoff helps to maintain phytoplankton in the euphotic zone thereby enhancing production” (Adams and Bate, 1999b).

More studies conducted by Adams and Bate (1994b) on the importance of freshwater for phytoplankton in the following South African Cape estuaries: Berg; Palmiet; Goukou; Gourits; Great Brak; Keurbooms; Gamtoos; and Sundays showed that in the Sundays Estuary high phytoplankton biomass was maintained if the nitrate concentration was greater than 200 µg.l-1. This was related to a consistent input of freshwater, high in nutrients, which was the result of agricultural fertiliser runoff from the catchment area. It was also evident that the retention time of water played a major role in phytoplankton productivity (Hilmer, 1990). Blooms have occurred in the estuary, in the upper reaches, when there was a water residence

31 time of 7 neap tidal cycles and at least 3 spring tidal cycles (MacKay and Schumann, 1990). These calm conditions were favourable for phytoplankton bloom formation because it allowed the phytoplankton time to take up the nutrients. These findings were also found for the Great Fish Estuary; where freshwater input results in nutrient input into the estuary and thus an increase in phytoplankton biomass. In other estuaries with no agricultural impacted catchments freshwater input results in high nutrients when flood pulses of water enter the estuary. This was evident in systems such as the Great Brak and Keurbooms estuaries. Freshwater pulses in these estuaries boosted phytoplankton biomass as they brought in nutrients. However, this was only a short term response. After the freshwater pulse the mean chlorophyll-a content increased from 0.2 to 13.3 and from 0 to 13.3 µg.l-1 in the Great Brak and Keurbooms, respectively. The corresponding mean nitrate was 10.6 – 23.4 µg.l-1 (Great Brak) and 4.3 – 61.1 µg.l-1 (Keurbooms).

When there are adequate nutrients, increased light intensity, warmer water, and minimal predation pressures from zooplankton, phytoplankton population explosions, or algal blooms, may occur. The phytoplankton will continue to bloom until one or more of the key factors promoting phytoplankton growth are no longer available (Ohrel and Register, 2006). This is usually not good for the health of estuaries.

Phytoplankton species distribution has been found to respond to salinity gradients. Stable stratified salinity conditions seem to be favourable for bloom formation (Hilmer and Bate, 1991), especially by the red tide species of the dinoflagellate group. This has been found in the Sundays and Gamtoos estuaries and these estuaries experience nutrient input from agricultural lands. This bloom formation could be toxic to other flora and fauna of the estuaries, as it is an indication of eutrophic conditions and thus anoxia developing. The dinoflagellate blooms in the Sundays Estuary were depressed during spring tides because vertical mixing occurred and dispersed the salinity gradients (Hilmer and Bate, 1991). Thus, in order for conditions like this to not persist, there should be a strong interaction between tidal cycles and river inputs in estuaries, otherwise irreversible eutrophic, detrimental conditions could persevere.

Ecological importance

“Phytoplankton is a fundamental actor in global biogeochemical processes, participating in the transformation and cycling of key elements” and, also, can affect turbidity, oxygen depletion, and the total productivity of the system (Los and Wijsman, 2007). Phytoplankton has been used extensively as a gauge of ecological condition and change. They have tremendous impacts on water quality and play a number of other major roles in many ecosystem processes (Los and Wijsman, 2007).

Its critical ecological function is being the primary producer that directly and indirectly fuels food webs; being the basis of the food web in estuaries. Without phytoplankton, the complex

32 web of estuarine plants and animals would collapse (Ohrel and Register, 2006). As photosynthesizers, they transfer the sun‟s energy into plant matter and provide nourishment for the next level of organisms (zooplankton), which in turn are consumed by other larger organisms. If the phytoplankton community is altered in composition or abundance, these changes may have serious ramifications throughout the food web and upset what may be considered a more favourable balance of life in these waters (Ohrel and Register, 2006).

Due to its fast population responses to changes in water quality, hydrology or climate, on top of being a widely used indicator of nutrient input changes, it is also effective in evaluating responses to many other environmental stressors (Domingues et al., 2008). Since different human activities in coastal areas generate wastes that cause changes in the natural hydrological conditions of the coastal system, inducing eutrophication, structural changes of the phytoplankton community are a good indicator of eutrophication effects because phytoplankton composition responds to fluctuations in environmental conditions (Livingston, 2001).

Phytoplankton influence oxygen concentrations in the estuary. During the day, when they photosynthesize, they produce oxygen, which is critical to all but a few estuarine organisms. When sunlight is unavailable (e.g. at night), phytoplankton respire, removing oxygen from the water. Oxygen is also consumed when bacteria work to decompose phytoplankton, which is a common outcome of excessive phytoplankton growth in estuaries. This in turn produces hypoxic conditions that may result in the deaths of many organisms (Ohrel and Register, 2006).

Disturbances and threats

Freshwater input

There are some anthropogenic activities that lead to a continual elevation of freshwater input and velocity, i.e. wastewater discharging activities. Increased freshwater input can be good for the productivity of an estuary because it brings in nutrients; however, it may also decrease biomass because of dilution and reduced residence time (Adams and Bate, 1999b). High flow rates can also be detrimental to phytoplankton by creating a turbid environment, which inhibits light penetration (Westeyn and Kromkamp, 1994). Due to its steep escarpment and narrow coastal plain, which results in many short and steep rivers, South Africa has many small estuaries. These are rapidly flushed by either river or seawater and thus, residence times for phytoplankton populations are subsequently low (Walker, 2003). Snow et al. (2000a) recorded that in the Gamtoos Estuary, maximum chlorophyll-a concentrations were found during times of relatively low flows (0.75 to 1.2 m3.s-1). Freshwater inflow into the Gamtoos is highly nutrient-enriched by run-off from agricultural lands, and the low flow increases the residence time of this enriched water and allows for phytoplankton to absorb the nutrients and subsequently increase in biomass. Conversely, when freshwater was released from the Mpofu

Dam to feed the freshwater-starved Kromme Estuary, no significant increases in phytoplankton biomass were recorded (Snow et al., 2000b). The conclusion from this study was, “the length of time that the freshwater influence was present prevented a significant increase in microalgal growth”, because the freshwater pulse came at very high speeds. Snow (2000) found that in the Gamtoos Estuary, low phytoplankton biomass was recorded at the flow rates that were greater than 2.3 m3.s-1 (the biomass was lower than 20 µg.l-1), while the greatest biomass was measured at 0.8 and 1.2 m3.s-1 (biomass of 47.5 and 49.9 µg.l-1 were measured, respectively), thus indicating that the ideal freshwater inflow rate was around 1 m3.s-1.

Increased nutrient inputs

In recent years, there has been an increasing presence of bloom-forming algae in estuaries worldwide. This has been attributed to increased levels of nutrients entering these waters. High nutrient inputs are usually due to an uncontrolled human activity; like leakage of sewage pipes into the system or agricultural return flow with high nutrient levels or deliberate discharges of sewage and industrial waste. As nutrient levels increase and the phytoplankton concentrations become denser, the water often takes on the colour of the algal pigments (e.g., reddish brown, green, brown) and odours become noticeable (Ohrel and Register, 2006). This condition is then called eutrophication, which tends to be unfavourable for the other living organisms in that estuary. This is because phytoplankton blooms lead to shading out of the water column, death of other organisms and degradation of these by bacteria and eventually to the depletion of oxygen in the water column and anoxic conditions. Specifically, this high nutrient input could be from increased nutrient loading in the freshwater input (e.g. due to inputs through human activities), increases in the quantity of freshwater input (e.g. high rainfall events, high return flow from agricultural lands, or transfer schemes), or increases in fresh/seawater flow rates such that benthic regenerated nutrients are mixed into the euphotic zone (Flint, 1985).

2.1.5 Benthic microalgae

Occurrence and distribution

Benthic microalgae dominate in estuaries with large intertidal areas and contribute to primary production in these estuaries (Adams and Bate, 1999b). It has been shown that in most South African estuaries the microphytobenthic biomass could even contribute much more than phytoplankton to primary production and its importance is now being recognized (Rodriguez, 1993; Walker, 2003). In Rodriguez‟s study, it was only in the Sundays Estuary that the total phytoplankton biomass exceeded the microphytobenthos (Table 2.1), and this could be due to this estuary being channel-like and lacking adequate exposed intertidal areas for the benthic microalgae to colonise. A variety of physical, chemical, and biological factors may regulate the standing stock of benthic microalgae in shallow aquatic ecosystems. Assessment of the 34 effects of these factors is thus important to understand and manage these ecosystems (McIntyre et a1., 1996). Since many benthic diatom species are cosmopolitan, there have been considerations that estuarine species tolerate a broad range of environmental conditions

Table 2.1: The comparison of phytoplankton and microphytobenthos biomass (as kg chlorphyll- a/estuary) for selected South African systems (adapted from Rodriguez, 1993)

Whole estuary phytoplankton Whole estuary benthic Estuary chlorophyll-a (kg) chlorophyll-a (kg)

Goukou 0.07 23 Gourits 0.04 16 Keubooms 0.06 20 Gamtoos 17 14 Sundays 86 14

Although a number of algal divisions are found in the microphytobenthic community, the most dominant group is usually diatoms (Rodriguez, 1993; Barranguet et al., 1998; Brotas and Plante-Cuny, 1998). Rodriguez (1993) found 75% of the microphytobenthic community to be consisting of diatoms in the Swartkops Estuary. The composition of the diatom community may however change along the length of the estuary and with the level of nutrient loading in a system (Walker, 2003). Some studies have shown that an increase in riverine nutrient input leads to an increase in benthic microalgal biomass. The phytoplankton community quickly absorbs these nutrients, however, and eventually the benthic microalgae become shaded out by phytoplankton blooms (Adams and Bate, 1999b). Nevertheless, it has been noted that even under conditions of low nutrient availability in the water column, benthic microalgae can survive by benefiting from the relatively high nutrient concentrations in the pore water of sediment (Meyer and Meyer-Reil, 1999). Thus, water column nutrients usually have no direct relation to benthic microalgal biomass.

Alternatively, Underwood et al. (1998) found relations between the distribution patterns of some epipelic diatom taxa and salinity and nutrient gradients, particularly ammonium. Underwood et al. (1998) found that some diatom species were located and abundant according to the estuarine salinity and nutrient gradients along the Colne Estuary. Navicula phyllepta and Navicula gregaria were abundant at the meso- and oligohaline sites respectively and Pleurosigma angulatum and Plagiotropis vitrea were found at the polyhaline sites. Species that were found at the seaward end were Nitzschia frustulum, Cylindrotheca signata and Navicula pargemina. Findings from a study by Minne (2003), where a number of eastern and western Cape estuaries in South Africa were investigated (including the Bushmans Estuary), also showed a division of the diatom species into four groups according to water column salinity and nutrients. The data was presented with a Canonical Correspondence Analysis (CCA). The groups were divided into species associated with high salinity (Amphora arcus, Amphora coffeaeformis, Amphora sublaevis, Cylindrotheca closterium, Entemoneis paludosa, Gyrosigma prolongatum, Halsea crucigera,

Haslea ostrearia, Navicula salinicola, Navicula tenelloides, Nitzschia angularis, Opephora minuta, Plagiotrphis tayrecta, Planothidium delicatum and Seminavis species 2), low salinity (Nitzschia littoralis), high ammonium (Amphora suacutiuscula), high nitrate and soluble reactive phosphorus (Achnanthes delicatula, Amphora helenensis and Navicula perminuta). However, statistically, only salinity and ammonium had significant effects on the distribution of the diatom taxa; the nitrate and soluble reactive phosphorus had little effect. Most of the Bushmans Estuary diatom taxa were indicative of high salinity and high ammonium.

Other factors that control benthic biomass are water currents, light flux, depth, physical disturbance, grazing, sedimentary composition and disturbances and water speeds (McIntyre et al., 1996; Adams and Bate, 1999b). In a survey of a number of South African estuaries, Rodriguez (1993) found no clear pattern when comparing subtidal with intertidal sites. Nel (1998), however, found that in the Gamtoos Estuary intertidal biomass was two to three times higher than subtidal.

Sediment type influences microphytobenthic biomass (Walker, 2003) because it influences nutrient availability for microphytobenthic organisms. Fine cohesive estuarine sediments usually have high organic matter content, with high rates of bacterial mineralization and high porewater concentrations of dissolved nutrients, while sandflats are more oligotrophic (Admiraal, 1984). Longitudinally, the distribution of the microphytobenthic biomass was found to be strongly influenced by position in the Kwelera Estuary by Walker (2003). It was concluded that the more turbulent hydrological conditions and loose sediment in the mouth area did not allow biomass to develop. The distribution of biomass showed a consistent pattern of low level in sandy sediments near the mouth and higher concentrations at the muddier stations (Walker, 2003). However, when these findings were compared to results from a Swartkops study (Rodriguez, 1993), it was found that in the Swartkops Estuary the benthic microalgal biomass was high in the sandy sediments. Walker (2003) thus concluded that the level of exposure also influenced benthic microalgae. The microalage in the Swartkops Estuary were found in the sandy sediment because it occurred in areas where there was low tidal action and thus the sediment was stable. Due to the coarser nature of sand, high absorption of light could have also played a role. The microphytobenthos tend to be dominant in the upper 1 cm euphotic zone, according to Rodriguez (1993), and benthic microalgae have the potential to remain dormant in the dark until favourable conditions prevail (Sundbäck and Granéli, 1988). When sediment containing microalgae was exposed to the dark, the chlorophyll-a content only decreased slightly and then remained constant for several weeks. The chlorophyll-a content increased rapidly when this sediment was re-exposed to light (Sundbäck and Granéli, 1988).

There are processes that may disturb the benthic community and temporarily lead to the benthic flora taking up a planktonic habitat (Baillie and Welsh, 1980; Adams and Bate, 1999b). These disturbances could be wind mixing and tidal currents. It appears that biomass is higher in sheltered, muddy habitats than in exposed, sandy regions (Miller et al., 1996).

Sandy sediment is more „loose‟ and can easily be moved by tidal currents and thus forms an unstable environment for microalgae to attach and settle in. Therefore, shelter may be an important factor in their distribution. Light limitations associated with increased wave height as well as the physical disturbance caused by waves are the main cause of reduced biomass, (Fielding et al. 1988, Rodriguez 1993). Desiccation, as a result of a drop in water level, has a serious effect on the succession of benthic algal communities. Some diatom taxa have, however, been reported to be able to resist long periods of desiccation (van der Molen, 2000).

Water speed determines sediment texture, which controls the accumulation of nutrient-rich organics. Higher flow rates lead to coarser, nutrient-poor sediments, while mud can sustain twice as much benthic microalgal production because it retains more nutrients (Adams and Bate, 1999b). Diatoms can attach themselves more easily to mud than sand and the solid environment provided by mud mitigates against mechanical damage (De Jonge, 1985). Davies and McIntire (1983) also found that biomass increased in sites with siltier sediment opposed to sandy sediments. A recent study on the Keurbooms, Mngazana, Gamtoos, Swartkops, Sundays and Mngazi estuaries by Snow (2007) showed that the microphytobenthic biomass had the strongest associations with sediment related variables. These were high organic content (> 3%) and fine sediment content (< 125 µm sediment contributes more than 20% of the sediment). These findings were similar to that reported for other systems. Benthic microalgal biomass was sampled in sediments collected from two sets of North Carolina estuaries, Massachusetts and Cape Cod bays, and Manukau Harbour in New Zealand. Comparisons of benthic microalgal biomass and sediment grain-size distributions in these coastal and estuarine ecosystems frequently showed a negative relationship between the proportion of fine-grained sediments and benthic microalgal biomass measured as chlorophyll-a. The highest sedimentary chlorophyll-a levels generally occurred in sediments with lower percentages of fine particles (diameter < 125 mm) (Cahoon et al., 1999).

Ecological importance

Benthic microalgae are now recognized as significant primary producers in shallow aquatic ecosystems (McIntyre et al. 1996). Their production and biomass can equal or surpass that of phytoplankton in the overlying water column. Rajesh et al. (2001) found the total annual production contributed by benthic microalgae to be more than that of the overlying water column (33.59 and 10.51 g C.m-2 respectively). Case studies on selected Cape estuaries show that the highest intertidal microalgal biomass was found in the Sundays Estuary (197 mg.m-2) and the highest subtidal in the Goukou Estuary (205 mg.m-2). In most of the estuaries the benthic microalgal biomass was higher than that of the phytoplankton, except for the Sundays and Gamtoos estuaries. Both these estuaries are surrounded by agricultural-dominated catchments and therefore had more phytoplankton biomass because of the fertiliser-enriched freshwater input (Adams and Bate, 1994b).

Benthic microalgae are likely to be important in nutrient cycling, and may act to control nutrient availability to other primary producers (Cahoon et al., 1999). They cycle nitrogen within the sediments by evolving oxygen during photosynthesis that inhibits denitrification (Dong and Underwood, 1998). Oxygen production by benthic microalgae also prevents the release of phosphate and ammonia from the sediment (Sundbäck and Granéli, 1988). Their photosynthetic activity has also been found to stimulate nitrification (conversion of ammonia to nitrate) in the sediment (Thornton et al., 1999). They support deposit feeders directly (Connor et al. 1982; McIntyre et al., 1996) and suspension feeders through resuspension by physical processes (Baillie and Welsh, 1980).

As producers of new organic matter, microphytobenthos can enter the benthic and pelagic food web; thus they form the key component of the carbon cycle in benthic environments. Furthermore, in shallow aquatic ecosystems, benthic microalgae provide a major food source for meio- and macro-invertebrate grazers such as annelids, nematodes, flat worms, crustaceans, mollusks and some demersal fishes and larvae (Rajesh et al., 2001). Finally, they play an ecological role in the stabilization of sediments through the production of extracellular polymeric substances during locomotion (Miller et al., 1996).

Disturbances and threats

Anthropogenic disturbances are increasingly placing the goods and services of the aquatic benthos under threat. The primary determinants of benthic microalgal composition and abundance (i.e. water depth, flow, sediment chemistry and grain-size) are being altered by the dredging of channels, the disposal of spoil, the discharge of pollutants and increased sedimentation from the land (Bishop, 2007).

Sediment disturbance

Benthic organisms are adapted to the natural processes of sediment movement, erosion and deposition (Turner and Miller, 1991; Miller et al., 1992). However, when sedimentation exceeds natural thresholds, then impacts may involve total loss of the community and subsequent colonization by pioneer species (Alongi, 1997; Miller et al., 2002). Extreme sediment disturbance events can result from man‟s modifications of the aquatic environment, and the scale and magnitude of these alterations can often greatly exceed that of natural occurrences. Probert (1984) listed the following human activities as being of great concern when it comes to disturbances on the benthic environment: dredging; spoil and mining waste disposal; marine mining; organic pollution; oil pollution; and bait digging.

Dredging activities for the purpose of opening estuarine mouths have become customary in temporarily open/closed estuaries for various reasons. Such disturbances result in flushing and sediment scouring, reducing the microalgal biomass stock. Anandraj et al. (2008)‟s study investigated the recovery of microalgal biomass and production following a breaching event

38 and the key environmental parameters influencing primary production during the open and recovery phases. This study established that the disturbance of the breaching event in the Mdloti Estuary drastically altered the dynamics of both benthic and pelagic microalgae. Following the breaching of the mouth, almost the entire stock (94–99%) of estuarine microalgal biomass was washed out to sea. However, after some time, the open mouth phase significantly stimulated the benthic microalgal production by almost an order of magnitude compared to the closed phase. This recovery of estuarine primary production and biomass appeared to be primarily influenced by optimal environmental conditions that prevailed during the open phase compared to the closed period, with the shallow waters (< 1 m) contributing to favourable light, temperature and nutrient levels for the benthic habitat (Anandraj et al., 2008). It should most importantly be noted that continuous flushing and sediment scouring prevent the buildup of microalgal biomass. Thus dredging, which is a proposed activity for the Bushmans Estuary, should be strongly regulated.

Licursi and Gómez (2009) assessed the effects of dredging on the structure and composition of diatom assemblages in a lowland stream (Rodrıguez Stream in Argentina). The results of the study showed that the effects of dredging in the stream involved two types of disturbances: the removal and destabilization of the substrate in the stream bed; and chemical changes and an alteration of the light environment in the water column. The physical and chemical modifications, after dredging, in the benthic habitat resulted in an immediate increase in the diversity and species numbers of the benthic diatoms, which decreased at the end of the study period. Also, in the post-dredging period, the sensitive benthic diatom species were replaced by species that seemed to be tolerant to organic pollution and eutrophication, resulting from the resuspension of sediment-trapped nutrients and organic matter due to the dredging (Licursi and Go‟mez, 2009).

Excessive sediment deposition can lead to burial, smothering or crushing of benthic organisms. Conversely, erosion removes sediment and organisms (Hall, 1994; Thistle et al., 1995). Both deposition and erosion can result in a change of bottom sediment level and possibly sediment grain size. Materials placement, for example dredge material disposal, should allow community responses to follow natural seasonal and successional trends and to exhibit minimal anthropogenic impacts (Miller et al., 2002). Unfortunately, there is little of the quantitative information necessary for predicting how materials placement, sediment deposition and erosion will affect the ecology of these environments.

Studies were conducted on the impacts of the introduction of the bivalve Tapes philippinarum in the Venice lagoon and its harvesting by hydraulic and mechanical dredges, which strongly increased the amount of sediment resuspension and settlement (Facca et al., 2002). This activity caused a marked increase in water turbidity and the disruption of the benthic microlayer of the lagoon bottoms. In the areas affected by the highest sedimentation fluxes, resuspension of many benthic taxa, such as Amphora, Cocconeis, Navicula, Nitzschia, Pleurosigma and Thalassiosira, occurred and so they were found in all the water column

39 samples and they were more abundant than exclusively planktonic diatoms (Facca et al., 2002).

Boating, like many other human activities, represents a major disturbance to the benthic habitat. There is a high susceptibility of sediment erosion due to boat activities, which cause waves that disturb intertidal sediments. Boating impact, through the suspension of sediment, can result in major alteration of bottom morphology and sediment grain-size (Osborne and Boak, 1999), resulting in sizeable effects on assemblages of the macrobenthos (Bishop, 2004). The disturbance and resuspension of the sediment are the major impacts, of which, in return decrease the biomass of the standing stocks of the benthic microalgae, due to physical removal and increases in turbidity.

Sediment contamination

Land-based nutrient pollution represents a significant human threat to aquatic environments. The findings of lab experiments on the reaction of benthic diatoms to artificial eutrophication showed a decrease of diversity and evenness with colonization time while species numbers initially increased but subsequently reached a plateau (Hillebrand and Sommer, 2000). This outcome was attributed to an increase in dominance of few species outgrowing their competitors (i.e. lowered evenness). The enhanced nutrient supply decreased microalgal diversity by increasing the dominance of single species in eutrophicated habitats. And thus, only few microalgal species respond significantly to nutrient enrichment.

Compared with macroorganisms, microorganisms are more sensitive to toxin contamination because their surface/volume ratio is higher and thus exposure to toxicants is more direct. Moreno-Garrido et al. (2007) conducted an investigation of the response of diatoms to sediments with different levels of pollutants, collected from the Aveiro Lagoon (Portugal). Benthic diatoms seem to be sensitive organisms in sediment toxicity tests. Among the microalgal species used, Cylindrotheca closterium seems to be a little more sensitive, although populations of all assayed species show comparable pattern responses when exposed to the different sampled sediments. The other species were Phaeodactylum tricornutum and Navicula sp. Concentrations of Sn, Zn, Hg, Cu and Cr (among all physico- chemical analyzed parameters) in the sediment had the most significant impacts on the benthic diatoms.

The results of an analysis done on the benthic diatoms of a metal-polluted stream in the Riou Mort watershed in France also showed that the diatom communities were negatively affected by the metal accumulation (consisting of cadmium and zinc) in the sediment (Morin et al., 2008). Morphological abnormalities were particularly evident in the genera Ulnaria and Fragilaria while the rest of the diatom community displayed induced tolerance, seen through structural impact and dominance of small, adnate species. This led to the reduction in species diversity as the species assemblages were characterized by taxa known to occur in metal-

40 polluted environments (i.e. high numbers of Eolimna minima, Nitzschia palea, Pinnularia parvulissima, Surirella brebissonii, Achnanthidium minutissimum, Navicula lanceolata and Surirella angusta were recorded).

2.1.6 Epiphytic microalgae

Occurrence and distribution

The presence of submerged macrophyte hosts is the most important factor that determines the overall epiphytic biomass of the estuary as they would determine the amount of suitable habitat on which epiphytes would attach (Gordon et al., 2008). Epiphytic algae occurring on macrophytes, both emergent and submerged, consist mainly of cyanobacteria, diatoms, crustose and ephemeral algae (Ballentine and Humm, 1975; Novak, 1984; Bologna and Heck, 1999). However diatoms form the dominant group on submerged macrophytes (Allan, 1995; Coleman and Burkholder, 1995). Colemen and Burkholder (1995) found that the diatoms were dominant under high nutrient enrichment and when the nutrient levels were low, red algae were dominant.

According to Muller (1999) the epiphytic community is divided into two main components: the adpressed and the upright components. The composition of the adpressed component is mainly by algae that position themselves against the leaf surface of the host plant (e.g. Cocconeis scutellum Ehrenberg), while filamentous algae, such as Cladophora, mostly comprise the upright component because they extend above the leaf surface of the host plant. On the top layers of the epiphytic community, some diatom species of the microphytobenthos and plankton community have been found and are considered to be transitory species that have settled due to reduced water flow rate (Meulemans and Roos, 1985; Muller, 1999). Examples of these are species from the free-moving, non-sessile Nitzschia, Navicula, Stephanodiscus and Cyclotella genera.

Factors that have been found to affect the epiphyte community are nutrients, light, grazing, water movement, and the biomass and diversity of the species are affected by the length of the time the host has been colonised by the epiphytes (Allan, 1995; Wright et al., 1995; Muller, 1999). Wright et al. (1995) found that the older leaf segments of the eelgrass Zostera marina had the greatest epiphyte biomass. Additionally, Wright et al. (1995) found that in estuaries with high nutrient loads, high epiphyte biomass was found. Nutrients also play a very important role in determining species composition (Wear et al., 1999). However, the extent to which the nutrients can affect the epiphytes depends on the external rate of nutrient supply, the degree of internal nutrient cycling and on environmental controls like light, temperature and water velocity (Worm and Sommer, 2000). Coleman and Burkholder (1995) demonstrated that epiphytic growth occurred immediately after a short-term nutrient enrichment under low flow conditions, but when the flow rates and nutrients were high epiphytic production was not stimulated. Additionally, epiphytes take nutrients up at higher

41 rates than the host because of their large surface area to volume ratio. Thus, epiphytes respond immediately to nutrient increases compared to submerged macrophytes and can indicate eutrophic conditions immediately through increased growth rates (Worm and Sommer, 2000).

Salinity affects the species composition of the epiphyte community and the submerged macrophyte hosts and cannot be directly correlated to reductions or increases in epiphytic biomass itself (Snoeijs, 1999). Within epiphytic communities, species would be affected by salinity according to their individual tolerance ranges (Gordon, 2006). However, salinity can indirectly affect the epiphytes through its effect on the host plant, since submerged macrophytes are greatly affected by salinity (Adams, 1994). Grazing could also indirectly affect the community biomass by affecting the species composition and species-specific production rates of the community (Coleman and Burkholder, 1995).

Light is another important factor that can control the biomass and composition of epiphytes. The biomass usually increases with increases in the mean daily irradiance until a saturation point is reached, where the rate that the biomass increases starts to decrease (Nelson and Waaland, 1997). The species composition of the epiphytic community is affected by light because the optimum irradiance level for different species varies (Patrick, 1977; Ruiz and Romero, 2001).

Flow rate could have a large effect on the distribution of epiphytes on the host and along the estuary. At higher flow rates the epiphytic community is usually dominated by smaller, adpressed epiphytes. And, on the host, at the apices the smaller epiphytes are found because that is where flow rate has the most effect and then the taller epiphytes would occur at the base of the host plant (Bergey et al., 1995).

The water column environmental variables can differ greatly in relation to the host leaf surface area. Also, each leaf of the host differs relative to its position or contact with the water column, sediment, desiccation and shading. Thus, epiphytic species composition differs greatly and is highly variable between individual hosts of the same species and between different species (Gordon, 2006). Gordon et al. (2008) found great variation in epiphyte biomass between the different sites and at the same sites over time in the St Luica Estuary and attributed this to the rapid change in environmental conditions and submerged macrophyte area cover. The submerged macrophyte beds would be exposed during one sampling period and inundated during the next depending on water level. Other studies have also indicated this epiphytic biomass variation. For example, Cattaneo (1983) found epiphyte biomass levels to fluctuate between 23 and 25 mg Chl a. m-2 in spring and autumn, while summer biomass levels were much lower, i.e. 3 mg Chl a. m-2.

In Gordon et al.‟s (2008) study, there were several factors that correlated to the distribution of the dominant diatom species. Nitzschia frustulum showed a significant correlation to water

42 column electrical conductivity (0.56, p < 0.01). N. frustulum is a brackish to marine diatom species normally found between 27 and 33 PSU with high requirements for chloride ions, medium to high requirements for potassium, sodium and magnesium ions and low to medium requirements for ammonium and silicate (Bate et al., 2004). Cocconeis placentula showed a strong correlation with dissolved oxygen concentrations within the water column and was found in the sites where the percentage oxygen saturation was high. This is a freshwater and brackish species, occurring at salinity values from 7 to 28 PSU (Sims, 1996; Bate et al., 2004). Navicula ramosissimo was the only dominant diatom species that showed a positive correlation with TOxN concentration. In this estuary, the epiphytic biomass and the composition of the diatom species did not show any strong statistical relationships with the environmental variables measured and thus it was concluded that biological factors, such as grazing and competition, may be more important than the physico-chemical environment in determining epiphyte biomass distribution (Gordon et al., 2008).

Ecological importance

Epiphytic algae on submerged macrophytes such as Zostera marina L. (eelgrass) have functional significance and they can act as indicators of biodiversity or ecological status. Although estimates of epiphyte productivity are variable, primary production by attached algae has been shown to represent between 22 and 61% of productivity in Zostera beds (Hemminga and Duarte, 2000). Epiphyte communities may act as indicators of anthropogenic environmental impacts (Coleman and Burkholder, 1995) because their overgrowth can reduce Zostera productivity (Sand-Jensen, 1977), potentially leading to degradation and loss of beds. And, in some cases, the process of overgrowth has been accelerated by eutrophication (Borum, 1985; Hauxwell et al., 2003).

Epiphytes form an important linkage in the food web structure of estuarine ecosystems (Humm, 1964; Penhale, 1977; Borum et al., 1984; Moncreiff et al., 1992; Coleman and Burkholder, 1995; Winning et al., 1999) due to their nutritional value which is higher than that of the submerged macrophytes themselves. Under the most suitable environmental conditions, epiphytes have the potential to contribute 218–283 kg chlorophyll-a at any point in time. Their importance is not only the standing biomass, but also the presence of a food source for juvenile fauna in a protected environment (Fry, 1984). This makes the epiphytes ecologically significant even during a drought situation (Gordon et al., 2008).

Epiphytic algae contribute significantly to biogeochemical processes and nutrient cycling within submerged macrophyte beds (Walker et al., 1988). They specifically contribute to the production of calcium carbonate and towards nitrogen fixation (Goering and Parker, 1972).

Disturbances and threats

Studies on the impacts of human activities on epiphytes are very rare. Human activities that would cause the most significant effect on epiphyte biomass and diversity would probably be activities that cause sediment resuspension and nutrient enrichment. Sediment resuspension would cause turbid conditions, which would reduce light availability for these macroalagae and thus hinder the productivity. High nutrient availability would result in the decrease in diversity because of competitive advantage of other species over others and their prolific growth. Natural processes that would impact on the distribution of epiphytic microalgae have been highlighted below.

Grazing

The structure of the epiphytic community and biomass can be largely affected by grazing, especially if this grazing occurs at uncontrollable intensities. This was shown in a study by Schanz et al. (2002), where the epiphyte biomass was higher in the absence of grazers and different epiphytic community composition was found on the grazed versus ungrazed submerged macrophytes. Where there were grazers only diatoms were found, and some organic and inorganic debris, while the macrophyte beds without grazers had filamentous red and green algae, diatoms and cyanobacteria. Most of the grazers that exist within submerged macrophyte beds prefer the epiphytes over the plants mainly due to the high nutritional value of the epiphytes (Tomas et al., 2005).

Desiccation

According to Gottlieb et al. (2005), desiccation affects the relative abundance of the epiphytic species, and thereby the productivity of the community. When the water level drops the submerged macrophyte host and their associated epiphytes often become exposed. Whether the host survives periods of desiccation depends largely on its ability to recover from exposure and desiccation. As the duration of exposure increases, the desiccation stress on the epiphytic community will increase and the community structure will change. Based on the tolerance of species to desiccation, communities can be divided into long or short hydroperiod groups, i.e. long or short inundation periods. Long hydroperiod communities are more affected by desiccation than are short-hydroperiod communities because the latter are better adapted to tolerate desiccation. Species composition of long hydroperiod communities is also more varied with diatoms dominating under normal conditions and cyanobacteria under desiccated conditions due to their quick physico-chemical response to desiccation and re-wetting (Gottlieb et al., 2005). This change in community structure leads to a change in the productivity of the community and the value of the community. The increase in cyanobacteria and decline in diatoms decreases the quality of food available to grazers (Lamberti, 1996) and also results in the input of nutrients from the community to the water column due to decomposition (Gottlieb et al., 2005).

Water movement

Variation in water movement, most notably flow rates, affects the recruitment of organisms throughout the estuary. At least, the motile organisms have the option of distributing themselves according to their specific tolerances, while the sessile organisms, like the epiphytes and their hosts, are not able to do that. Since epiphyte recruitment is a passive process, higher numbers of epiphytes cannot settle on hosts if the rate of water movement is high and thus does not allow. Also, besides settlement, higher flow rates can result in lower epiphytic biomass due to the removal of the epiphytes (Cambridge et al., 1986). High flow can decrease the diversity of the epiphytic community, leading to dominance of the smaller epiphytic species and thus changing community composition (Bergey et al., 1995).

2.2 The botanical importance of estuaries

Botanical importance can be used as an indicator of ecosystem status because plants are the primary producers and will influence the status or condition of all higher trophic levels (Colloty, 2000). According to Coetzee et al. (1997), the condition and extent of plant communities can influence the overall ecological condition of an estuary. Coetzee et al. (1996) developed the botanical importance rating (BIR) index. In this index four plant communities were considered. Species were not considered separately because, unlike in most terrestrial environments, estuarine plant communities consist of either one, or very few, dominant species (O‟Callaghan, 1994). The communities were supratidal salt marsh, intertidal salt marsh, reeds and sedges and submerged macrophytes (Coetzee et al., 1996). The principle behind the scoring of the index was: the greater the area covered by a plant community, the fewer impacts associated with it, and the greater the number of communities (community richness), the higher the final score. The elements of the index were area covered, condition, importance, and richness of the plant communities.

Colloty et al. (1998) revised Coetzee‟s (1997) BIR index, firstly, to incorporate more communities. The swamp forest, mangroves, macroalgae and microalgae were included. They also introduced functional importance, which means that an estuary will have a high functional importance score if it contains a large or diverse number of habitats. Subsequently, Colloty et al. (1998) and Colloty (2000) decided that an assessment of botanical importance of an estuary should include functional importance, species richness, plant community richness and habitat rarity. Finally the functional importance, as part of the final BIR index score, was calculated as follows:

Functional Importance = a(Areasup) + b(Areaint) + c(Areasub) + d(Arearee) + e(Areaman) +

f(Areaben) + g(Areaphy) + h(Areamacro) + i(Areaswamp) where:

Areasup; Areaint; Areasub; Arearee; Areaman ; Areaben; Areaphy; Areamacro; and Areaswamp are the areas of supratidal salt marsh, intertidal salt marsh, submerged macrphytes, reeds and sedges, mangroves, benthic microalgae, phytoplankton, macroalgae and swamp forest respectively

a – i are the productivity values of each of the communities and they will not be discussed in this section.

The final botanical importance rating index would include the functional importance score, the species richness, plant community richness and habitat rarity scores.

2.3 The health of estuaries

Estuaries are highly productive coastal environments, providing habitat for a wide variety of fish and invertebrates (crabs, prawns and oysters) of commercial and recreational importance as well as food for many resident and migrant birds (WSAA, 2009; Davidson et al., 1991). They are also important for fisheries, wildlife and recreational leisure. Estuaries worldwide provide a focal point around which many coastal communities develop and grow (Svensson et al., 2007; Wolanski et al., 2004), placing estuarine environments under increasing pressure from human disturbances.

The concept of „health‟ has been used to signify whether environments are natural (healthy) or highly polluted and disturbed (unhealthy). Thus when it is said that an estuary is “unhealthy” it does not imply that it is full of pathogens and is dangerous for people to swim in; it has probably been completely transformed from its natural state. Pantus and Dennison (2005) see the measurement of ecosystem health (estuarine health) as expressing the degree to which the actual state of an ecosystem diverges from the ideal state as defined in the management objectives.

Based on an Estuarine Health Programme that was conducted in Australia (WSAA, 2009) it was evident that estuarine health assessments are important for providing regional information on the state of the environment, assess whether water quality is being maintained or improved by local and regional environmental programs, and to quantify environmental changes caused by pollution and habitat modifications. “Primarily an estuarine health monitoring program would be of benefit to national, state and local authorities responsible for environmental protection and nature conservation. However, it will have widespread secondary benefit through improved management of estuaries, which are an important commercial and recreational resource for many” (WSAA, 2009).

Therefore, it all revolves around a monitoring plan. A systematic and well-planned monitoring programme can identify water quality and other problems and help answer the

46 questions critical to solutions. Useful monitoring data can reveal the current physical, chemical, and biological status of an estuary (Ohrel and Register, 2006).

Estuarine health can be monitored in a number of ways, depending on the activities that might impact a certain estuary and the extent of available data for the system. For example, to monitor change in the trophic state of an estuary using change in macrophyte distribution there must be prior data that explain how the system‟s macrophyte distribution was before the estuary‟s trophic level changed.

2.2.1 The South African Estuarine Health Index

In South Africa Resource Directed Measures were designed to protect aquatic systems as part of Chapter Three of the National Water Act (Act No. 36 of 1998). According to the Act, there should be water of good quality and quantity set aside for basic human needs and for the ecology of that concerned water resource. These are respectively called the Human Reserve and the Ecological Reserve. In the process of determining the Ecological Reserve, the Resource Directed Measures procedure is followed (DWAF, 2008a). The Ecological Reserve Determination and Implementation method acknowledges that, in order for management to be effective, it is necessary to identify the present ecological state (health) of the systems, related to a reference or pristine state. The Estuarine Health Index (EHI) is used for that purpose and is the most crucial component of the process as it determines the health of the estuary and forms the basis for further management.

In recent years in South Africa, monitoring of the health of estuaries has been based on the improved Ecological Health Index, which is outlined by Turpie (2004) in the Resource Directed Measures manual. This makes it easier to take the information and apply it in the ecological reserve studies. This eventually leads to finding solutions for health problems experienced by estuaries and maintaining estuaries in the most attainable health status.

Assessing estuarine health in South Africa is relatively new. One of the earliest indices to have been formulated was the Community Degradation Index (CDI) (Ramm, 1988), which was based on comparing “the present fish community to the community that would exist in the absence of (or prior to) degradation”. Later Cooper et al. (1994) compiled an estuarine health index, which was a combination of a biological health index (Cooper et al., 1993), water quality index (House, 1989) and an aesthetic health index. This biological health index was based on Ramm‟s CDI in which existing fish communities were surveyed and compared to historical data on the same estuary. For the water quality index, selected chemical and bacteriological parameters were used. These were: dissolved oxygen, oxygen absorbed, ammonia, E. coli counts, nitrate, phosphate, and chlorophyll-a concentration. The aesthetic health index aspect considered the appearance of an estuary compared to its pristine state. The parameters that were taken into account were: floodplain landuse, degree to which channel margins were natural, persistent odours, presence of exotic vegetation, oil sheen,

47 mouth stabilization, presence of bridges and degree of visual impact from industrial or residential buildings. A botanical importance rating index (Coetzee et al., 1996) was also developed. Although originally not intended as a health index, it has, however, been used in this manner (Turpie, 1999).

After a workshop with other scientists and environmental managers Turpie (1999) formulated an approach for assessing estuarine health based on the above-mentioned indices, and others that were not included above (CERM‟s physical health index and the estuarine integrity index). Potential variables were identified and grouped into abiotic and biotic variables. The abiotic variables are: hydrology (mean annual runoff and river inflow patterns); hydrodynamics and mouth condition; water quality (salinity, nitrate and phosphate concentrations, suspended solids, organic and inorganic toxic water quality constituents, dissolved oxygen, pH, temperature and faecal coliforms); physical habitat alteration (sediment structure and distribution, estuary bed and channel modification, infilling, and human-induced impacts like disturbance of habitas and biota and constructions). Together these abiotic variables make up a habitat health score where each is rated according to its deviation from natural conditions. From the biotic variables, a biological health score was produced by looking at the deviation from natural of the biomass, diversity and community composition of the primary producers (microalgae and macrophytes), invertebrates, fish and birds.

Each variable is defined as a percentage of the pristine state and is weighted and then aggregated. The final score should reflect the state as a percentage of pristine (Turpie, 2004). The EHI score would then be a combination of the habitat and biological health scores, as outlined in Table 2.2 below. This score is used to determine the present ecological status of the estuary (present condition) that is divided into 6 categories (See Table 2.3). These categories range from A (unmodified, natural) to F (extremely degraded). This easily outlines the health of the estuary to managers and what the future desired state can be, after certain considerations.

Table 2.2: Calculation of the Estuarine Health Score

Variable Score e.g. Weight Abiotic (habitat) variables 1 Hydrology 41 25 2 Hydrodynamics and mouth condition 80 25 3 Water quality 59 25 4 Physical habitat 80 25 1. Habitat health score = weighted mean 65 50 Biotic variables 1 Macrophytes 60 20 2 Microalgae 60 20 3 Invertebrates 70 20 4 Fish 60 20 5 Birds 90 20 2. Biological health score = weighted mean 70 50 ESTUARINE HEALTH SCORE = weighted mean of 1 and 2 67.5

Table 2.3: Ecological Management Categories

Estuarine Health Present Ecological Description Index Score Status Category 91 – 100 A Unmodified, natural

76 – 90 B Largely natural with few modifications

61 – 75 C Moderately modified

41 – 60 D Largely modified

21 – 40 E Highly degraded

0 – 20 F Extremely degraded

2.2.2 Monitoring

Monitoring is a process of regularly collecting quantitative and qualitative information on key elements and processes. It is used to track long-term changes in key physical, chemical, biological and socio-economic factors that govern the health of South African estuaries (Adams and McGwynne, 2004; Ohrel and Register, 2006). Monitoring should form a crucial part of the environmental management process. It should tackle questions (objectives) that usually arise during the strategic assessment phase, and deliver the answers back to the management cycle during the evaluation phase (Adams and McGwynne, 2004). Therefore, a management decision will be as good as the information it is based on and that information is obtained through monitoring (Adams and McGwynne, 2004).

Monitoring can be done to assess ecosystem conditions based on specific goals, or it can be pursued for the compilation of continuous long-term records of environmental change. The latter can be very valuable for recording and differentiating between natural and human induced change, and for generating testable hypotheses on observed patterns and relationships (Wolfe et al. 1987) as long-term records are often rare. However, an objective framework should also guide this type of monitoring in order to avoid the feedback of little information to guide management decisions (Adams and McGwynne, 2004). When monitoring is driven by management objectives, it can be very purposeful. While it is measuring the compliance of the environment to specific conditions associated with a future desired state, the effectiveness of management strategies to achieve this state can also be determined (Adams and McGwynne, 2004).

Overall, and most importantly, monitoring programmes serve to bring problems to light, educate the public, provide a scientific basis for specific management decisions and strategies, contribute to the broad base of scientific information on estuary functions and the effects of estuary pollution, document the impacts of pollution control measures, and provide data needed to determine permit compliance (Adams and McGwynne, 2004; Ohrel and Register, 2006).

Biomonitoring vs. Physical and chemical monitoring

Management of an estuary to maintain a healthy environment requires maintaining the physical and chemical environments in the range normally experienced by the plants and animals in the natural community (biological component). This sets the stage for selecting the environmental parameters that will indicate estuarine health. Therefore, a monitoring programme can either look directly at the plants and animals (biomonitoring) to ensure a normal community is being maintained, or at the physical and chemical variables to ensure they stay in the range required by the community (WSAA, 2009). In an ideal situation the quality of water should be assessed by the use of physical, chemical and biological parameters in order to provide a complete spectrum of information for appropriate water management. However, such a study needs much more time and expenses than the study of the biological parameters, which, as it is widely accepted, can give reliably all the information about the aquatic system of concern (Iliopoulou-Georgudaki, 2003).

Chemical monitoring alone may not detect a short period between two sets of samples where environmental conditions changed to such an extent that all the animals in the estuary were killed. Another problem with chemical monitoring is that it is difficult to detect small changes because the variables naturally fluctuate over time. As a result, physical and chemical monitoring is best at detecting whether an estuary has experienced major pollution, but poor when it comes to detecting if the environment has been slightly modified and there has been a change in the plants and animals present. Changes to the plant and animal communities frequently occur below the detectable levels of a physical and chemical monitoring program (WSAA, 2009). The advantage of monitoring with the use of bioindicators is that biological communities reflect overall ecological quality and integrate the effects of different stressors providing a broad measure of their impact and an ecological measurement of fluctuating environmental conditions (Iliopoupou-Georgudaki, 2003). Overall routine monitoring of biological communities is reliable and relatively inexpensive compared to the cost of assessing toxicant pollutants (Iliopoupou-Georgudaki, 2003).

Although biomonitoring is labour intensive; it has advantages over chemical monitoring. Organisms integrate environmental conditions over a long period of time and can reveal a great deal about an estuary‟s health. For example, the widespread distribution of submerged aquatic vegetation can suggest that turbidity, or excessive nutrients, may not be a problem in the area. The presence of other organisms, such as high levels of bacteria, however, could cause concern because they can indicate the presence of pathogens in the water – a potential human health risk (Ohrel and Register, 2006). Conditions of intermittent or mild pollution, which are hard to detect by chemical means, may also be readily detected by changes in the biota. Biomonitoring also measures the actual effects on biota, whereas physical and chemical methods must eventually be interpreted on a biological basis (WSAA, 2009). Monitoring procedures based on the biota measure the health of an aquatic system and the

50 ability of that system to support life instead of simply characterising the chemical and physical components of a particular system. “This is the central purpose of assessing the biological condition of aquatic communities” (Harding et al., 2005). Cairns (1981) recognised that the condition of individual species and of communities of indigenous biota was one of the best measures of the health of an ecosystem, which is why biomonitoring is ideal.

Biomonitoring can be carried out on individual species or different community types. It is recommended that for estuarine health, monitoring at a community level would be more appropriate. Communities that are frequently used in monitoring programmes are macrobenthos, plankton, fish, birds and diatoms. Financial constraints usually limit monitoring programmes to one or two of these assemblages (WSAA, 2009).

Use of indicators

Estuaries are complex systems with a large range of habitats, animal and plant species, and physical and chemical conditions. As a result, a large number of monitoring parameters are being used to evaluate estuaries. Several parameters describe the basic chemical, physical, and biological properties of an estuary (Ohrel and Register, 2006).

“Indicators are central to the process of monitoring and represent the quantifiable tools used to collect and focus information” (Adams and McGwynne, 2004). Ecological indicators are commonly used to supply synoptic information about the state of ecosystems. They are quantitative representations of the forces that drive a system, of responses to forcing functions, or of previous, current, or future states of a system (Salas et al., 2006). They should therefore, according to Adams and McGwynne (2004), describe the extent of major pressures, the condition of ecosystems resulting from the pressures and the response of society to changes in the condition of the ecosystem. When they are used effectively, indicators are expected to reveal conditions and trends that help in development planning and decision-making. Most often they reveal an ecosystem‟s structure and/or functioning (Adams and McGwynne, 2004), such as nutrient concentrations, water flows, macro-invertebrates and/or vertebrate diversity, plant diversity, plant productivity, erosion symptoms and, sometimes, ecological integrity at the system‟s level (Salas et al., 2006).

According to UNESCO (2003), Salas et al. (2006) and Adams and McGwynne (2004), for an indicator to be effective it should have certain characteristics. Some of these characteristics might be repetitive, but indicators must: (a) have an agreed, scientifically sound meaning; (b) provide a representative view of important environmental conditions, pressures and societal responses; (c) provide valuable information with a readily understandable meaning and provide an early warning of unacceptable change; (d) be meaningful to external audiences and provide a basis for local, regional and national comparisons; (e) help focus information to answer important questions; (f) assist decision-making by being efficient, cost-effective and

51 easy to understand; and (g) have relevance to policy and management needs. Additionally, each indicator should be established with upper and/or lower limits of acceptable change that have been set scientifically and, sometimes, in communication with stakeholders (Ward et al., 1998).

Salas et al. (2006) states: “the main feature of an ecological indicator is to combine numerous environmental factors in a single value, which might be useful in terms of management and for making ecological concepts compliant with the general public understanding”. This is where indices come in. Different types of indices are used for different reasons. Some which have been used internationally will be discussed.

Indices based on indicator species

Indices based on indicator species attempt to characterize environmental conditions by analyzing the dominance of species indicating some type of pollution in relation to species considered as indicative of an optimal environmental situation. It is not advisable to use these indices because often such indicator species may occur naturally in relatively high densities. “The point is that no reliable methodology exists to know at which level one of those indicator species can be well represented in an unaffected community, leading to a significant exercise of subjectivity” (Salas et al., 2006).

“River ecology has an established long tradition in applying macrobenthos as bio-indicators” (Borja et al., 2000) and the majority of studies utilize soft-bottom communities to construct the indices. This is because macrobenthic animals are relatively sedentary (and cannot avoid deteriorating water/sediment quality conditions), have relatively long life-spans (thus, indicate and integrate water/sediment quality conditions, with time), consist of different species that exhibit different tolerances to stress, and have an important role in cycling nutrients and materials between the underlying sediments and the overlying water column (Dauer, 1993). This formed the basis of a Marine Biotic Index (AMBI) constructed by Borja et al. (2000). The AMBI was designed to establish the ecological quality of European estuarine and coastal environments by analyzing the response of soft-bottom communities to natural and man-induced changes in water and sediment quality, integrating long-term environmental conditions. This is derived from the proportions of individual abundance in five ecological groups related to the degree of sensitivity/tolerance to an environmental stress gradient (Borja et al., 2000). These groups are: species very sensitive to organic enrichment and present under unpolluted conditions; species indifferent to enrichment, always present in low densities with non-significant variations with time; species tolerant to excess organic matter enrichment; Second-order opportunistic species (mainly small sized polychaetes – subsurface deposit-feeders, such as cirratulids); and First-order opportunistic species (these are deposit-feeders, which proliferate in reduced sediments). The distribution of these ecological groups, according to their sensitivity to pollution stress, provides an AMBI with eight levels, from 0 to 7.

Another example of a species indicator index is the Ecological Evaluation Index (EEI) by Orfanidis et al. (2001). EEI was introduced for the evaluation of ecological status of transitional and coastal waters in accordance with the European Water Framework Directive. This index is based on the classification of the marine benthic macrophytes in two ecological state groups (ESGs I and II) representing alternative ecological states (pristine and degraded, repectively). ESGI represents a pristine state with late-successional species (or a high ecological status class {ESC}) while a degraded state is represented by the opportunistic species of ESGII. The marine benthic macrophytes (seaweeds, seagrasses) were used as bioindicators of ecosystem shifts because, as photosynthetic sessile organisms, they respond directly to the abiotic and biotic aquatic environment, and thus represent sensitive bioindicators of its changes (Orfanidis et al., 2001 and 2003).

Indices based on ecological strategies

Different organisms tend to apply strategic methods in order to counteract environmental stress. For this reason some indices are based on the ecological strategies of organisms (Salas et al., 2006). This is the case of the Feeding Structure Index (FSI) described by Petrov and Shadrina (1996) and Word‟s Infaunal Index (Word, 1979), both of which are based on the different feeding strategies of the benthic organisms.

According to Harding et al. (2005), the use of diatom indices, as a tool to provide an integrated reflection of water quality for aquatic waters, has rapidly grown in the last decade throughout the world. This has even led to the replacement of invertebrate-based indices as the biomonitoring method of choice in certain situations, e.g. canalised waterways (Prygiel and Coste, 1993). In France a biological diatom index (BDI) was developed in order to manage water quality (Bate et al., 2004). The system relies on community relationships of the epilithic diatoms, which was assembled over a 17-year period. The data comprise 1048 taxa. Rare species were then excluded to reduce this large number. The chemical parameters that were considered were: water temperature; pH; EC; TSS; BOD; COD; dissolved oxygen;

NH4; NO2; NO3; PO4; and chloride. This index has not been applicable in South Africa because, mainly, there is not enough information to determine rare taxa (Bate et al., 2004).

Indices based on species biomass and abundance

Other approaches consider the variation of an organism‟s biomass as a measure of environmental disturbances (Salas et al., 2006). These approaches include methods such as species–abundance–biomass–curves (SAB) (Pearson and Rosenberg, 1978), consisting of a comparison between the curves resulting from ranking the species as a function of their representativeness in terms of abundance and biomass (Salas et al., 2006). Warwick‟s (1986) abundance–biomass curves (ABC) method also involves the comparison between the cumulative curves of species biomass and abundance. He proposed the ABC method as a

53 technique for detecting the impact of pollution on communities of marine macrobenthos. The method was later developed to apply to organisms other than the macrobenthic. When Penczak and Kruk (1999) applied the method in the Pilica River, a tributary of the Vistula, they found it useful for assessing the impact of stresses on fish populations. Generally, the method proved a useful tool for estimating disturbances in fish communities caused by point source sewage inputs and all impacts of a dam.

Indices that integrate environmental information

From a more holistic point of view, some authors proposed indices able to potentially integrate the whole environmental information (Salas et al., 2006). Vollenweider et al. (1998) developed a Trophic Index (TRIX) integrating chlorophyll-a, oxygen saturation, and total nitrogen and phosphorus concentrations, in order to characterize the trophic state of coastal waters. TRIX works like a multimetric index. It combines certain arrays of an ecosystem and acts as an overall indicator of the quality of an ecosystem, with respect to the reference situation of absence or negligible presence of environmental stresses (Giovanardi and Vollenweider, 2004). Moreover, it offers the advantage of utilising, as components, environmental variables directly measured and routinely collected.

The Estuarine Biotic Integrity Index (EBI) reflects the relationship between anthropogenic alterations in the ecosystem and the higher trophic levels status. The underlying ecological principle of the EBI is that higher trophic levels (in aquatic systems, usually fish) require a diversity of intact ecosystem functions and processes to survive, grow, and reproduce (Deegan et al. 1997). The metrics that compose the EBI reflect water quality, habitat structure, water flows, food webs, and biotic interactions. The scores are poor, good, or excellent in comparison to a reference fish assemblage (Karr and Dudley 1981; Karr 1991). Deegan et al. (1997) suggested that the EBI is a useful indicator of estuarine ecosystem status because it reflects the relationship between anthropogenic alterations in estuarine ecosystems and the status of higher trophic levels.

Ferreira (2000) proposed an index for integrated evaluation of estuarine quality, based on an aggregation of four different components. The index is called EQUATION, meaning Estuarine QUAlity and conditION. The components that it is based on are: vulnerability, measuring the physical capacity of the system to react to change; water quality, which examines trophic status and eutrophication aspects; sediment quality, which looks at the sediments and benthic fauna; and trophodynamics, which addresses the quality and value of the top levels of the trophic web. The data requirements are reduced by the application of models and heuristic grading, and the four components are combined into an overall grade between 1 (worse) and 5 (better). The index was tested on very different estuaries in the US and Europe. The test ecosystems were chosen to study a range of physiography, tidal regimes, organic loading and contamination by persistent pollutants. Scores ranged from Low (grade 2) for the Elbe estuary to Excellent (grade 5) for Tomales Bay in California. The

54 methodology provides a useful synthesis of the basic descriptors of estuarine quality: physical aspects, water quality, benthos and higher trophic levels (including socio-economic aspects of fisheries). However, this index is not designed for detailed management of a particular system, which needs a completely different approach, focusing on specific problems and potential solutions.

Ecological indicators thermodynamically oriented or based on network analysis

The majority of indicators are based on a particular species or ecosystem component, e.g., phytoplankton, zooplankton, macrophyes, macrozoobenthos, etc. Even though these indicators provide useful information on ecosystem status, they do not reflect the real complexity of ecosystems (Austoni et al., 2007). It is therefore necessary for the indicators to include structural, functional and system-level aspects. In the last two decades, several functions have been proposed as holistic ecological indicators, intending (a) to express emergent properties of ecosystems arising from self-organization processes in the run of their development, and (b) to act as orientors (goal functions) in models development, as already referred above (Salas et al., 2006). Such proposals resulted from a wider application of theoretical concepts, following the assumption that it is possible to develop a theoretical framework able to explain ecological observations, rules, and correlations on the basis of an accepted pattern of ecosystem theories (Jorgensen and Marques, 2001). This is the case of Exergy (Jorgensen, 1997), a concept derived from thermodynamics, and can be seen as energy with a built in measure of quality. Exergy is defined as the amount of work a system can perform when it is brought to thermodynamic equilibrium with its environment or reference state (Jorgensen, 1997). Austoni et al. (2007) found that Exergy has recently been used to assess ecosystem health in freshwater, transitional and coastal ecosystems. It has also been applied to assess changes due to eutrophication. The energy of an ecosystem cannot be measured, but may be computed for each system component by multiplying its concentration measured in terms of average standing biomass, with its genetic information content.

2.2.3 The South African monitoring protocols

The existing South African monitoring protocols are integrative of all aspects of aquatic ecosystems (Taljaard et al., 2003; Adams and McGwynne, 2004). They do not focus on single parameters as indicators; they try to incorporate most of the indicators so as to have a holistic view of the ecosystem changes in an estuary. In South Africa, so far, there are two monitoring programmes that have been developed specifically for South African estuaries.

The Eastern Cape Estuaries Management Programme Monitoring Protocol

As part of the Eastern Cape Estuaries Management Programme, Adams and McGwynne (2004) developed a monitoring protocol, applicable to South African estuaries, that meets the needs of the ecosystem, society and management institutions, realising that there is a

55 discrepancy between the requirements of environmental managers and the information generated by existing monitoring programmes. However, it must be accepted that each estuary will require its own specific monitoring protocols because of differing management objectives.

The core of the monitoring protocol is to track long-term change in key physical, chemical, biological and socio-economic factors that govern the health of South African estuaries (Adams and McGwynne 2004). The protocol presents a rational and practical procedure to measure progress towards sustainability, a concept that has been interpreted as being the outcome of maintained ecological, societal and institutional integrity. Based on the dominant issues that threaten each of these domains, the protocol uses an objectives hierarchy to outline a pathway for reaching a future desired state defined in terms of operational goals (Figure 2.1). To monitor the achievement of these goals, a suite of indicators was developed through a series of workshops with scientists and managers and tested in co-operation with stakeholders associated with the Swartkops and Tyolomnqa estuaries in the Eastern Cape. The indicators are presented in 10 categories: (1) hydrodynamic and sedimentary processes; (2) water quality; (3) biodiversity; (4) human population growth; (5) control of human activities; (6) planning and development; (7) law enforcement; (8) co-operative governance and co-management; (9) effective management; (10) and satisfaction of basic human needs.

Managers working together with, or as part of, co-management forums are the target user group and, to accommodate limits in their resources, the indicators are presented in three levels of increasing skill requirements, time and cost. The indicators can be customised to meet management objectives that are bound to differ between estuaries. Each indicator is embedded in an interpretive framework where defined end-points signify thresholds of potential concern. Monitoring is an integral part of management (demonstrated in the estuarine management framework in Figure 2.2) and Adams and McGwynne (2004) state that this management protocol is iterative and adaptive and will provide managers and stakeholders with opportunities for learning through experience.

Monitoring key elements over an extended time period (>5 years) can outline long-term trends and identify possible cause-effect relationships that can be investigated through goal- directed research. Long-term monitoring can build a scientifically and morally defensible case for the introduction of restorative action where necessary that will illustrate South Africa‟s commitment to its global environmental obligations and to its people now and in the future (Adams and McGwynne, 2004).

Figure 2.2: An estuary management protocol that incorporates the monitoring protocol proposed for South African estuaries (taken from Adams and McGywnne, 2004, modified from Rogers and Biggs, 1999).

Resource Monitoring Procedures for Estuaries for Resource Directed Measures

The CSIR developed Resource Monitoring procedures for estuaries for application in the Ecological Reserve determination and implementation process. The aim of the project was to define guidelines and procedures for designing resource monitoring programmes for estuaries as part of the Ecological Reserve Determination process for estuaries, including baseline studies and long-term monitoring programmes (Taljaard et al. 2003).

These resource monitoring procedures are outlined for both baseline study purposes and long- term monitoring because the programme was developed for the Reserve determination and implementation. The baseline studies are required as the starting point for reserve determinations; when data is still being collected to assess the current state of the estuary. The purpose of long-term monitoring programmes, in this instance, is to assess (or audit) whether the Ecological Specifications (defined as part of the Ecological Reserve determination process) are being complied with after implementation of the Reserve. In addition, these programmes can also be used to improve and refine the Ecological Reserve measures, in the longer-term through an iterative process.

Given that each estuary is somewhat unique in its characteristics, the project provides generic sampling procedures (including recommended spatial and temporal scales) for each abiotic and biotic component, to be applied when a component is selected for inclusion in either baseline studies or the long-term monitoring programme of a particular estuary. Below is the list of the components that are usually assessed as part of Reserve determination process and thus form the crucial components of monitoring (also shown in Figure 2.3).

The abiotic components are: Hydrology; Sediment dynamics; Hydrodynamics; and Water Quality. The biotic components are: Microalgae; Macrophytes; Invertebrates (including zooplankton, benthic invertebrates and macrocrustaceans); Fish (ichthyofauna); and Birds (avifauna).

When drafting the resource monitoring programme, for each of the above abiotic and biotic components, the following should be specified in the procedures: Sampling procedures; Recommended spatial scales, i.e. selection of sampling stations; and Recommended temporal scales, i.e. frequency of sampling specific for baseline studies or long-term monitoring.

WATER QUALITY Salinity

SS/Turbidity/Light RIVER INPUT Sediment Load Toxic substances

Water Quality Dissolved oxygen

Flows Inorganic nutrients

HYDRODYNAMICS ESTUARINE VEGETATION INVERTEBRATES FISH BIRDS Water circulation Micro algae Zooplankton Plankton feeders Plankton feeders Stratifiction Invertebrate feeders Macrophytes Benthic invertebrates Benthic feeders Currents Piscivores Piscivores Macrocrustaceans

SEDIMENT DYNAMICS MARINE INPUT Gain size distribution Influence of tides State of Mouth Sediment Load Bathymetry Water Quality

Figure 2.3: Conceptual framework of the anticipated abiotic and biotic processes and interactions relevant to estuaries (from Taljaard et al., 2003)

Although baseline studies and long-term monitoring programmes have different purposes, it is extremely important that long-term monitoring programmes follow on from similarly structured baseline studies. In essence, the monitoring activities selected for the long-term monitoring programme should be derived from the monitoring activities conducted as part of the baseline studies, but implemented on less intensive spatial and/or temporal scales. The baseline studies that are carried out for an Ecological Reserve determination study at Comprehensive level may be considered as the baseline data against which the long-term monitoring is carried out. A list of abiotic indictors that should always be included in long- term monitoring programmes to allow for proper identification of „cause and effect‟ links, in particular links to river inflow and water quality are:

River inflow (i.e. flow gauging) Continuous water level recording at the estuary mouth (recording the state of the mouth, a key driver for most biotic components) Water quality of river inflow Water quality and flow rate of effluent discharges into the estuary Salinity distribution patterns under different river flow ranges.

Aerial photographs, collected on an annual basis, are also considered as key components in the long-term monitoring of estuaries, as these provide useful information on both abiotic and biotic components (Taljaard et al. 2003).

The inclusion of other abiotic indicators such as sediment characterisation, bathymetric and topographical surveys, water quality (other than salinity) in the estuary and accumulation of toxic substances in sediments will depend on the sensitivity of a particular system to changes in such indicators as well as the level of interaction of such indicators with the selected biotic indicators. Another consideration for inclusion would be any particular abiotic indicator that is on a „trajectory of change‟.

For the selection and prioritisation of the biotic indicators, the criteria that should be considered in the selection for long-term monitoring programmes include:

Ecological Specifications (i.e. Resource Quality Objectives for Biophysical Environment), specified for a particular estuary as part of the Ecological Reserve Determination process; The biotic indicators should be particularly sensitive to potential impacts associated with changes in river inflow and water quality, such as state of the mouth, tidal variation, sedimentation/erosion, salinity distribution patterns and deterioration in water quality; Biotic components considered to be on a „trajectory of change‟ or that are particularly sensitive to abiotic components that are on a „trajectory of change‟, should also be considered for inclusion as indicators in long-term monitoring programmes. Biotic components that are of regional or national biodiversity importance are also suitable indicators, particularly when also sensitive to changes in river inflow and water quality (Turpie et al., 2002 indicates the biodiversity importance of most South African estuaries); Biotic indicators should also be representative of the important food chains present in a particular system. This will ensure that monitoring programmes provide resource managers with appropriate data to establish „cause and effect‟ links, a key requirement for effective management of estuaries; The selection of biotic indicators should also present a balance between indicators that provides „early warning‟ signals and those that reflect longer-term, more cumulative effects. For example, fish are often considered to be useful „early warning‟ indicators, while macrophyte distribution patterns are often better indicators of cumulative, longer-term changes in estuaries.

Chapter 3: The botanical importance and present status of the Bushmans Estuary

3.1 Introduction

Various natural and human activities can shape the community composition of estuaries. For example, the Bushmans Estuary has dense, continuous stands of seagrass even up to the upper reaches of the estuary, which usually only occur in the saline lower reaches of most permanently open estuaries. As a result, the botanical status of an estuary can be used as an indicator of ecosystem status as the plants are the primary producers and their condition will influence the status of higher trophic levels (Colloty et al., 1999).

The objective of this study was to evaluate the changes over time of the botanical communities of the Bushmans Estuary, taking into consideration the low freshwater inputs and the marine sedimentation identified in recent years. Changes in the botanical communities were assessed from past aerial photographs and the application of a formula to calculate the relative botanical importance score for the estuary under natural (undisturbed) and present conditions. For a more holistic examination of the current condition of the estuary, the biomass, distribution and species composition of the microalgae was assessed, together with the physico-chemical variables that drive estuarine ecology. The biomass of Zostera capensis was determined to investigate the effect of the physico-chemical characteristics of the estuary on the distribution of the submerged macrophytes. As a measure of change in the salt marsh community over time, the species composition from 1995 was compared with a survey completed in 2008. This included documenting the distribution, percentage cover and species composition of salt marsh along the length of the estuary.

3.2 Materials and Methods

3.2.1 Assessment of the present status of the estuary

The present condition of the estuary was assessed during four sampling trips; one after a major flood event in 2006 and three under high and low rainfall periods. This was in August 2006, May and October 2008, and February 2009. Microalgal and macrophyte biomass, composition and physico-chemical variables of the water column were investigated. Sampling sites were chosen (Figure 3.1) distributed evenly from the mouth to as far as the estuary could be navigated by boat; 32.5 km from the mouth. In August 2006, after a flood, Mandilakhe Mdodana (NMMU Botany Honours student) obtained samples for the analysis of phytoplankton, benthic and epiphytic microalgae as well as the physical drivers. This sampling schedule was repeated in May 2008 during a dry period, with help from Maryanne Croker, who collected the epiphytic microalgae. In October 2008 and February 2009, a high rainfall period, Zostera biomass and salt marsh distribution were also investigated as well as the microalgae. The water column variables and phytoplankton samples were measured at different water depths. The depths were the surface (0 m), at half a meter (0.5 m) and then every meter to the bottom at 1 m intervals.

Figure 3.1: The location of the sampling sites along the estuary for: microalgae and physico-chemical (environmental) variables (M&E1-7; in red); Zostera capensis (Z1-9; in black); and salt marsh (Lower, Middle and Upper Estuary; in grey).

3.2.1.1 Environmental factors

Physico-chemical variables

The environmental variables that were measured in the water column of the estuary from the mouth to the upper reaches were secchi depth, which was measured with a secchi disc, and using a CTD (conductivity, temperature and oxygen) Meter and/or YSI Multiprobe, the temperature, salinity and dissolved oxygen concentration were measured at each depth.

Sediment analysis

Intertidal and subtidal sediment samples were collected from the same sites that the benthic microalgal samples were collected in order to determine the approximate sediment composition (sand, silt and clay) of the sediment along the length of the estuary. The samples were taken in October 2008 and February 2009 and compared with results from August 2006.

The hydrometer method by Gee and Bauder (1986) was used where approximately 40 g of previously air-dried sediment was accurately weighed out in a pre-weighed beaker and allowed to equilibrate with the atmosphere overnight. The following day, 100 ml of a 50 g.l-1 solution of Sodium hexametaphosphate (Na3PO)6 and 250 ml distilled water were added to the sample. The beakers were placed on a mechanical shaker for 1 hour prior to the start of the experiment. The sediment mixture was then placed in a 1 litre measuring cylinder and the volume made up to 1 litre with distilled water. The cylinder was closed off at the mouth and shaken by hand for at least 1 minute. Two drops of amyl alcohol were added to remove the foam on top of the sample. A hydrometer was inserted after 30 seconds, 60 seconds, 3 minutes, 1.5 hours and 24 hours. A blank containing a similar solution was also prepared but without sediment. The temperature of the solutions was taken using a mercury thermometer. The readings were then used in the following equations to calculate the percentage size fractions in the sample.

Determine C, the concentration of sediment in suspension in g.l-1, using: C = R – RL Where: R is the uncorrected hydrometer reading (in g.l-1) and RL is the hydrometer reading of the blank solution.

The determination of P, the summation percentage for the given time interval, using: P = (C/Co) * 100 Where: Co is the oven dried weight of the sample.

Determine X, the mean particle diameter in suspension in μm at time t using: X = _θt -1/2

Where: θ is the sedimentation parameter (μm min ½) and is a function of the hydrometer settling depth, solution viscosity and particle and solution density.

1/2 θ = 18ηh` / [ g(ρs –ρl)] 3 Where: h` is the hydrometer settling depth (cm), ρs = soil particle density (g/cm ), ρl = solution density (g/m3), g = gravitational constant (cm/s2) and η = fluid viscosity in poise (g.cm-1.s-1). The relationship of the settling depth to the hydrometer dimensions were approximated by h` = -0.164R + 16.3 Where: R is the uncorrected hydrometer reading. The summation percentage is calculated as follows,

P2μm = m ln(2/X24) + P24

Where: X24 is the mean particle diameter in suspension at 24 hours, P24 is the summation percentage at 24 hours and m were determined using the following equation

m = P1.5 – P24 / ln (X1.5 – X24) Where: m is the slope of the summation percentage curve between X at 1.5 hours and X at 24

hours. X1.5 is the particle diameter in suspension at 1.5 hours and P1.5 is the summation percentage at 1.5 hours. The procedure was repeated for the 30 sec and 60 sec readings.

Nutrients

Filtered water samples were collected from the sampling sites along the length of the estuary. These were treated with mercuric chloride and then frozen until further analysis. The samples were analysed for total oxidised nitrogen (TOxN) using the reduced copper cadmium method as described by Bate and Heelas (1975), and then ammonium, soluble reactive phosphorus (SRP) and silicate were analysed using standard spectrophotometric methods according to Parsons et al. (1984).

3.2.1.2 Phytoplankton

Biomass

Phytoplankton biomass in the estuary was measured as chlorophyll-a. Water samples (500 or 250 ml, depending on the turbidity of the water at the sites) were collected at the same depths as the nutrients. The water was filtered through Whatman GF/C filter papers and the filter papers were wrapped with foil and stored in the dark in a cooler box until they could be frozen in the laboratory .The chlorophyll-a was extracted by placing the frozen filters into 10 ml of 95% ethanol (Merck 4111). After extraction for 24 hours, spectrophotometric

65 determinations of chlorophyll-a were performed according to Nusch (1980). Before the spectrophotometric readings were done the extracts were once again filtered through the Whatman GF/C filter papers and then the new extracts were analysed. Absorbance was measured at 665 nm before and after acidification of extracts with 0.1 N HCl. The following equation was used to calculate phytoplankton biomass:

-1 Chl-a biomass (µg.l ) = (Eb665 - Ea665) x 29.6 x (v/(Vxl)) Where:

Eb665 = absorbance at 665 nm before acidification

Ea665 = absorbance at 665 nm after acidification 29.6 = constant calculated from the maximum acid ratio (1.7) and the specific absorption coefficient of chlorophyll-a in ethanol (82 g.l-1.cm-1) v = volume of solvent used for the extraction (ml) V = volume of the sample filtered (l) l = path of spectrophotometer cuvette (cm)

Community composition

Samples were taken from the different depths to evaluate the structure of the phytoplankton community. They were preserved with glutaraldehyde and stored at 4oC until the counts were done. The Coulon and Alexander (1972) method was used to settle 60 ml of the sample overnight for phytoplankton in 26.5 mm diameter settling chambers. To stain the cells 6 drops of Rose Bengal were added. After settling, the Zeiss IM 35 inverted microscope was used to count and identify the microalgal groups at a maximum magnification of 630x (Snow et al. 2000), where either 200 frames were counted per sample or 200 cells was counted. The area of the frame was approximately 3.142 mm2.

The actual counts for the different phytoplankton groups were calculated using the following equation, taken from Snow (2000): Cells.ml-1 = ((лr2)/A) x C/V Where: A = area of each frame (mm2) C = number of cells in each frame V = volume of sample in settling chamber (ml)

3.2.1.3 Benthic microalgae

Biomass

Chlorophyll-a was used as a measure of benthic microalgal biomass. Four replicate intertidal benthic samples were collected at low tide from each site by scraping the surface sediment

(<2 mm depth) just above the estuarine water level. Samples from Site 7 (the last site, at 32.5 km from the mouth), in the upper reaches of the estuary, could not be taken because this area had no accessible intertidal and subtidal area as the banks were lined with reeds and the bottom was rocky. Four subtidal samples were collected from each site using a 20 mm internal diameter corer attached to an extension pole and the surface sediment was scraped from the core. Both intertidal and subtidal samples were stored in the dark, in a cooler box until they could be frozen. The samples were freeze-dried; approximately 0.2 g was added to 4 ml of 95% ethanol (Merck 4111) and then stored for 24 hours at 0 °C. Once the chlorophyll-a had been extracted the samples were filtered through Whatman GF/C filters and the extract was analysed on a spectrophotometer at 665 nm before and after the addition of two drops of 1N HCl. The benthic microalgal biomass was calculated using the following formula:

-1 Chl-a biomass (µg.g ) = (Eb665 - Ea665) x 29.6 x (v/(m x l)) Where:

Eb665 = absorbance at 665 nm before acidification

Ea665 = absorbance at 665 nm after acidification 29.6 = constant calculated from the maximum acid ratio (1.7) and the specific absorption coefficient of chlorphyll-a in ethanol (82 g.l-1.cm-1) v = volume of solvent used for the extraction (ml) m = mass of the sample (g) l = path of spectrophotometer cuvette (cm)

Diatom species composition

Collection of diatom samples

The benthic microalgal community was sampled based on the method described by Round (1981) and the details described in Bate et al. (2004). The intertidal and the subtidal communities were sampled. The mouth of the collecting vial was used to scrape the surface and allowed to fill with a mixture of surface sediment and water. The mixture was stored in a plastic container and then some of the settled material was placed in a Petri dish overnight. The following day 5 clean cover slips were placed on top of the wet sediment and left for about four hours. Thereafter they were carefully removed and placed in bottles and stored until they could be digested.

Digesting

The diatom coated cover slips were placed in heat-resistant beakers to digest the samples on the cover slips. In order to get rid of other materials that would interfere with the identification of the diatom species (i.e. detritus; silt), 5 ml of saturated Potassium Permanganate was added to the samples. Then, another 5 ml of concentrated Hydrochloric Acid was added to the samples. The digestion was done according to Taylor et al. (2007a),

67 where the beakers with the material were boiled on a hot plate for approximately one hour, or until the mixture became pale yellow. The cooled solutions from each beaker were transferred into centrifuge tubes, measured up to 10 ml and centrifuged for 10 min at 2500 rpm. This process was repeated five times; after each centrifuge round the top 8 ml of the solution was discarded and then the tubes filled to the 10 ml mark with water. After the last round the top 8 ml were discarded and the remaining samples were transferred to micro-tubes and stored until they could be used for slide preparation.

Slide preparation and reading

A drop of the digested sample, from the bottom of the tube that has not been disturbed or shaken, was placed on a pre-cleaned cover slip and left to air-dry over a couple of days. After the water samples had dried, three drops of diluted Naphrax were placed on the centre of a pre-cleaned and marked slide and the cover slip lowered onto the drop with the dried sample side. The slide was then heated on a hot plate until bubbles formed, in order to fix the mountant. After this the slides were then left to dry over a couple of days, after which, the diatom frustules were examined under a Zeiss Axioplan light microscope with Differential Interference Contrast (DIC) optics. Using a television camera (Imaging Source DFK41F02), images of the different species that were found in the prepared slides from each site were visualised using the Imaplan V 2.06 image analysis programme (IMATEC Elektronische Bildanalysesysteme GmbH ©2004) and captured for identification and counts. The number of cells per species that occurred in each sample from each site was counted and the dominant species identified (those with > 10% relative abundance).

3.2.1.4 Epiphytic microalgae

Biomass

Zostera capensis host material was collected from five sites along the estuary from the lower to the upper reaches. Three replicates of the material were collected at the edge and the centre of the Z. capensis bed and from each replicate five leaves were scraped for the measurement of epiphyte chlorphyll-a and organic content.

To determine epiphyte biomass, the scraped material was filtered through Whatman GF/C filters and the filters were placed in foil and frozen for analysis in the laboratory. In the laboratory, 10 ml of 95% ethanol was added to the filter papers to extract chlorphyll-a overnight at 4oC. Then the extract‟s chlorophyll-a was measured with a spectrophotometer at 665 nm before and after acidification with 1N HCl. The following equation, from Hilmer (1990), was then used to calculate epiphyte biomass:

-2 Chlorophyll-a (mg.m ) = (Eb665 – Ea665) x 29.6 x (v/A) x 1000 Where:

Eb665 = Absorbance at 665 nm before acidification

Ea665 = Absorbance at 665 nm after acidification 29.6 = Constant calculated from the maximum acid ratio (1.7) and the specific absorption coefficient of chlorphyll-a in ethanol (82 g.l-1.cm-1) v = Volume of the solvent used for the extraction (ml) A = Area of the sample (mm2) 1000 = Correction factor (µg.mm-2 to mg.m-2)

Organic content

The scraped material was filtered through pre-combusted Whatman GF/C filters for the measurement of organic content. The filters were combusted in an oven for two hours at a temperature of 550 oC before they were used for filtering. The filter papers with the material were then placed in foil and frozen until further analysis in the laboratory. They were then weighed in the laboratory before and after drying at 60 oC for 24 hours and then after ashing at 460 oC for three hours. This allowed for the determination of both the organic and the inorganic fraction. For both the organic content and chlorphyll-a measurements, the area of the scraped leaf blades was measured.

Diatom species composition

From the same sites that the material was taken for biomass and organic content, plant material was collected for the determination of epiphytic diatom species composition. Five Z. capensis leaf blades were scraped from each site and the scraped material was placed in a glass vial and fixed with gluteraldehyde. These samples were then stained with Rose Bengal in the laboratory and the microalgal community composition was assessed by taking a drop of the stained sample and viewing it under a Zeiss Axioplan light microscope with Differential Interference Contrast (DIC) optics at a magnification of 400x. The community was found to be almost 100% composed of diatom cells.

Thereafter, following the guidelines from Taylor et al. (2007a) of diatom cell and slide preparation, approximately 2 ml of the concentrated bottom of the sample vials was taken and digested for the analysis of diatom species composition. Then, 2 ml of saturated Potassium Permanganate was added to the samples and left overnight. another 2 ml of concentrated Hydrochloric Acid was added to the samples and the digestion continued with placing the material, in glass test tubes, in a water bath and then boiling them on a hot plate for approximately one hour, or until the mixture became pale yellow. The cooled solutions from each beaker were transferred into centrifuge tubes, filled up to 10 ml with distilled water and centrifuged for 10 min at 2500 rpm. The centrifuging process was repeated five times; after each centrifuge round the top 8 ml of the solution was discarded and then the tubes filled to

69 the 10 ml mark with distilled water. After the last round the top 8 ml was discarded and the remaining samples were transferred to micro-tubes and stored until they could be used for slide preparation. The preparation of the slides was the same as that used for the benthic diatoms described in section 3.2.1.3.

3.2.1.5 Salt marsh cover and distribution

Salt marsh sampling in October 2008 was based on a study that was conducted on the Bushmans Estuary in 1995 by Oelofsen as a third year Botany Department research project. Twenty salt marsh sites along the length of the estuary were chosen for investigation, and data from these sites was grouped into results for the lower, middle and upper estuary reaches (Figure 3.1). These were the same sites used by Oelofsen (1995). At these salt marsh sites a visual estimation of the percentage cover of the different salt marsh species was done in the field. Unknown species were recorded and were taken back to the laboratory for identification. The results were then compared to those from Oelofsen (1995) in order to note if there have been any major changes in percentage cover or community composition.

3.2.1.6 Zostera capensis biomass

Leaf length biomass determination

In October 2008 and February 2009 nine sites were chosen along the length of the estuary from the lower reaches until 15 km upstream from the mouth, near the Ghio Bridge for Zostera capensis biomass measurements. The sampled sites are shown in Figure 3.1. At each site, plants were harvested from three quadrats (35 x 35 cm). The harvested material was then placed in plastic bags and transported back to the laboratory for analysis.

The leaf length method was used to determine the biomass of the seagrass. From the contents of each quadrat, 30 of the longest leaves of the plant material were measured and the leaf length was used to calculate biomass based on a regression analysis comparison, between leaf length and dry weight biomass. The February 2009 results were used to calculate the relationship between leaf length and biomass. A linear relationship was found (Figure 3.2) and the resulting formula was used to determine the October 2008 biomass by inserting the measured leaf length into the following formula:

Y = 0.308x + 9.234 where: Y = total biomass (g. m-2) x = leaf length (mm)

Figure 3.2: Regression comparison of the leaf length and biomass of Zostera capensis in 2009.

Dry weight biomass method

Destructive sampling was undertaken in February 2009. At the same nine sites and using the same quadrat size (35 x 35 cm), Z. capensis above and below ground material within the quadrats was dug out using a spade. The spade was pushed into the sediment to 15 cm in order to sample the below ground biomass. The contents were placed in plastic bags. The sediment and organic material (including some small animals) were washed off as much as possible from the plant and then the remainder was weighed and dried at 60oC in the oven for four days or until no moisture was evident. The mass of the plant material from each quadrat was measured after drying and the difference from the wet mass was determined and used to calculate the above and below ground biomass of Z. capensis from each site in grams (mass) of material per quadrat area (m2).

3.2.1.7 Human activities

Human activities along the length of the estuary were noted in October 2008. These activities were described for different reaches of the estuary namely; the lower, middle and upper reaches. This was done to identify where disturbance was taking place as this would influence the present state and health of the estuary. The described activities were those used by Forbes (1998) in a study on the recreational activities in a number of estuaries which also included the Bushmans Estuary. The number of developments along the estuary, jetties and houses that could be impacting the intertidal area, were also counted. From this a percentage of the disturbed intertidal area was estimated to be used in the Estuarine Health Index section.

3.2.1.8 Statistical Analysis

Statistica (Version 8) statistical package software was used to perform statistical analyses o fthe data. After the normality tests of the data were performed, the Students t-test was used to compare significant differences between two data sets and the Tukey‟s test was used for more than two sets, of parametric data. For non-parametric data, the Mann-Whitney Rank Sum test was used for the comparison of two data sets, then Kruskal-Wallis Anova on Ranks test for more than two sets. Pearsons Product Moment Correlation was used to test the strength of association between variables.

The Canonical Correspondence Analysis (CCA, in CANOCO for Windows, Version 4.5, 2002) was used for the representation and analysis of the results of identified diatom species from the benthic and epiphytic communities. This ordination technique was used in order to identify patterns in species distribution that corresponded with patterns in the measured water quality variables. The CCA diagrams were plotted with CANODRAW (v. 4.5, 2002). In the graphs the environmental variables are represented by arrows and the length of the line represents the importance of each variable. The position of the sites and species relative to these variables indicates their correlation with each other and the closer the species or sites are to a variable, the stronger the correlation with that particular variable.

3.2.2 Change in the botanical importance of the estuary

3.2.2.1 Assessment of habitat areas from aerial photographs

Aerial photographs of the Bushmans Estuary, from the Department of Surveys and Mapping, for 1942, 1966, 1973 and 2004 were assessed as well as another set of colour aerial photos that were taken in the 1990s (exact year not specified). The area covered by the different plant communities was digitised, using Geographical Information Systems software (ArcGIS, version 9.2, 2006) and the area covered by the different plant community types was calculated for the different years. The area covered by development and sand banks (as marine sedimentation has been identified as one of the problems in the estuary) in the estuary was also digitised. Only the 1966, 1973 and 2004 aerial photos were digitised along the entire length of the estuary. The quality of the 1942 photos was poor and made it difficult to digitise the boundaries of the habitat areas, especially where the estuary narrowed in the middle to upper reaches. Thus, for 1942, only the lower reaches of the estuary was digitised as well as for the 1990s photos because these photos did not show the entire estuary. The lower reaches in these years were digitised and compared with the other years (i.e. 1966, 1973 and 2004). The lower reaches showed the greatest change in habitat and thus it was important to track these changes over time.

Due to the poor quality of the assessed aerial photographs, some of the communities were under-represented while others were over-represented. To cater for such errors, after a visual

72 assessment of the estuary during a sampling trip, an estimation of the areas that were erroneously lost and gained during the digitising of the images was considered and has been included in the final habitat areas. This was done in order to ensure that the final botanical importance results were properly represented.

3.2.2.2 The Botanical Importance Rating Index

The change in the botanical importance of the estuary was determined for the whole estuary, using 1966, 1973 and 2004 aerial photographs. The botanical importance scoring method of Colloty (2000) was used to estimate the change in botanical importance over time. However, quantification of the botanical importance of the estuary was only completed up to and including the functional importance step. The rest of the formula was not used as these are fixed values and this study wanted to highlight the changes over time.

Botanical importance = Functional importance + species richness + community richness + sum of habitat rarity scores of each plant community type where:

Functional importance = a(Areasup) + b(Areaint) + c(Areasub) + d(Arearee) +

e(Areaman) + f(Areaben) + g(Areaphy) + h(Areamacro) + i(Areaswamp)

With: Areasup; Areaint; Areasub; Arearee; Areaman ; Areaben; Areaphy; Areamacro; and

Areaswamp as the areas covered by supratidal salt marsh, intertidal salt marsh, submerged macrophytes, reeds and sedges, mangroves, benthic microalgae, phytoplankton, macroalgae and swamp forest respectively. No swamp forest or mangroves occurred in the Bushmans Estuary.

a – i are the productivity values of each of the community types (Table 3.1).

Table 3.1: The productivity values for the different plant community types from Colloty et al. (1999) Productivity Productivity value Plant community type abbreviation (g.m-2.yr-1) Supratidal salt marsh a 863 Intertidal salt marsh b 651 Submerged macrophytes c 1 316 Reeds and sedges d 1 400 Mangroves e 1 811 Intertidal benthic microalgae f 124 Phytoplankton g 163 Macroalgae h 196 Swamp forest i 1 780

Functional importance gives a good representation of how the botanical importance of an estuary changes over time and so it was the only parameter of the BIR index that was applied in the current study to track the changes in the Bushmans Estuary. Below is the method applied to approach the other steps in the calculation of the BIR index.

Species richness = total number of plant species in estuary, and the scores were calculated according to Table 3.2.

Table 3.2: The scoring system for species richness based on the number of species present No. of species Species richness score 1-4 40 5-11 60 12-16 80 17-20 90 >20 100

Community richness = total number of community types present in the estuary. The scores were calculated from Table 3.3.

Table 3.3: The community richness scores based on the number of communities present No. of plant community types Plant community richness score 1-3 20 4 40 5 60 6 80 7-9 100

Habitat rarity = Sum of the number of times each community type in the estuary is found in other estuaries in South Africa. The score is attained by assigning a weight, which would be, for example; if the estuary has a mangrove community and mangroves are only found in 20 estuaries in South Africa, then the weight for the mangrove community would be 1/20 (i.e. 0.05). Weights for the other community types would be calculated and then, from the sum of the known weights the habitat rarity score would be assigned as indicated in Table 3.4.

Table 3.4: The scoring system for plant community type rarity Plant community type Sum of weights rarity score 0-0.017 20 0.018-0.027 40 0.028-0.04 60 0.041-0.049 80 0.05-0.067 95 0.068-0.133 100

Since only the habitat areas were assessed, only the functional importance of the estuary was determined and the steps following that were not applied, that is, the species and community richness, and habitat rarity steps.

3.3 Results

3.3.1 Assessment of the present status of the estuary

3.3.1.1 Environmental variables

Freshwater inflow to the Bushmans Estuary was low throughout 2008 (Figure 3.3a). In May flow was 0.013 m3.s-1 and in October it was 0.010 m3.s-1. In 2006 the flow was also low for most of the year, which is why there was no significant difference between 2006 and 2008 (p = 0.187). However, there was a flood in August 2006 which increased the mean monthly flow to 5.061 m3.s-1. The estuary returned to its low flow conditions by December (0.016 m3.s-1). The historical flow profile (Figure 3.3b) shows that there used to be more frequent high flow events in the estuary before 1990, after that when the flow did peak it was never as high as before. Additionally, the clarity of the water column (measured as secchi depth, Figure 3.4) in October 2008 was significantly higher than in August 2006 (p = 0.007) and February 2009 (p = 0.013, Figure 5). In October 2008, the water column was clear up to 1.7 m in the lower reaches of the estuary whereas it was only up to 51 and 64 cm in 2006 and 2009 respectively. In 2008 the estuary was sampled three days after a spring high tide that was on the 28 October. There was no significant difference in the overall clarity of the water column between August 2006 and February 2009 (p = 0.433), especially in the middle reaches of the estuary.

The horizontal salinity gradient, shown in Figure 3.5, from the mouth to the head of the estuary in August 2006 was significantly different to that found in the estuary in May 2008 (p <0.001), October 2008 (p <0.001) and February 2009 (p <0.001). Under low flow conditions there was no horizontal salinity gradient and no significant difference in salinity (p >0.05) between May 2008, October 2008 and February 2009. During all these times the flow of freshwater was low compared to the high flow in 2006. Figure 3.5 also shows that the estuary lacked stratification, and was mostly well mixed from the mouth to the upper reaches in May 2008, October 2008 and February 2009. During these times, salinity was above 30 PSU for most of the estuary and only decreased at 28.1 km from the mouth in October 2008 and at 25.2 km in May 2008 to an average of 25 PSU, while in February 2009 the salinity was 32.3 PSU even at the head of the estuary. However after the flood salinity stratification was evident in 2006 until 14.3 km from the mouth, where the top and bottom salinity was approximately 2 PSU, indicating freshwater conditions at this site.

There were significantly different temperature conditions in the estuary in August 2006 compared with that in May 2008 and February 2009 (Figure 3.6; p < 0.05). Since the temperature patterns in August 2006 were similar to those found in October 2008 (mean temperature 21.3o C, Figure 3.6), there was no significant difference detected between these sampling trips (p = 0.999). There was also vertical temperature stratification in the water

76 column in 2006 which started at 2.1 km from the mouth whereas the stratification in October 2008 only started at 25.2 km, with a maximum difference of 2.5oC from top to bottom. There was no vertical temperature stratification in the estuary in May 2008 and February 2009, the surface and bottom temperatures were the same along the entire length of the estuary showing how uniform the water column was. The average temperatures were 19.0oC and 26.8oC in May 2008 and February 2009 respectively and were close to this mean for most of the estuary.

In May and October 2008 the concentration of dissolved oxygen in the water decreased from the mouth to the upper reaches whereas in 2006 it increased towards the upper reaches (Figure 3.7). In 2006 the dissolved oxygen concentration was 6.0 mg.l-1 at 14.3 km upstream from the mouth. In May, oxygen decreased from 7.5 to 3.5 mg.l-1, from the mouth to the head of the estuary, while in October 2008 it decreased from 7.3 to 4.8 mg.l-1. The concentration of oxygen was not statistically different in May and October 2008 (p = 0.546). The dissolved oxygen in the water column was significantly lower in August 2006 compared with May and October 2008 (p<0.05). In most instances the concentration of dissolved oxygen was higher in the surface than bottom waters.

The dominant type of sediment in the estuary from 2006 to 2009 was silt (Figure 3.8), consisting on average of 61, 63 and 37 % of the sediment in August 2006, October 2008 and February 2009. The silt content decreased in 2009 and the clay content increased significantly from 16 % in 2006 to 33 % in 2009 (p = 0.012; F = 5.168; df = 33). The sand content also increased in 2009 to 29 % compared with the previous years of 23 and 17 % respectively (p = 0.016; F = 4.764; df = 33).

Figure 3.4: Clarity of the water column indicated by Secchi depth in August 2006, October 2008 and February 2009.

Figure 3.5: Top and bottom water column salinity along the length of the estuary in August 2006, May, October 2008 and February 2009.

Figure 3.6: Top and bottom water column temperature along the estuary in August 2006, May, October 2008 and February 2009.

Figure 3.7: Top and bottom water column dissolved oxygen concentrations along the estuary in August 2006, May and October 2008. No data were available in 2009 as the meter was faulty.

Figure 3.8: The distribution and composition of sediment particle types along the intertidal and subtidal areas of the estuary in August 2006, October 2008 and February 2009.

Nutrients

The estuary had significantly higher ammonium concentrations in May 2008 (Figure 3.9a), than in February 2009. The lowest concentrations were recorded in August 2006 (p < 0.001; F = 54.942; df = 42). Ammonium concentrations were low in August 2006 and October 2008 and there was no significant difference between the two months (p = 0.135). In February 2009 the highest ammonium concentrations were found at the head of the estuary (6.3 µM), decreasing towards the mouth. However, in May 2008 ammonium was high at the head and in the mouth region, and lower in the middle reaches, indicating that there were ammonium sources from the sea as well.

Figure 3.9b indicates that the overall total oxidised nitrogen (TOxN) concentration of the estuary was significantly higher in 2006 after the flood than for the other sampling periods (p < 0.001; F = 174.668; df = 242). The highest concentrations were found towards the upper reaches, with 24.0 + 0.3 µM at 14.3 km from the mouth decreasing to 17.8 + 0.3 µM at the mouth. However, the sea had nitrate concentrations higher than the estuary (26.2 µM). The lowest concentrations were measured in October 2008 and February 2009 (1.9 + 0.0 and 1.9 + 0.2 µM respectively), and these levels were not significantly different. The second highest concentrations were recorded in May 2008, but the concentrations of the TOxN had a variable distribution along the estuary with the highest concentrations detected 2.1 km from the mouth (15.2 + 2.7 µM) and at the head of the estuary, 16.1 µM.

The soluble reactive phosphorus (SRP) (Figure 3.9c) was highest in the estuary in 2006 compared to May 2008, October 2008 and February 2009; with averages of 55.3 + 2.7, 2.3 + 0.3, 1.3 + 0.1 and 0.2 + 0.0 µM respectively (p <0.001; F = 92.725; df = 42). There was no significant change in SRP from May 2008 to February 2009 (p>0.05). When there were detectable concentrations of the nutrient in 2006, the concentration of the nutrient in the sea (61.7 µM) was higher than the sites in the lower reaches of the estuary. The highest concentration was recorded at 14.3 km from the mouth (84.4 + 1.9 µM) decreasing down the estuary towards the mouth to 35.9 + 0.3 µM.

There was no significant change in the silicate concentration of the estuary amongst the sampling periods (p = 0.661; F = 0.419; df = 33), shown in Figure 3.9d. There was, however, a horizontal gradient that was evident along the channel of the estuary, from 2.7, 5.4 and 5.0 µM at the mouth to 102.2, 114.5 and 90.3 µM at the head in May 2008, October 2008 and February 2009. Overall, the silicate concentrations in the estuary increased towards the head of the estuary.

3.3.1.2 Phytoplankton

Biomass

The biomass of the phytoplankton community, represented by chlorophyll-a (Figure 3.10), was significantly higher after the flood in 2006 than the other sampled periods; decreasing from an average of 9.0 + 0.9 µg.l-1 in August 2006 to 2.1 + 1.0, 3.4 + 0.6 and 3.9 + 0.4 µg.l-1 in May 2008, October 2008 and February 2009 respectively (p < 0.001; F = 15.365; df = 42). After the flood the phytoplankton biomass of the estuary in 2008 and 2009 was similarly low and no significant difference was detected amongst those periods (p>0.05). In May 2008, the biomass was so low that it was mostly below 2 µg.l-1 in the estuary, until 25.2 km from the mouth, where it increased to 4 µg.l-1. The phytoplankton biomass of the estuary was mostly situated at the surface and was lower in the bottom waters in August 2006 and in May 2008. However, in October 2008 and February 2009, the biomass was higher in the bottom waters. This could have been due to the disturbance of the sediment and the subsequent resuspension of benthic microalgae. The bottom water column samples were analysed for diatom species composition at the sites where the bottom water biomass was higher than the top (i.e. for the sites at 7.4 and 25.2 km from the mouth). This showed that approximately 83 % of the diatom species in those water column samples were benthic diatom species indicating that the high biomass in the bottom water samples was due to disturbance and resuspension of the benthic biomass.

The correlation analysis showed that the occurrence and distribution of the phytoplankton biomass did not correlate to any environmental variable in 2006 (Table 3.5). However, in May 2008 phytoplankton biomass was related to high concentrations of silicate, phosphate and ammonium in the water column, increased water temperatures and reduced salinity. The strongest correlation for biomass during this sampling trip was with silicate. The only significant driver of the biomass in October 2009 was the high phosphate while in February 2009 it was high ammonium and low salinity. The relevant R-values are indicated in Table 3.5.

Community composition

The occurrence and distribution of the phytoplankton community in 2006 was not the same as was found in 2008 and 2009 (Figure 3.11). However, the dominant groups were always diatoms and flagellates. The abundance of diatom cells decreased since 2006 (p = 0.018; F = 3.790; df = 42). Diatom abundance was significantly higher in August 2006 (average of 1787 cells.ml-1) than in May 2008, October 2008 and February 2009 (p = 0.020; p = 0.015; and p = 0.022 respectively). After 2006 there was no great change in diatom numbers for the 2008 and 2009 sampling periods (with 169, 411, 170 cells.ml-1 in May, October and February respectively). Flagellate cells were significantly higher in 2006 (p<0.001; F = 9.159; D.F =

42) compared with that found in May (p = 0.003), October (p<0.001) and February 2009 (p<0.001), decreasing from 4156 in 2006 to 2132, 1587 and 821 cells.ml-1 respectively.

In 2006, the diatom abundance was the same as the flagellates (p = 0.051), but the distribution was different (Figure 3.11a-d). The diatoms were dominant in the lower reaches while the flagellates were dominant in the upper reaches. In 2008 and 2009 the flagellates were dominant throughout the estuary. Additionally, in February 2009 (Figure 3.11d), at the head of the estuary a Euglena species (relative abundance: 29%) was more abundant than the diatoms (relative abundance: 17%). The measured salinity at this site was, on average, 34.5 PSU.

R – Value Aug 2006 May 2008 Oct 2008 Feb 2009 Diatoms -0.516 0.725 0.0585 0.515 Flagellates 0.245 0.684 0.241 0.140 Temperature 0.404 0.661 0.492 -0.038 Salinity -0.292 -0.868 -0.106 -0.820 Dissolved oxygen 0.492 -0.552 -0.459 ___ Ammonium -0.608 0.660 -0.199 0.611 TOxN -0.097 -0.232 -0.488 0.243 Phosphate 0.068 0.650 0.596 0.194 Silicate ___ 0.855 0.560 0.407

Figure 3.10: The top and bottom phytoplankton chlorophyll-a concentrations (average + standard error) along the estuary in August 2006, May and October 2008, and February 2009.

3.3.1.3 Benthic microalgae

Biomass

There was no significant difference in the biomass of the benthic microalgae (Figure 3.12) for the different sampling dates (p = 0.240; F = 1.476; df = 35). Benthic microalgal biomass of the estuary was the same in August 2006, May 2008, October 2008 and February 2009 despite other environmental differences. In August 2006 the biomass of the intertidal area at 12.4 km from the mouth was significantly higher than the other sites (p = 0.007). Benthic chlorophyll-a at this site was 39.3 + 21 µg.g-1 while the average of the intertidal biomass was 13.9 µg.g-1. In most instances the intertidal area had higher benthic microalgal biomass than the subtidal (Figure 3.12). However, the differences between the intertidal and subtidal biomass were not statistically different (p = 0.069; F = 2.099; df = 43). For all sampling trips the lowest biomass compared to the rest of the estuary, was measured at 0.6 km from the mouth (intertidal and subtidal). Correlation analysis was used to test for significant effects of the environmental variables on the biomass of the benthic microalgae. A positive correlation with ammonium was found in 2006 (p = 0.024; R = 0.773). In October 2008 and February 2009 positive correlations were found with silicate and total oxidised nitrate (p = 0.031; R = 0.679; and p = 0.006; R = 0.799 respectively).

Diatom species composition

In total, 43 diatom species were identified in the benthic community of the estuary; only those species with a relative abundance greater than 10% were identified. Out of this, only a few were dominant (relative abundance > 20%) and common (found on more than one sampling trip and/or more than one site along the estuary) (Table A.1, Appendix A). Table 3.6 shows the benthic diatom species that were found in the estuary on more than one sampling trip and the species that were dominant in the benthic environment along the estuary. There were only six diatom species that were found more than once in the estuary during the sampling period. These were Seminavis sp., Amphora acutiuscula, Mastogloia exigua, Navicula gregaria, Nitzschia frustulum, and Planothidium delicatulum (Table 3.6). All these species, except for Seminavis sp., were common between 2008 and 2009 and occurred at more than 20% dominance at the sites where they were found, compared to the other species at those sites.

There were also species that were dominant in the 2006 sampling trip but were not found in the other years (Table 3.6), which could possibly be responding to the conditions of the estuary at that time. In 2006 the most common diatom species was Tryblionella constricta. The distribution of this species was widespread as it was found in the lower, middle and upper estuary. Navicula species were dominant in the upper estuary while Hippodonta cf. gremainii was dominant in and restricted to the lower reaches of the estuary.

In 2008 the species that was distributed throughout the estuary and at high abundance was Amphora. acutiuscula while Planothidium delicatulum was dominant in the middle to upper reaches of the subtidal area of the estuary. A. acutiuscula was dominant in the middle and upper reaches of the estuary in 2009 while it was Nitzschia. frustulum that was dominant in the lower to middle estuary. Fragillaria elliptica was also found more than once in 2009 in the lower reaches but its relative abundance was not high (11 and 10% at Sites 2 and 3 respectively). Overall, what was noticeable from the results was that the Navicula species were mostly distributed in the subtidal areas of the estuary while the Nitzschia species were limited to the intertidal benthic habitat (see Table A.1 in the appendices).

The ordination analysis of the benthic diatoms against the environmental variables (Figure 3.13) indicated that the distribution of the benthic diatoms in 2006 was driven by different factors compared with that in 2008 and 2009. The 2006 diatom species either responded strongly to soluble reactive phosphorus (SRP) or total oxidised nitrogen (TOxN), forming Groups 1A and 1B (Figure 3.13). The most abundant species in 2006 were also found in this group, showing their preference for high concentrations of SRP (Tryblionella. constricta and Hippodonta. cf. gremainii in Group 1A) and TOxN (Navicula sp, in Group 1B). Ordination confirmed that the estuary in 2008 and 2009 had similar conditions, the benthic diatom species from these years formed two mixed groups; containing both 2008 and 2009 species (Figure 3.13). The first group (Group 2A, Figure 3.13) showed a response to high salinity, conductivity and silicate and consisted of the dominant but less frequent (found only in one site) Nitzschia flexa, Navicula tenneloides and Diploneis elliptica found in 2009, and Amphora subacutiuscula and Nitzschia coarctata in 2008. Group 2B mostly comprised of the highly abundant and frequent species (found more than once in the estuary and that were at > 20 % relative abundance at the sites compared to the other species) identified in the estuary in 2008 and 2009 (i.e. Amphora acutiuscula, Nitzschia. frustulum, Planothidium delicatum). This group‟s distribution corresponded with high ammonium, temperature and the clay sediment type.

Figure 3.12: The average intertidal and subtidal benthic chlorophyll-a along the estuary in August 2006, May, October 2008 and February 2009 (average + S.E).

Table 3.6: Benthic diatom species list showing the common species for the different sampling trips and the dominant species (species with > 20% relative abundance).

Species name 2006 2008 2009

Amphora acutiuscula  

Cocconeis placentula v euglypta 

Diploneis elliptica 

Hippodonta cf. gremainii 

Mastogloia exigua  

Navicula gregaria  

Navicula sp 

Navicula tenelloides 

Nitzschia closterium 

Nitzschia coarctata 

Nitzschia frustulum  

Planothidium delicatulum  

Planothidium delicatulum 

Seminavis sp.  

Tryblionella constricta 

SPECIES ENV. VARIABLES 2006 Sites 2008 Sites 2009 Sites

1.0 Coc_plac 064S

Group 2B Nav_sali Fall_sp. Nit_font 092S Group 1A Fra_sopo Nav_cryp Try_litt Ope_hors 083I 065I Navi_sp. Fra_elli 092I Nit_frus 065S Nav_ammo 086S Dip_smit Pla_deli 094I Nit_clos 082I 093I 063S Clay 094S Semi_sp. Dip_bomb Nav_greg 084I 064I SRP Temp 093S 083S Amp_acut 085S Ast_punc MPB_Chla Sand Amp_coff NH4 084S Nit_flex TOxN 082S Salinity 095I 063I Cond Coc_scut pH Stau_sp. Mas_exig D.O. 062I Nav_penn Si2+ Silt Group 1B Amp_suba 062S Try_cons 086I Nit_coar 061S 085I Hip_grem Group 2A Dip_elli 096S Nav_tene 096I 061I 095S Nav_ramo 081S 081I Amp_wise -1.0

-1.0 1.5

Figure 3.13: Ordination (Canonical Correspondence Analysis) of the benthic diatom species found in the Bushmans Estuary in August 2006, October 2008 and February 2009 (Example of abbreviated names: 061S = Subtidal Site 1 sampled in 2006). The full names for the epiphytic diatom species are in Appendix A. The following codes for the environmental variables stand for: MPB_Chla = benthic microalgal chlorophyll-a (biomass); Temp = temperature; NH4 = ammonium; Cond = electrical conductivity; Si2+ = silicate; D.O. = dissolved oxygen; TOxN = total oxidised nitrogen; SRP = soluble reactive phosphorus.

Table 3.7: Summary of the CCA for the benthic diatom species against environmental variables found in the estuary in 2006, 2008 and 2009.

Axes 1 2 3 4 Total inertia Eigenvalues 0.971 0.944 0.932 0.895 17.520 Species-environment correlations 0.994 0.987 0.979 0.972

Cumulative percentage variance: of species data 5.5 10.9 16.3 21.4 of species-environment relation 11.8 23.3 34.7 45.6

Sum of all eigenvalues 17.520 Sum of all canonical eigenvalues 8.216

3.3.1.4 Epiphytic microalgae

Biomass and organic content

The average biomass of the epiphytes, represented by chlorophyll-a (Figure 3.14), in both August 2006 and February 2009 was significantly higher than in May 2008 (p = <0.001; F = 10.670; df = 44). The average epiphytic biomass in the estuary was 88.0 + 17.7 mg.m-2 in 2006, 1.7 + 0.8 mg.m-2 in 2008 and 61.8 + 14.4 mg.m-2 in 2009.

It was evident that the overall epiphytic organic content was significantly higher in May 2008 (28.6 + 2.1 %) than in August 2006 (4.7 + 1 %) (p < 0.001; df = 8; t = -7.714) but there was no difference in organic content between May 2008 and February 2009 (p = 0.657) (Figure 3.15). Overall, the organic content increased significantly from 4.7 + 1 % in 2006 to 28.6 + 2.1 % in 2008 then to 30.4 + 4.7 % in 2009 as freshwater input decreased (p<0.001; F = 27.316; df = 44).

There were no differences between sites for both epiphytic biomass and organic content in August 2006 and in May 2008 (p >0.05 for both), indicating that the organic content and biomass was the same throughout the estuary in those years. However, in 2009, there was a decrease in biomass (p<0.001) and organic content (p<0.001) in the estuary from the mouth to the upper reaches, decreasing from 165.1 + 45.8 to 18.9 + 3.5 mg.m-2 and from 52.9 + 3.5 to 17.6 + 2.6 % respectively. Peaks for both biomass and organic content were recorded at 7.4 km from the mouth. The results of the correlation analysis illustrated that there was no statistical relationship between the epiphytic biomass and the environmental variables.

Epiphytic diatom species composition

A sum of 69 epiphytic diatom species were identified in the estuary, but very few were dominant and a great number of the species were identified in 2009 (Table A.2, Appendix A). Table 3.8 shows the common and the dominant species that were found in the estuary. The four most dominant epiphytic diatom species in the estuary in 2006 were Cocconeis placentula v euglophyta, Navicula species, Fragilaria investiens and Navicula ramosissima with relative abundances of 27.8, 15.7, 11.8 and 11.1%, respectively. Nitzschia frustulum had the highest relative abundance (43.5%) in the estuary in May 2008, Navicula ramosissima (27.6%), Cocconeis scutellum v scutellum (7%), and Nitzschia closterium (6.4%) followed. The dominant epiphytic diatoms in February 2009 were Synedra species, Cocconeis placentula, Fallacia sp. and Nitzschia frustulum with relative abundance of 16.2, 12.9, 12.2 and 7.5 % respectively.

The only species common for the three sampling trips was Nitzschia frustulum which was the dominant species in 2008 (relative abundance 43.5%). Seminavis sp. was also found in 2008 and 2009 but not in 2006. There were four common species between 2006 and 2008. These 93 were Cocconeis scutellum v scutellum, which was abundant in 2008; Fragillaria investiens that was dominant in 2006 and Navicula ramosissima and Nitzschia frustulum were dominant in 2008 (Table 3.8). There were also four species common between 2006 and 2009 (Amphora sp., Cocconeis placentula v. euglophyta, Fragillaria sp. and Licmophora sp.).

The ordination analysis in Figure 3.16 showed that there were three groups of epiphytic diatom species that responded to different environmental factors. Group 1 consisted of species that were only found in 2006 and which responded to high levels of TOxN and SRP. This group mostly consisted of the diatoms with low relative abundance (i.e. Licmophora sp., Cocconeis sp., Navicula paeninsulae, and Tryblionella constricta) found in the estuary in 2006. In Group 2 there were mostly epiphyte diatom species that were recorded in 2008, which also had low relative abundance except for Nitzschia frustulum which had the highest relative abundance in 2008 (43.5 %). This group was associated with high ammonium, silicate, salinity, and organic content. The distribution of the epiphytic diatoms in 2009, which formed Group 3 were related to high temperature and low dissolved oxygen.

Figure 3.14: The chlorophyll-a (biomass) of the epiphytic microalgae along the estuary in August 2006, May 2008 and February 2009 (Average + S.E. bars).

Figure 3.15: The organic content of the epiphytic microalgae along the estuary in August 2006, May 2008 and February 2009 (Average + S.E. bars).

Species Name 2006 2008 2009 Amphora sp.   Campylodiscus sp.  Cocconeis placentula v euglophyta   Cocconeis scutellum v scutellum   Cocconeis sp.  Coscinodiscus sp.  Fragilaria investiens   Fragillaria sp.   Licmophora gracilis v gracilis  Licmophora sp.   Navicula paeninsulae  Navicula ramosissima   Navicula sp.  Navicula sp. 1  Navicula sp. 2  Nitzschia closterium  Nitzschia frustulum    Nitzschia sp.  Pleurasigma sp.  Seminavis sp.   Tryblionella constricta  Berkeleya fennica  Cocconeis engelbrechtii  Fallacia sp.  Synedra sp  95

Species Environmental Variables

2006 Sites 2008 Sites 2009 Sites

S309 1.0 S409 Syne_sp. Gram_sp3 S209 Navi_sp Salinity Org_cont Fall_sp. Coc_enge Group 3 Group 2 S509 Si+2 Ber_fenn NH4 S108 Amp_jost Temp Nit_clos S109 Frag_sp. S208 Nit_frus S308 Coc_plac S506 Camp_sp. S408 S508 Navi_sp1 Coc_scut S206 Amph_sp. Lic_grac Secchi Fra_inve Licm_sp. D.O. S106 Cocc_sp. Nav_paen Epi_chla S306 Try_cons SRP Nav_ramo Navi_sp. -1.0 S406 Group 1 TOxN

-1.0 2.0 Figure 3.16: Ordination (Canonical Correspondence Analysis) of the epiphytic diatom species and water column environmental variables in August 2006, May 2008 and February 2009 (Example of abbreviated names: S108 = Site 1 sampled in 2008). The full names for the epiphytic diatom species are in Appendix A. The following codes for the environmental variables stand for: Epi_chla = epiphytic microalgal chlorophyll-a (biomass); Org_cont = epiphytic organic content; Secchi = secchi depth; Temp = temperature; NH4 = ammonium; Si2+ = silicate; D.O. = dissolved oxygen; TOxN = total oxidised nitrogen; SRP = soluble reactive phosphorus.

Table 3.9: Summary of the CCA for the epiphytic diatom species against environmental variables found in the estuary in 2006, 2008 and 2009.

Axes 1 2 3 4 Total inertia Eigenvalues 0.87 0.808 0.592 0.496 4.509 Species-environment correlations 0.993 0.977 0.982 0.911

Cumulative percentage variance: of species data 19.3 37.2 50.3 61.3 of species-environment relation 22 42.4 57.4 69.9

Sum of all eigenvalues 4.509 Sum of all canonical eigenvalues 3.957

Epiphytic and benthic diatom species comparison

Ordination was used to compare distribution relationships for the benthic and epiphytic diatom species. In 2006, there was a mixed group of intertidal and subtidal benthic species which was correlated with high salinity and conductivity (Group 1, Figure 3.17a) and occurred in the lower reaches of the estuary at Sites 1 and 2. This group consisted of the species that had high relative abundance in the benthic habitat in 2006, i.e. Hippodonta cf. gremainii and Tryblionella constricta. Group 2 contained Cocconeis placentula, which was the dominant epiphyte in 2006 (27.8 %) and at subtidal Site 4 in the benthic area (with 60 % relative abundance), and was highly correlated with epiphytic organic content and temperature. The other species in this group were not strongly associated with the environmental factors. Group 3 consisted of the benthic diatoms from the subtidal Sites 3 and 5. The three dominant species were Tryblionella cf. littoralis, Diploneis bombus and D. smithii.

Most of the species in 2008, epiphytic and benthic, showed a correspondence to TOxN, salinity, and ammonium (Group 1, Figure 3.17b). Group 2 was a small group of intertidal benthic species that was strongly correlated with high silicate concentration and temperature. This group consisted of Nitzschia coarctata and Amphora subactuiscula which were both only found at intertidal Site 5 at 26 % relative abundance.

In 2009, most of the intertidal benthic and epiphytic diatom species formed a large group (Group 1, Figure 3.17c) that was strongly influenced by epiphytic biomass and organic content, salinity, SRP and ammonium. Group 2 was a subtidal benthic outlier group that did not appear to correspond with any of the environmental variables measured.

From the comparisons, it was also noticeable that the epiphytic and intertidal benthic diatom species grouped together and responded to common environmental variables or alternatively the intertidal and subtidal benthic diatoms formed a group. Also, some of the subtidal species did not respond to any of the measured environmental variables, and thus they were sometimes outliers.

A total of 14 diatom species were common between the epiphytic and benthic habitats, and they were: Amphora acutiuscula; Amphora coffeaeformis; Amphora sp.; Cocconeis placentula v euglyphyta; Cocconeis scutellum; Fallacia sp.; Navicula ramosissima; Navicula sp.; Navicula tenelloides; Nitzschia closterium; Nitzschia frustulum; Planothidium delicatulum; Seminavis sp. and Tryblionella constricta.

3.3.1.5 Current salt marsh cover and distribution

The results in Figure 3.18 show that the salt marsh species that were present throughout the estuary in 1995 were Bassia diffusa, Sarcocornia spp. and Spartina maritima. In the lower reaches S. maritima was the dominant species while Sarcocornia species were dominant in the upper reaches. Maximum B. diffusa cover was 15% and less throughout the estuary. In the lower reaches of the estuary in October 2008 S. maritima was dominant and Sarcocornia species were dominant in the upper reaches. However, the area covered by S. maritima in 1995 was much greater than in 2008, especially in the lower reaches; with 67% in 1995 and 47% in 2008. In 2008, there were more species than those found in 1995 in the lower reaches; these species might have replaced the areas that were covered by S. maritima. There were more Triglochin and Limonium linifolium plants in 2008 and also a large area of Phragmites australis that took up 35% of the salt marsh area at 0.9 km from the mouth. Also, there was a presence of Disphyma crassifolium in the supratidal area in 2008 that was not recorded in 1995 which may indicate drying of the marsh as this is a more supratidal / terrestrial fringe species.

A salt marsh wetland area influenced by freshwater inflow occurred approximately 5 km from the mouth, in a sheltered habitat. This consisted of a dense community of freshwater reeds, sedges and rushes (see Figure 3.19). This wetland consisted of stands of P. australis, Typha capensis (a freshwater rush species), and Bolboschoenus maritimus (brackish/freshwater sedge). Besides this noticeable wetland of freshwater plants, there were also other isolated areas were small stands of P. australis or B. maritimus and sometimes mixed stands were noticed along the left (east) bank of the estuary‟s mid-lower and mid- middle reaches. The upper reaches of the estuary consisted of mostly unhealthy looking pure stands of P. australis and sometimes mixed stands of the reed with Bolboschoenus maritimus (shown in Figure 3.20). It was also noticed that from the mid-lower reaches to the top of the middle reaches there are agricultural fields above the left bank of the estuary and depressions in the land surface, which have formed gullies. Thereafter, all the way to the head of the estuary, surrounding land on either sides of the estuary were old agricultural lands which were not in use anymore and have now become degraded.

Figure 3.18: Dominant salt marsh species in the lower, middle and upper reaches of the Bushmans Estuary in 1995 and 2008. Full species names can be found in Appendix C.

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Figure 3.19: Areas in the lower reaches of the estuary that had stands of freshwater reeds and sedges.

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Figure 3.20: The reeds and sedges communities along the estuary in the upper reaches. a) The unhealthy pure stands of P. australis; b) Mixed stands of the reed and sedge, B. maritimus

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3.3.2.6 Zostera capensis biomass

In October 2008 there was a significant increase in the biomass of Zostera capensis from the lower to the upper reaches (p< 0.001; F = 215.552; df = 269); increasing from 133.7 + 9.2 to 306.1 + 13.0 g.m-2 (Figure 3.21). Most of the leaves from the lower reaches had long dead tips and got easily broken off during measurements compared to the leaves from the upper reaches that were healthy and strong, otherwise, the biomass could have been uniform throughout the estuary as biomass was based on a relationship with leaf length. Conversely, in February 2009 (Figure 3.21) the biomass of the submerged macrophyte decreased significantly from the lower to the upper estuary (p<0.001; F = 72.035; df = 239), with a biomass of 154.1 + 10.8 g.m-2 just below the R72 Bridge to 70.6 + 1.6 g.m-2 at the last site in the upper reaches. Total seagrass biomass was significantly higher in 2008 (183.4 g.m-2) compared to 2009 (114.9 g.m-2). Ruppia cirrhosa was noted at Site 7 in Quadrat 3 in 2009. In this quadrat R. cirrhosa biomass (93.1 g.m-2) was greater than that of Z. capensis (65.5 g.m-2). No R. cirrhosa plants were found upstream or downstream of this site.

Samples were also collected to check if there would be a difference between the leaf length method and dry weight method of measuring biomass (Figure 3.22). It was found that both methods illustrated a similar distribution of biomass (p = 0.116; t = 1.601; df = 46) along the estuary. Figure 3.22 showed that the dry weight method indicated the same decline in biomass from the lower to upper reaches that was illustrated in the leaf length method, ranging from 156.3 + 13.7 to 92.0 + 4.5 g.m-2 from the mouth to the upper estuary.

Correlations of Z. capensis with physical variables of the overlying water showed strong correlations between the biomass of the submerged macrophyte and the nutrients in October 2008 (with phosphate R = 0.988 and p < 0.001 and with nitrate R = 0.683 and p = 0.043) and a negative correlation with sandy sediment (R= -0.975; p < 0.001).

3.3.1.8 Human activities

Most of the human activities were concentrated in the lower reaches of the estuary, especially in the mouth area and the first 7 km of the estuary where development was concentrated (Table 3.10). No human activities were recorded in the upper reaches and very few in the middle reaches. More people participating in activities were recorded on Saturday 28th of February 2009 (70 people). This was a greater number than that recorded on Saturday 1 November in 2008 (with 36 people recorded). The 28th of February was in the summer season while it was still spring and chilly on the 1st of November. On a week day, Friday 27th of February, only 22 individuals were active on the estuary and 17 of them were digging for bait. The most popular activities in the estuary were bait collecting, fishing and boating. Swimming was quite popular as well in the mouth region, about 0.8 km from the mouth, on Saturday 28th February with 20 people participating in this activity.

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The human impacts that were recognized along the estuary in the lower reaches were approximately 22 private jetties, a slipway, a small boat marina, a floating restaurant on the channel taking up a portion of the salt marsh area and several houses built on the supratidal salt marsh after the R72 bridge. Below the bridge alongside the banks all the way to the mouth, there are houses built below the 5 m contour line (also see the digitised maps in Section 2.4.1). In the middle reaches, 46 privately owned jetties were counted and a lodge below the 5 m contour line was also noted. Next to the restaurant is a parking area encroaching on the intertidal zone that has also disturbed the salt marsh. The upper estuary had one jetty on a salt marsh area. About 20% of the intertidal and supratidal area of the estuary has been lost to these developments including a portion of housing development on either sides of the estuary in the lower reaches, in the first 2 km of the estuary. Figure 3.23 are photographs which show some of these disturbances in the estuary

Table 3.10: Anthropogenic activities recorded in the estuary during the sampling trips and the total number of the individuals that were participating in these activities, as well as the number of jetties on the intertidal area as an estimate of the extent of development.

Number of people in estuarine reaches 01-Nov 2008 27-Feb 2009 28-Feb 2009 Activities Lower Middle Upper Lower Middle Upper Lower Middle Upper Walking on salt marsh 7 1 Swimming 4 20 Bait digging 3 17 8 Fishing 4 6 2 21 3 Boating 6 1 1 13 4 Canoeing 5 1 1

Total number of people 29 7 0 22 0 0 62 8 0

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Figure 3.21: Z. capensis biomass along the estuary in October 2008 and February 2009 measured using the leaf length method.

Figure 3.22: Z. capensis biomass along the estuary in February 2009 comparing the leaf length method and the dry weight biomass method.

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a) b)

c) d)

e) f)

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3.3.2 Change in the botanical importance of the estuary

3.3.2.1 Assessment of habitat areas from aerial photographs and changes over time

The greatest changes were noticeable in the lower reaches of the estuary as indicated by the digitised boundaries of the habitat areas in the 1942, 1966, 1973, 1990s and 2004 aerial photographs. The area covered by bare sand in the estuary decreased from 99.3 ha in 1942 to 58.0 ha in 2004 while the area covered by salt marsh increased from 44.4 to 72.9 ha respectively (Table 3.11 and Figure 3.24). There was a presence of 0.1 ha of a reed and sedge community in the lower estuary in 2004, which was not present in the previous years. There were also some other reeds and sedges stands that occurred in other parts of the lower estuary even though these areas are highly saline (above 30 PSU). The area of the open water has gradually decreased since 1942, declining from 89.1 to 77.4 ha in 2004. This habitat is now occupied by seagrass and intertidal salt marsh. The area covered by the seagrass (Zostera capensis) has gradually increased over the years.

Overall, when the habitat boundaries along the whole estuary were outlined (see Figure 3.25 and Table 3.12 for area covers in 1966, 1973 and 2004), it was evident that the salt marsh area has increased in size in the estuary and most of the increase in its cover was in the lower estuary. The area covered by the submerged macrophytes was also highest in 2004 (44.6 ha) compared to the other years, 20.7 and 27.0 ha in 1966 and 1973 respectively. On the other hand, the reeds and sedge community cover of the whole estuary has decreased since 1966, from 24.6 ha in 1966 to 16.8 ha in 2004.

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Table 3.11: Change in the cover of the different habitats in the lower reaches of the Bushmans Estuary.

Habitat area cover in the lower reaches (ha)

Habitat type 1942 1966 1973 1990 2004

Main water channel 89.1 83.0 80.5 72.1 77.4

Submerged macrophytes 11.4 17.0 19.4 25.8 25.6

Sand and mud banks 99.3 88.8 85.2 65.0 58.8

Salt marsh 44.4 45.3 50.4 72.4 72.9

Reeds and sedges 0.8

Developed area 15.6 24.5 24.4 21.9 27.3

Terrestrial vegetation 132.1 130.8 129.1 96.1 127.3

Total digitised area 391.9 389.4 389 353.3 389.3

Table 3.12: Overall change in the cover of the different habitat types from 1966 to 2004 along the entire length of the estuary.

Habitat area cover along the whole estuary (ha)

Habitat type 1966 1973 2004

Main water channel 165.1 156.6 151.9

Submerged macrophytes 20.7 27.0 44.6

Sand and mud banks 93.2 87.2 59.2

Salt marsh 86.0 89.5 126.0

Reeds and sedges 24.6 35.5 16.8

Developed area 24.5 24.4 27.3

Terrestrial vegetation 233.8 224.7 227.1

Total digitised area 648.8 644.9 652.9

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Figure 3.24a: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (1942 and 1966). 109

Figure 3.24b: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (1973 and 1990s). 110

Figure 3.24c: Maps of the lower reaches of the estuary comparing the changes in this section of the estuary over time (2004).

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Figure 3.25a: Map illustrating the changes along the length of the estuary in 1966. 112

Figure 3.25b: Map illustrating the changes along the length of the estuary in 1973 113

Figure 3.25c: Map illustrating the changes along the length of the estuary in 2004. 114

3.3.2.2 The Botanical Importance Rating index

The calculated botanical importance score for the whole estuary showed that the estuary has become 32.3 % more important than it was in the earlier years (i.e. 1966) (Table 3.13). This score represents change in the estuary over the years, and thus the estuary has changed by 32.3 % from its near-natural state. The habitat areas that contributed mostly to this change, due to increase in cover were the submerged macrophytes and salt marsh. The functional importance of these communities increased from 27241.2 to 58693.6 g.m-2.yr-1 per ha for submerged macrophytes and from 127882.0 to 187362.0 g.m-2.yr-1 per ha for salt marsh.

Table 3.13: Botanical importance of the entire estuary.

1966 1973 2004

Area (ha)

Phytoplankton 165.1 156.6 151.9

Submerged macrophytes 20.7 27.0 44.6

Intertidal benthic microalgae 93.2 87.2 59.2

Salt marsh 86.0 89.5 126.0

Reeds and sedges 24.6 35.5 16.8

Area x productivity

Phytoplankton 26911.3 25525.8 24759.7

Submerged macrophytes 27241.2 35532.0 58693.6

Intertidal benthic microalgae 11556.8 10812.8 7340.8

Salt marsh 127882.0 133086.5 187362.0

Reeds and sedges 34440.0 49700.0 23520.0

BIR 228031.3 254657.1 301676.1

Normalised score 100.0 111.7 132.3

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3.4 Discussion

3.4.1 Assessment of the present status of the estuary

The average flow in the Bushmans Estuary has been low in the past two years (less than 0.02 m3.s-1) because the Eastern Cape was experiencing a drought. However, drought or not, the estuary has low freshwater inflow due to high water storage by 30 weirs and farm dams in its upper catchment. This study showed that the estuary had poor water transparency most of the time, except in October 2008, when the clarity of the estuary could have been influenced by the spring tide event that had occurred three days prior to the sampling session. This poor water clarity can be attributed to high silt and suspended solids in the water column as a result of deposition from disturbed land in the catchment. Along the banks of the estuary, above the 5 m contour line from the lower reaches to the head of the estuary, there are agricultural lands and degraded bare lands (which are old agricultural lands that are not in use anymore) (Figure 3.26). These could be the source of suspended solids because there is no proper vegetation holding the topsoil of these areas, and thus these lands can be easily eroded by water runoff and be blown away by wind. The sediment would be deposited in the estuary, causing a silty water column. The particle size analysis showed that the sediment composition of the Bushmans Estuary was predominantly silt, thus showing an accumulation of this type of sediment on the benthos due to the deposition from these lands.

Figure 3.26: Evidence of agricultural activity near the estuary and abandoned land that has become degraded.

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In the Kromme Estuary the dissolved oxygen concentration ranged between 6 and 8 mg.l-1 throughout the estuary (Snow 2007). The Kromme is another estuary that has been freshwater-deprived due to abstraction by major impoundments and is thus marine- dominated. The concentrations of oxygen in the Bushmans were approximately 6 mg.l-1 and below, reaching near hypoxic (dissolved oxygen < 2 mg.l-1) conditions. Prinsloo (unpublished data) found a high abundance of filamentous algae in the Bushmans Estuary compared to the other estuaries sampled in the Eastern Cape. High nutrients from surrounding developments and seepage from the wastewater treatment works as well as raw sewage spillages could be responsible for the dense macroalgal growth. Algal mats consisting of filamentous algae are harmful, as they rapidly induce anoxia (Raffaelli et al., 1998). The filamentous algae gradually detach towards the end of their lifecycles and sink to the bottom often below the photic layer, where algal degradation proceeds (Salovius and Bonsdorff, 2004).

In 2006 the estuary exhibited horizontal and vertical salinity gradients because of the increase in freshwater inflow entering the estuary due to the flood event (flow was measured at 5.1 m3.s-1 at the flow gauge station). Otherwise, most of the time the estuary functions as an arm of the sea, similar to the Kromme Estuary where salinity can exceed that of seawater under drought conditions. The Bushmans Estuary is usually well mixed with no vertical stratification (Bornman and Klages, 2004). Analyses after the flood in 2006, in 2008 and 2009, have shown this to be true. In a stratified estuary there would be horizontal and vertical gradients of salinity, moving from the saline marine influenced mouth to the riverine influenced brackish to freshwater upper reaches. Robertson (1984) showed that the Bushmans Estuary has marine characteristics in dry conditions, but can be reset by small floods. In Robertson‟s study the flood event that reset the estuary was equivalent to 9.4 m3.s-1 of flow. After a month salinity distribution reverted back to the homogenous state. Interestingly, a horizontal salinity gradient was still evident in Robertson‟s assessment when the flow decreased to approximately 1 m3.s-1 and the current study showed that a 5 m3.s-1 freshwater flow event in 2006 established a horizontal salinity gradient.

Temperature patterns were similar to that of salinity. This is largely because temperature is usually determined by the amount of freshwater and seawater influence in an estuary. Temperature is also influenced by season with the water being warmer in the summer months and cooler in winter. This was the case in the Bushmans Estuary, the highest temperature of 26.8 oC was found in the summer month of February 2009. The water was the coldest in May 2008, which was the end of autumn with winter approaching. There was a horizontal temperature gradient in the estuary when there was freshwater influence in August 2006. Otherwise, in the low freshwater inflow periods no gradients were recorded. In 2006 the lower reaches were colder than the upper reaches due to the colder marine water entering at the mouth and the warmer freshwater entering at the head.

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Freshwater inflow into an estuary at its tidal head or via tributaries is often the most prominent nutrient source for an estuary. However, there is the ever-increasing threat of increased water abstraction and urbanisation, inadequate sewerage systems and increasing industrial and agricultural developments contributing to material input into estuaries and thus modifying the system (Schaler and Baird, 2003). Comparisons of three permanently open estuaries (Kromme, Sundays, Swartkops) subjected to different land use and/or anthropogenic activities highlighted the importance of freshwater inflow for the contribution of nutrients into estuaries and for dilution purposes in polluted systems (Scharler and Baird, 2003).

The results showed that the Bushmans Estuary is predominantly oligotrophic but that increased freshwater inflow introduced nutrients to the estuary. This was shown by increases in soluble reactive phosphorus (SRP) and total oxidised nitrogen (TOxN) after the flood event in 2006. This compares with the freshwater-starved Kromme Estuary (with mean flow mostly 0.0.07 m3.s-1 in the absence of major floods). Scharler and Baird (2003) showed that the Kromme Estuary had the lowest concentrations of nitrate and phosphate most of the time (averaging 8.8 and 0.6 µM respectively), as it suffers from freshwater impoundment in the catchment by two dams with a joint capacity of 133% of the MAR. However, it was also found that the nutrient concentrations become elevated when freshwater reaches the estuary, during maybe a significant flood event. Thus, highlighting the importance of freshwater input. The Berg Estuary also showed variability in nutrient concentrations with freshwater input; with high concentrations recorded in the high rainfall winter season and reduced nutrient levels in the dry (or low flow) summer period (Snow, 2007).

The second highest concentrations of TOxN in the Bushmans Estuary were detected in May 2008, which was a low flow period. This was, however, a month after raw sewage was reported to be spilling into the estuary (The Herald, February 2009). During this sampling period, the TOxN concentrations were especially high downstream of the seepage point (at 2.1 km from the mouth). The sewage spillages were observed from a conservancy tank between Bushman‟s River Mouth and Riversbend, about 4 km from the mouth. Raw sewage was noticed overflowing from the tank into a gully leading to the Bushman‟s Estuary (The Herald, February/January 2009). It was reported that, this situation commonly occurs after the holiday periods when the sanitation system reaches its maximum carrying capacity due to a larger number of users, of which during this period (May, 2008 sampling period) the April Easter holidays would have been the cause. Scharler and Baird (2003) highlighted that point sources along the Swartkops Estuary, i.e. canals draining residential and industrial areas and the polluted Chatty River tributary that drains an informal settlement, contributed more nitrate, ammonia and nitrite than the Swartkops River. Fortunately, freshwater abstraction from the Swartkops River catchment is limited, leaving a reasonable mean flow of 1.52 m3.s-1 to dilute the impacts that could result. The highest phosphate concentration was measured in the Swartkops Estuary (0.8 to 6.8 µM from mouth to the upper reaches).

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Other sources of nutrient input would be the agricultural lands adjacent to the estuary since many gullies drain into the estuary (see Figure 3.27) which could contribute to agricultural return flow. Return flow usually has high levels of nutrients. The Sundays Estuary experiences phytoplankton blooms due to agricultural return flow contributing nutrient-rich freshwater into the estuary, via the river. Scharler and Baird (2003) recorded the highest nitrate concentrations in the Sundays Estuary, and also the highest phytoplankton biomass. At least the Bushmans Estuary has a wider intertidal area to absorb these return flows drained by the gullies along the estuary.

The presence of the brine disposal pipe in the mouth region of the estuary might lead to inputs of nutrients. According to Du Plessis et al. (2006) salt is not the only constituent that is concentrated in brine. Several components present in the brine at high concentrations are sodium, calcium, magnesium, copper, nitrates, phosphorus, total dissolved solids (TDS), and fluoride (Du Plessis et al., 2006). This would explain why there was high phosphorus and nitrate at the sea and not at 0.6 km from the mouth. If the sea naturally had high nutrients, then the nutrients would be high as well at this site because then the sea would be contributing its high nutrients to the estuary, instead the nutrients were high in the upper reaches, decreased towards the mouth and were high again in the sea.

Figure 3.27: Gully/valley formations towards the banks of the estuary.

3.4.1.1 Phytoplankton

South African studies have shown that phytoplankton biomass in estuaries was positively correlated with freshwater input (Alanson and Read, 1995; Hilmer, 1984; Hilmer, 1990; Hilmer and Bate, 1990) because the phytoplankton communities are supported by the freshwater inflow which supplies nutrients and maintains stable stratified conditions. This has been the case in the Bushmans as well, with the highest biomass (9.0 + 0.9 µg.l-1) recorded under high flow conditions in 2006 (the mean flow was 5 m3.s-1 in August 2006 due to the

119 flood) when the nutrient concentrations were highest. Phytoplankton biomass decreased to 2.1 + 1.0 µg.l-1, 3.4 + 0.6 µg.l-1 and 3.9 + 0.4 µg.l-1 in May 2008, October 2008 and February 2009 respectively when freshwater input became relatively low (with flow less than 0.1 m3.s-1 during those periods). The same trend has been recorded in the Kromme Estuary, which is also a permanently open estuary that receives little freshwater input due to the capacity of the dams being equivalent to the mean annual run-off of the catchment. Snow (2006) measured particularly low phytoplankton chlorophyll-a (3 µg.l-1) in the freshwater-deprived period when sampling was undertaken (November 2003 and July 2004); indicated by mean salinity concentrations that were similar to or higher than seawater. In an earlier study the average chlorophyll-a concentration increased to a maximum of 5.6 + 0.3 µg l-1 almost a month after a 2 x 106 m3 pulse of freshwater in 1998/1999 (Snow et al. 2000a). This increase in biomass was relatively small and was related to the rate that these nutrients were coming in; high flow rate does not give the microalgae enough time to absorb the nutrients. Hilmer (1990) noted that retention time plays a major role in phytoplankton activity and that a continuous freshwater flow of about 1 m3.s-1 was adequate to maintain phytoplankton biomass in the Sundays Estuary. In Snow (2000), the assessment of the Gamtoos Estuary showed that phytoplankton biomass in the estuary was more pronounced at flow rates of about 0.8 and 1.2 m3.s-1 (resulting in chlorophyll-a values of 47.5 and 49.9 µg.l-1 respectively). At the higher flows from 2.3 m3.s-1 to 30.5 m3.s-1 mean phytoplankton biomass was lower than 20 µg.l-1. Lowest biomass was measured at the low flow of 0.3 m3.s-1 in this estuary. Thus, also confirming that the lower the freshwater inflow the lower the productivity in that estuary, as was found in the Bushmans Estuary.

In 2006 and May 2008 the highest phytoplankton biomass was recorded at the surface compared with the bottom waters which could be attributed to light penetration. Adams and Bate (1999b) mention that light reduction from heavy silt load will act to reduce phytoplankton production. The Bushmans Estuary is a highly turbid estuary, it was declared as such by Doudenski (2004) who measured unacceptable TDS levels of 2200 mg.l-1 (compared with the 200 mg.l-1 that is usually measured in the headwaters) and it was also evident from the low secchi depth measured in the current study (less than 100 cm most of the time, while the estuary was mostly 2.5 to 3 m deep). This then acts to reduce light penetration to the bottom of the water column and leads to reduced phytoplankton productivity at depth. However, in October 2008 and February 2009 some of the sites had higher biomass at the bottom than the surface, which is not usually the expected response. To investigate this bottom water samples from these sites were analysed for diatoms species composition. It was found that approximately 83 % of the species in these sites consisted of those diatoms usually associated with the benthic environment (i.e. pennate species; Taylor et al., 2007b). Therefore, the high phytoplankton biomass in bottom waters was probably due to resuspension of the benthic species due to disturbance of the sediment while sampling.

Low phytoplankton biomass was associated with low phytoplankton cell numbers (diatoms and flagellates). Based on Margalef‟s (1978) findings, the flagellates in the Bushmans

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Estuary could be the dominant microalgae because of the low levels of nutrients. However, Margalef‟s study also found that the flagellates increase in abundance when stratification has set in and the water has become depleted of nutrients and the Bushmans lacks stratification most of the time. Margalef (1978) also states that diatoms are favoured in nutrient-rich, well- mixed waters during spring tides and most of the time the Bushmans is well mixed but is not nutrient-rich. Thus, in 2006 the diatoms and flagellates were co-dominant because the water had adequate nutrients (the source being freshwater coming in at high volumes due to the flood) for the diatoms to grow abundantly and at the same time, the estuary had vertical stratification, evident from the salinity gradient, which favoured the flagellates. After the 2006 flood, the diatoms probably decreased in numbers because of reduced nutrients and the flagellates decreased because the estuary was well-mixed and lacked favourable stratified conditions. However, overall, the flagellates were still more abundant than the diatoms because of their tolerance to low nutrients. Furthermore, Hlaili et al. (2006) recorded a healthy growth of flagellates under nitrogen deficient conditions while diatom growth and biomass was stimulated by nitrogen enrichment. The number of phytoplankton cells in the estuary decreased as the nutrients decreased, thus indicating that the Bushmans microalgae are autotrophic (they require nutrients to support their growth). If they were heterotrophic cells they would have been able to survive and maintain their growth by feeding on other organisms.

In a study by Adams and Bate (1994b), in all the estuaries that were studied (Berg, Palmiet, Goukou, Gourits, Great Brak, Keurbooms, Gamtoos and Sundays) the phytoplankton community was dominated by flagellates. In their study the levels of nitrate, which was found to be the favoured nutrient when it comes to phytoplankton biomass, varied greatly among these estuaries and so the nutrient levels could have not determined the species composition. However, all the estuaries displayed horizontal and vertical salinity gradients, which must be an indicator that the estuaries provided the stratified condition that flagellates relate to. So, from this study it was concluded that phytoplankton biomass relates better to water column nitrate concentrations and the salinity gradient determined the species distribution (Adams and Bate, 1994b).

3.4.1.2 Benthic microalgae

The average benthic biomass in the estuary was low but, when compared with other estuaries, the Bushmans Estuary has comparable benthic biomass. Sometimes the intertidal biomass of the estuary was even higher than some of the most nutrient-rich estuaries, like Sundays, Gamtoos and Swartkops (Table 4.14). Table 3.14 is an insert from Snow (2007) displaying his data on some of the South African estuaries he has worked on and measurements from the current study have also been included for comparative purposes. It thus turns out that, even with the low nutrient concentrations and the marine dominated environment, benthic microalgae can be productive indicating that benthic biomass may be independent of freshwater input and the characteristics of the overlying water column. Meyer and Meyer-

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Reil (1999) also noted that under conditions of low nutrient availability in the water column, benthic microalgae can survive by benefiting from the relatively high nutrient concentrations in the pore water of sediment. The water column nutrients usually have no direct relation to benthic microalgal biomass. Also, from a recent study done on the Keurbooms, Mngazana, Gamtoos, Swartkops, Sundays and Mngazi estuaries, Snow (2007) established that the microphytobenthic biomass had the strongest associations with sediment related variables than the water column variables. These were high sediment organic content and fine sediments such as silts and mud. This would then explain the fact that the biomass of the benthic microalgae in the Bushmans Estuary was comparable with the nutrient-rich estuaries (i.e. Gamtoos, Swartkops), even though its water column nutrient concentrations were very low (oligotrophic). This suggested that the benthic microalgal biomass was reliant on the sediment-trapped nutrients and other sediment characteristics. It should also be noted that while there was no significant correlation between benthic biomass and the distribution of the silty-type sediment in the estuary, there were statistically significant correlations between biomass and nutrients. This correlation was perfomed to test whether associations existed between the subtidal benthic biomass and environmental variables recorded in the deepest part of the water column and between the intertidal biomass and the surface water variables, as this would represent the interface between sediment and water column. The results showed positive correlations of benthic biomass with ammonium in 2006, silicate in October 2009 and TOxN in February 2009. This indicated that benthic microalgal biomass at each site was controlled by the immediate environment at that site. The conditions of the sediment- water interface and sediment characteristics influence benthic biomass. Further, investigations will however be required to confirm this.

Other factors that have been found to control benthic biomass are water currents, light flux, depth, physical disturbance, grazing, sedimentary composition and disturbances and water speeds (McIntyre et al., 1996; Adams and Bate, 1999b). Thus, the intertidal area of the Bushmans had more biomass than the subtidal most of the time because of the turbid nature of Bushmans water. Since the intertidal habitat is usually shallower, the microalgae are able to receive greater light than in the subtidal area where they are shaded by the silty water. Cahoon et al. (1999) found that loadings of fine sediment may reduce the biological productivity of shallow water ecosystems. On the other hand, Walker (2003) attributed higher intertidal biomass to the instability of the subtidal sediments because there would be more disturbances from tidal action in the subtidal habitat.

Sediment type influences microphytobenthic biomass (Walker, 2003) because it influences nutrient availability for microphytobenthic organisms. The coarser sandy sediment is easily stirred by water movement and thus cannot maintain nutrient resources in the porewater. Whereas mud can sustain twice as much benthic microalgal production because it usually has high organic matter content, with high rates of bacterial mineralization and high porewater concentrations of dissolved nutrients (Admiraal, 1984; Davies and McIntire, 1983; Adams and Bate, 1999b). Thus, the low biomass in the mouth area in the Bushmans Estuary would

122 be due to the type of the sediment in the area (sand dominated) and the strong water movement due to tidal currents, which results in oligotrophic sand flats (Admiraal, 1984). The higher benthic biomass was found towards the upper estuary where the sediment gets muddier and tidal flows are reduced. Walker (2003), also found lower biomass in the mouth region of the Kwelera Estuary and concluded that the more turbulent hydrological conditions and loose sediment in the mouth area did not allow benthic biomass to develop.

Table 3.14: Comparison of the measured benthic microalgal biomass in the Bushmans Estuary with that of other estuaries as measured by Snow (2007).

Estuary Sampling date Intertidal chl a (µg.g-1) Bushmans August 2006 13.9 + 7.3 01 May 2008 18.8 + 5.9 31 October 2008 18.1 + 4.7 29 February 2009 22.4 + 2.3 Kromme 24 November 2003 12.9 + 2.5 30 July 2004 4.9 + 0.4 Gamtoos 7 August 2002 27.7 + 3.0 21 February 2003 10.4 + 0.9 Keurbooms 25 August 2002 9.5 + 0.8 Mngazana 22 June 2003 27.2 + 2.3 25 January 2003 16.4 + 2.5 Mngazi 26 January 2003 12.9 + 0.9 Sundays 19 February 2003 9.1 + 1.0 25 July 2002 10.2 + 1.0 Swartkops 5 December 2002 3.7 + 0.6 12 February 2002 11.2 + 0.9 15 August 2003 22.5 + 1.9 28 November 2001 5.9 + 0.9 30 October 2001 8.2 + 0.8

Ordination results showed a separation of the 2006 benthic diatom species from the 2008 and 2009 species and their correlation with high TOxN and SRP concentrations, which were higher in the estuary in 2006 compared to the other years. It can thus be assumed that these nutrients were the important influencing factors in the estuary in 2006, which contributed to the species composition and distribution. The Bushmans Estuary is usually marine-water dominated and it is thus expected that the species would have responded to the incoming nutrients when the flood event occurred in August 2006. The presence of Tryblionella litorallis and Diploneis smithhii, which are brackish water species (Taylor et al., 2007b), in the 2006 group of species indicates that this group favoured the freshwater diluted nature of the estuary, with the higher nutrients available at that time.

The large 2008 and 2009 group of benthic diatom species were related to high ammonium, high temperature and the clay content of the sediment. The Navicula spp. found in the Bushmans Estuary and listed in Taylor et al. (2007b) are tolerant to critical levels of pollution, mostly eutrophic conditions. Watt (1998) also found out that the sediments of the

123 severely polluted Manzimtoti Estuary contained large numbers of small Navicula and Nitzschia spp. while the oligotrophic Mhlabatshane Estuary was dominated by Achnanthidium and Diploneis spp. In the current study Navicula gregaria was in this 2008 and 2009 group, which is listed by Taylor et al. (2007b) and Tornés et al. (2007) as a good indicator of eutrophic to hypertrophic freshwaters with moderate to high electrolyte content and polluted conditions. Navicula crpytotenelloides was also in this group, which is also a species that is usually found in meso- to eutrophic calcareous streams and lakes (Taylor et al., 2007b).

Nitzschia. frustulum was also part of the group and it has been listed as a species that is tolerant of poor water quality, according to Harding et al. (2004)‟s water quality class list of diatoms (as per the proposed European Guidance Standard). John (2002) also identified this species as abundant in highly nutrient enriched water and Taylor et al. (2007b) and Tornés et al. (2007) described it as being tolerant to, and an indicator of, critical levels of pollution. At Bushmans Estuary, this species was dominant in the lower to middle estuary sites in 2008 and 2009 indicating the possibility of pollution in this vicinity. These sites were in the vicinity of the site were raw sewage had been leaking into the estuary due to spillage from an unmaintained conservancy tank. Responses to these spillages were not noticed in the water column microalgal biomass and nutrient concentrations, however the presence of this diatom species in the sediment can confirm nutrient rich conditions. There was also Fragilaria elliptica in the sediments of the estuary in 2009, which is a freshwater species (thus nutrient- liking) and at that time the estuary had been in a marine state for more than two years. This species was also found at sites in the vicinity of the sewage spill (at 2.1 and 7.4 km from the mouth) and possibly occurred there because of the input of the contaminated freshwater. The benthic microalgae as well as the fringing vegetation would absorb the nutrients from this point source input before it could be detected in the water column of the estuary. In all instances, Nitzschia frustulum was found in the estuary in the intertidal zone where it responded to these point source inputs.

Some of the benthic diatoms in the estuary responded to high ammonium and silicate concentrations, low SRP and TOxN, as well as the high salinity in 2008 and 2009. These species were Diploneis elliptica; Nitzschia flexa, Cocconeis scutellum and Amphora subacutiuscula. This can be attributed to the fact that during that period the catchment had been in a dry period, with less freshwater entering (thus highly saline) and contributing nutrients. Usually TOxN and SRP enter the estuary through river inputs and ammonium is generated within the estuary by nutrient cycling. Thus, during this dry period the source of nutrients in the estuary would have been the cycled ammonium and the silicate arising from the river (silicate concentrations had not changed in the estuary since 2006). As a result, these conditions favoured the oligotrophic (Watt, 1998; Taylor et al., 2007b) and fresh to brackish water diatoms such as Diploneis elliptica (Patrick and Reimer, 1975; Sims, 1996; Bate et al., 2004), and the brackish to marine Cocconeis. scutellum (Sims, 1996; Lange-Bertalot, 2000;

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Bate et al., 2004) and A. subacutiuscula (Lange-Bertalot, 2000; Bate et al., 2004). The dominant benthic diatom species in the estuary were either pollution tolerant species (i.e. Nitzschia frustulum, Navicula tenelloides; Taylor et al., 2007b) or had a wide range of salinity tolerance, like D. elliptica (fresh to brackish water), Amphora acutiuscula (brackish to marine salinity) and Planothidium delicatum (10 – 36 PSU) (Patrick and Reimer, 1975; Sims, 1996; Lange-Bertalot, 2000; Bate et al., 2004).

3.4.1.3 Epiphytic microalgae

The lowest epiphytic biomass was recorded in May 2008 and this could have been because epiphytes are less productive in winter than in summer. The case was the same in Nelson and Waatland (1997)‟s study, the lowest epiphyte biomass was recorded in winter and the highest in summer, ranging from 2.1 to 202.3 mg.m-2, respectively. Correlation analysis showed no statistical relationship between the epiphytic biomass and the physico-chemical variables of the estuary, suggesting that other factors could be controlling the epiphyte biomass of the estuary. Gordon et al. (2008) also found this in the St Lucia Estuary and attributed the biological factors, such as grazing and competition for resources between the host and epiphytes, to be more important than the physico-chemical environment in determining epiphyte biomass distribution. Alcoverro et al. (1997) also found that the biological factors played a bigger role in the variability of epiphyte biomass distribution. Findings from their study showed that grazing on the epiphytes by the micrograzers and on the shoots of the submerged macrophytes (that the epiphytes attach on) caused variability in the distribution of the epiphyte biomass along the estuary from one macrophyte bed to the next and within the beds. Seasonality in seagrass growth also played a role, as more epiphyte biomass was found in the older shoots, as the substratum for epiphyte colonization would have been available for a longer period (Alcoverro et al., 1997). In general, since the epiphyte biomass of the Bushmans Estuary did not show correlations with the measured environmental variables, there is a strong possibility that biological activity might be the controlling factor and further investigation into this is recommended.

The ordination analysis indicated that the distribution of the epiphytic diatom species (i.e. species composition) responded to different environmental conditions, depending on the state of the estuary during each sampling period. The ordination results showed that the epiphytic diatom species from the different years were controlled by different environmental variables because species from each year formed their own separate groups. The distribution of the 2008 group of species was mostly controlled by high salinity, ammonium and silicate concentrations and this was due to the lack of freshwater input into the estuary, which would bring in the TOxN and SRP nutrients that favoured the diatom distribution in 2006. In the 2006 group of species that were correlated with TOxN and SRP, Navicula ramosissima was the only species with a relative abundance of more than 10 % (most dominant); the abundance of the other species in this group were below 10 %. Gordon et al. (2008) also found this species at St Lucia Estuary and it was the only dominant species that showed a

125 positive correlation with TOxN concentration there as well. However, Nitzschia and Navicula spp. are not of epiphytic origin. They are considered as transitory species that usually settle on the top layers of the epiphytic community due to the reduced water flow rate (Meulemans and Roos, 1985; Muller, 1999).

Nitzschia frustulum which was a dominant benthic diatom species was also a dominant species in the epiphyte community in 2008 and 2009, therefore emphasizing the polluted nature of the estuary and that it was also evident in the species composition of the epiphyte community. Fallacia sp. was another dominant genus in the estuary and Taylor et al. (2007b) found this genus in a wide variety of aquatic conditions, some species tolerant to critical levels of pollution, which would explain why it was the third dominant species in 2009, after Synendra sp. and Cocconeis placentula v euglophyta.

Cocconeis placentula v euglophyta was the most abundant species in the estuary in 2006. It was associated with a high water flow rate because like C. scutellum, they form part of the adpressed component of the epiphyte community where these species position themselves against the leaf surface of the host plant. They stay attached to the host plant when flow rate is high compared to the species that extend upright on top of the adpressed layer. Taylor et al. (2007b) also found this diatom species in high abundance on plants, wood and stone.

According to Taylor et al. (2007b) C. placentula occurs in meso- to eutrophic waters, thus supporting the possibility of polluted/eutrophic conditions in the Bushmans in February 2009. The ordination analysis (CCA) showed that the species distribution was correlated with high temperature and low dissolved oxygen concentration. Low dissolved oxygen is usually a result of eutrophic conditions and species usually also require warm temperature under these conditions in order to grow optimally. This species was the second dominant epiphyte in 2009 (after Synendra sp.) and therefore its dominance was due to these favourable conditions for their growth. The most dominant species in the epiphyte community in 2009 was Synendra sp., and about 80 % of the species from this genus were found by Taylor et al. (2007b), in the benthos of meso- to eutrophic rivers and lakes. Taylor et al. (2007b) also described the species as being easily resuspended from the benthos due to their large surface area, which explains its presence in the epiphytic community at high numbers.

3.4.1.4 Salt marsh cover and distribution

A number of changes were noted when the macrophyte assessment from this study was compared with an assessment of the Bushmans Estuary by Oelofsen (1995) (Section 3.3.1.5). The decrease in Spartina maritima and possible replacement by Triglochin spp. and Phragmites australis, especially in the lower reaches, might be due to an increase in elevation of the intertidal area and seepage of freshwater respectively. At 0.9 km from the mouth S. maritima cover was 80% in 1995 and in 2009 it was just 10% and the reeds increased in cover to 35% of the area, including the area up to the intertidal zone. These reeds were

126 present because of freshwater input possibly from continuous seepage from septic tanks. Furthermore, the only other species at this site in 1995 was Sarcocornia sp. and in 2008 Disphyma crassifolium, Sporobolus virginicus, Limonium spp. and Bassia diffusa were present. These salt marsh species are characteristic of the upper intertidal and supratidal area (Adams et al., 1999). The sedimentation that has been noticed in the estuary might have contributed to an increase in the elevation gradient from subtidal to supratidal, leaving a reduced area for the intertidal S. maritima and creating a greater elevated level for the supratidal salt marsh to develop. Thus salt marsh zonation was evident here corresponding to the different levels of inundation (Boorman et al., 2001). Gray (1992) attributes elevation as the most important factor in salt marsh vegetation zonation, and the elevation results from sediment build up and stabilisation over time. Sediment stability influences macrophyte colonisation (Adams et al. 1999). Thus, the increase in species richness of the salt marsh at 0.9 km from the mouth and the increase in species richness of the other salt marsh areas in the lower reaches – where these areas now have supratidal species which were not there in 1995 – can be attributed to succession. The intertidal salt marsh plant S. maritima would have colonised the bare sand and mudflats and their sediment trapping function increased sediment stability which acts to reduce the impact of flow in the area (Pierce, 1982; Sanchez et al., 2001). Suitable environments were created for colonisation by subsequent salt marsh species according to their preferences of inundation (i.e. mean water high neap to mean high water spring species) (Dijkema, 1987; Adams et al., 1999).

Early in 2008 complaints were forwarded to the local municipality about sewage spillages occurring between Bushman‟s River Mouth and Riversbend, about 4 km from the mouth, from a conservancy tank. Raw sewage was noticed overflowing from the tank into a gully leading to the Bushmans Estuary (The Herald, February/January 2009). This can occur in holiday periods when sanitation systems reach their maximum carrying capacity due to a larger number of users. This is the area where the stands of reeds, sedges and rushes were noticed indicating that a brackish wetland has been created in this area due to anthropogenic effects. Plants indicative of brackish conditions occur here due to freshwater seepage with high loads of nutrients from the sewage spillages into this area. This plant habitat would perform an important nutrient filtering function, preventing high nutrients from reaching the estuary water column. These plants thrive in this environment even though the salinity of the tidal water can be above 30 PSU. Adams and Bate (1999a) showed that this is possible as the roots of the plants are located in low salinity sediment. The other areas where small pockets of reeds and sedges were found in the lower and middle reaches occurred at gully sites along the lateral boundaries of the estuary on the left bank. These gullies drain agricultural fields, which occur on this side of the estuary. Nutrient-rich seepage from these areas would support the growth of brackish plant species. This has also been recorded in the Kromme Estuary, which is also marine dominated but has pockets of reeds in areas with freshwater seepage occurs because the root system is located in fresh or brackish water (Adams and Bate, 1994b). Thus, the reeds in the Bushmans Estuary can be used as indicators of sites of freshwater seepage. In the areas with no freshwater seepage, where there was no agricultural activity in the proximity of the estuary and no house developments with sanitation problems

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(mostly the upper reaches), the reeds were struggling and looked unhealthy especially in the dry summer months, as was noted in February 2009 (Figure 3.20). In most studies Phragmites australis is said to have winter die-back (Allirand and Gosse, 1995; Choy et al., 2008). Kadlec and Knight (1996) indicated that for plants such as Typha spp. and P. australis the obvious annual cycle of aboveground biomass would be new shoots starting from zero biomass in early spring and growing at a maximum rate in spring and early summer. Late summer is a period of reduced growth, and complete shoot die back occurs in the fall. Adams et al. (1999) also mentioned that P. australis and Schoenoplectus triqueter can die back in summer, releasing particulate matter into the water. Consequently the presence of the unhealthy looking reeds noted in the estuary in February 2009 in the upper reaches. This natural phenomenon was probably exacerbated in the Bushmans Estuary because, additional to the estuary receiving close to zero freshwater and being highly saline, more evaporation occurs in summer due to the high temperatures which increases salinity thus stressing the plants. The salinity was highest in the hottest month of February 2009 (average 35.5 PSU) compared with the other months (24.1; 32.7; 30.8 PSU), probably due to increased evaporation. Additionally, it has been shown that P. australis dies completely when inundated for 94 days at 30 PSU water (Benfield, 1984). This is because, increased salinity generally results in overall reduced plant performance and shoot height (Adams et al., 1999) because of a diversion of energy away from active meristematic growth to the maintenance of osmotic balance (Hellings and Gallagher, 1992). The reeds in the upper reaches of the Bushmans Estuary were dying back because of salinity stress unlike the reeds in the lower and middle reaches at the sites of anthropogenic disturbance and nutrient-rich freshwater seepage.

3.4.1.5 Zostera capensis biomass

During its life time, Z. capensis continually forms new leaves, and old leaves become senescent and detach, especially in late summer (Verhagen and Nienhuis, 1983). This could have been what happened in the estuary in February 2009 (which was late summer), resulting in an overall decrease in biomass compared to samples collected in mid-spring, October 2008. Figure 3.6 illustrates that the temperature of the estuary was higher in February 2009 than the other months, with a mean of 26.8oC compared with 19.0, 19.0 and 21.3oC measured during the other months. Thus, the leaves that were collected in February were the shorter, young leaves with lower biomass than that which the old leaves would have shown. Another attribute would have been the recent end of the December/January holiday season, which Hooker (1996) discovered to have a big impact on Z. capensis biomass due to the high number of people partaking in recreational activities. Forbes (1998) and Hooker (1996) found that bait digging and people walking on Zostera beds disturbs the plants and this needs to be examined because recreation is a major activity in the Bushmans Estuary, especially in the December/January and April (Easter) holiday season. Hooker (1996) also found that boating activities (especially power-boating) resuspend sediment in the Bushmans and caused an increase in the turbidity of the water as well as damage to the submerged macrophytes. Thus during the February sampling, the biomass of the submerged macrophyte could have been

128 affected by a number of factors together due to the use of the estuary prior to the February sampling trip. Documentation of the human activities in the estuary also showed a higher number of active people in the estuary in February 2009 than that recorded in the October off-season period.

Hodgson (1986) recorded that Z. capensis only penetrated the Bushmans Estuary 5 km up the estuary. However, in the current study the macrophyte has been recorded in the lower, middle, and upper reaches of the estuary and this could be accredited to the homogenous marine-dominated nature of the estuary which could have led to colonisation by the submerged macrophyte over time. Hodgson (1986)‟s findings might be doubtful because Robertson (1984) had indicated that marine conditions prevailed even then and salinity gradients were only evident in the event of major rainfall floods. However, there was a flood in November 1985 that removed Zostera sp. in the Kwelera Estuary (Talbot et al., 1990). Since this estuary is also in the Eastern Cape (East London area) and not that far away from the study site, it is thus expected that the same intensity floods were experienced in the Bushmans Estuary which might have also led to the submerged macrophyte only being detected up to 5 km during Hodgson‟s study. Flow data from the Department of Water Affairs, measured at the flow gauge station in the upper Bushmans River, showed that in November and December 1985 the peak flows were 22.5 and 31.2 m3.s-1 respectively. This showed that the Bushmans River experienced the same intensity floods that removed submerged macrophytes in Kwelera Estuary. Adams et al. (1999) also stated that submerged macrophytes are sensitive to flows greater than 1 m.s-1, which result in their removal, while they grow optimally at flows less than 0.1 m.s-1. These are the types of flow that prevail in the Bushmans Estuary, and thus, would also contribute together with the homogenous marine conditions to the colonisation of this submerged macrophyte into the upper reaches of the estuary.

Z. capensis is usually dominant in permanently open tidal estuaries because of its ability to survive daily periods of exposure (Adams and Bate, 1994b) and thrives under the saline conditions (30 PSU) of the lower reaches in these estuaries. However, reduced freshwater inflow has increased the water column salinity to the extent that submerged marine macrophyte communities have encroached to the upper reaches and displaced brackish communities (Wortmann et al., 1997), which is why they occur up to the upper reaches of the saline Bushmans Estuary.

According to Adams et al. (1992) and Adams and Bate (1994b) Ruppia cirrhosa is the dominant submerged macrophyte in temporarily open estuaries that are characterized by fluctuating salinity, however, it can also occur in the calm, brackish upper reaches of permanently open estuaries, and this was where the plant was found in the Bushmans Estuary in the current study. This is expected because the estuary is calm, except it was not brackish in February 2009 when the plant was found, but rather saline and even higher than the rest of the estuary at this site (with 37.8 PSU salinity). Laboratory experiments done by Adams

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(1994) showed that this macrophyte can actually tolerate high salinity, with new growth still recorded at 55 and 75 PSU.

The results have shown that you get similar results to the dry weight method when measuring Z. capensis biomass using the leaf length method. Thus, in order to decrease the destruction of the submerged macrophyte habitat, it would be more suitable to use the leaf length method for biomass measurements in future studies particularly those that require frequent monitoring.

3.4.1.6 Human activities

Most of the human activities take place in the lower reaches of the estuary because development in the Bushmans River is concentrated around the mouth, with Kenton-on-Sea on the east bank and Bushmans River Mouth on the west which are both popular holiday resorts (Forbes, 1998). This is also the area of highest habitat diversity and therefore this region is vulnerable to impacts, especially the salt marsh area and seagrass beds. Bait digging for mud prawns and pencil bait can disturb the salt marsh and seagrass habitat as well as impact on the bait resource.

Surveys were conducted which counted the number of people undertaking various activities in the estuary. Forbes (1998) found that the majority of people in the Bushmans Estuary were involved in swimming (24.2 %) and then angling (20.9 %). Meanwhile, in the current study the majority mostly participated in fishing (28.1 %), followed by boating (19.5 %) and then swimming (18.8 %). Results from Forbes‟ questionnaires showed that angling was the first choice/preference for people in the Bushmans Estuary; meaning that it was the first reason that people went to the estuary. Approximately 87 % of the estuarine anglers admitted to collecting their own bait (Forbes, 1998). This could be the reason why bait digging was found to be the second highest activity (21.9 %) in the counts for 2008 and 2009.

Forbes (1998) recorded a municipal jetty and slipway, 23 private jetties and a small boat marina near the bridge. In the current study 69 jetties were recorded along the length of the estuary. There has been a rapid increase in development and recreational use of the Bushmans Estuary.

3.4.2 Change in the botanical importance of the estuary

The earlier aerial photographs (i.e. 1942, 1966, 1973) show large areas of exposed bare sediment in the lower estuarine reaches, mostly marine sand. Reduced flooding and an increase in sediment stability allowed for salt marsh and seagrass colonisation. These communities add more value to the estuary because of their contribution to productivity. Compared to the open water phytoplankton‟s productivity of 163 g.m-2.yr-1, salt marsh and submerged macrophytes respectively contribute 1487 and 1316 g.m-2.yr-1 to the estuary (these

130 are average productivity values from Colloty et al., 1999). Sedimentation and reduced flow rate triggered the habitat changes over time. The following events might have particularly contributed to these changes. Firstly, Baird et al. (1981) detected that the estuary experiences marine sediment input due to marine sand moving up and down the channel as bed-load during flood and ebb tides. They identified that an average of 20 m3 of sand is transported into the estuary over a single spring tidal cycle (Baird et al. 1981). According to Baird et al. (1983) building the R72 bridge would prevent flushing of fine sediment because before it was built spring tides and freshwater spates were able to overtop the „island‟ and thus flush away fine sediment. Now, the central part of the bridge sits on this island and prevents flushing. Thus a build-up of fine material immediately up and downstream of the bridge and stable sediment conditions was expected. The distribution of the sediments along the estuary is also a factor to consider. The sediments 3.5 km up the estuary consist almost entirely of mud with a small component of sand, of terrestrial origin (Gerber, 1992), which would be favourable for submerged macrophyte establishment and growth. Also, the New Years River Dam was constructed in 1959 and many of the farm dams in the area were also constructed after the chosen reference state, 1942. Thus, it can be assumed that when all these dams were built, the flow rate of incoming freshwater decreased thus leading to stable sediments because of reduced flushing by floods. The R72 Bridge also decreased the flushing effect of freshwater and thus stabilized the sand banks on the lower side of the bridge. So, the now stable sand banks became suitable environments for salt marsh to establish, which is why most of the increase in salt marsh after 1942 occurred in the sandy lower reaches of the estuary. Seagrass expansion and encroachment from the lower to upper reaches could also be related to the increase in sediment stability, as well as the high salinity of the upper estuary.

Recently reed and sedge habitats have developed in some areas in the lower reaches due to the anthropogenic effects of nutrient rich freshwater seepage. There is no industrial input into the system, though phosphorus and nitrogen may be washed into the river in run-off from fertilised agricultural lands (as well as herbicides and pesticides) (Palmer, 1980). This explains the occurrence of the reeds and sedges in the other areas in the highly saline lower reaches where there is no evident sewage contamination. Previous studies show that P. australis forms dense beds in the brackish upper reaches (<15 PSU) of South African estuaries that have a gradient of decreasing salinity up the length of the estuary (Adams et al., 1992). In marine dominated estuaries, salinity is uniform (35 PSU) from the mouth to the head of the estuary. In these systems, P. australis beds have been found only at the confluence of small freshwater streams and seepage areas flowing into the estuary. The plants are, however, tidally inundated with saline water (Adams and Bate, 1999a). This has been recorded in the marine dominated Kromme Estuary, in which the reed is tidally inundated by 35 PSU salinity water. This has been possible because it was also found that the root system was immersed in fresh or brackish water (Adams and Bate, 1994b). Thus, groundwater seepage plays an important role in marine-dominated estuaries like the Kromme and Bushmans as this creates points of biotic diversity where both brackish and saline species may occur. The same findings were recorded in the Goukou and Keurbooms estuaries, where

131 pore water salinity was lower than surface water salinity within Phragmites stands and a decrease in the height of P. australis was associated with an increase in the pore water salinity towards the water‟s edge (Adams and Bate, 1999a).

Nutrients trapped in fine textured sediment have contributed towards greater above ground biomass of Cyperus esculentus and Scirpus validus compared with coarser sediments (Barko and Smart, 1978) because sandy sediments have a lower ability to trap nutrients than fine sediment. Thus, caution should be taken towards increased anthropogenic nutrient increases, which end up being trapped in sediments and can contribute to monospecific stands of emergent macrophytes that reduce species diversity. This has been the case in the Bushmans, where there are known nutrient seepages into the estuary (i.e. sewage leakages and agricultural runoff) but not evident in the water column. Thus proving that they are trapped in the sediment and benefiting the growth of organisms that get their nutrients directly from these sediments (i.e. benthic microalgae, reeds and sedges), while the plankton are disadvantaged.

3.5 Conclusion

The lack of continuous freshwater input into the estuary has led to the greatest changes in the estuary over time; low flow and reduced flooding has reduced vertical and horizontal salinity gradients. These gradients are required to ensure species biodiversity and a functioning healthy ecosystem. The well-mixed, nutrient-depleted (or oligotrophic), marine nature of the estuary has led to poor species diversity in the water column, which was largely dominated by flagellates and low microalgal cell numbers.

The fact that no algal blooms were recorded in the water column (indicator of nutrient rich conditions), and that the pollution of the Bushmans Estuary could only be detected from the sediment (benthic) diatom species was strange. This could be due to the large influence of seawater in the estuary, which might quickly dilute the seeping polluted water before it is used up phytoplankton. Then again, due to the clay nature of most of the sediment in the estuary, this pollution gets retained in the sediment, where the pollution tolerant species get to colonise and dominate over the sensitive species.

Increased recreational activity in the Bushmans Estuary is of concern as at time too many people are using the system and the carrying capacity of the estuary is exceeded. This occurs during peak seasons such as the Christmas and Easter holidays. Many boats and those travelling at high speeds can lead to disturbance of the Zostera capensis beds and erosion of the salt marshes.

The hypotheses for the study were accepted. These were:

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Freshwater abstraction and an increase in salinity have increased the distribution and cover of the seagrass, Zostera capensis, which is usually found in the saline lower reaches. Lack of floods and increased freshwater abstraction has increased sediment stability, which has allowed the seagrass to colonise the entire length of the estuary Marine sedimentation in the lower reaches of the estuary has increased available habitat for salt marsh, which has increased in cover over time. This has compensated for the loss of habitat due to jetties and retaining walls. Low freshwater input and associated nutrients has decreased the species richness and biomass of microalgae.

Zostera capensis cover has increased and this plant has encroached into the upper reaches due to the saline nature of the estuary and the lack of floods, which have increased sediment stability over time and allowed or favoured colonisation by the submerged macrophyte. In the lower reaches of the estuary there has been marine sediment accumulation due to reduced flooding. Aerial photograph analysis showed that in some areas the sediment has stabilised and become colonised by salt marsh. Overall, salt marsh cover has increased in the estuary.

Low freshwater input has decreased the richness and biomass of the phytoplankton as low biomass was measured in the estuary together with low nutrient concentrations in the water column. However, with the known contamination (sewage leakages and agricultural return flows) and the high levels of nutrients that come with it, there should be phytoplankton blooms but this was not evident. That then led to the belief that the muddy/clay nature of the sediment in the estuary has led to the trapping of this contamination in the sediment thus preventing the nutrients from entering the water column; since these types of soils have a high holding capacity due to their fine particle size. This has been indicated by the presence of the pollution tolerant benthic diatom species and the occurrence of pockets of reed and sedge communities (usually occurring in fresh to brackish waters) in the highly saline lower reaches of the estuary, showing that they are supported by the nutrient-rich freshwater in the sediment they are established in.

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Chapter 4: Application of the Estuarine Health Index and identification of monitoring indicators for the Bushmans Estuary

4.1 Introduction

Monitoring is important in providing feedback to ensure effective management of water resources (Adams and McGwynne, 2004). When the health status of resources is not adequately protected, the managers need to know whether this results from inadequate implementation, inadequate standards, or inadequate strategies (Granger et al., 2005). The Regional Water Affairs office in East London wants to initiate a monitoring programme in the Bushmans Estuary. This monitoring is required for the provision of baseline information in order to understand how this estuary functions before any activities such as dredging can be allowed. Baseline data are also needed so that an effective management programme can be initiated in the long term.

A successful monitoring protocol would be one that is based on a minimum number of easy to implement indicators that can adequately reflect core issues (Adams and McGwynne, 2004). However, before a monitoring programme can be implemented it is important that the present status (health) of the estuary is assessed. This is because, the starting point of any monitoring programme is to assess the current condition or status of an estuary relative to a natural or normal non-degraded state, thus the assessment of the health status of the Bushmans Estuary is an important first step.

This study has provided detail on the botanical characteristics of the estuary which can provide input to the assessment of estuarine health. Following the estuarine health index guidelines, a preliminary assessment of the estuary‟s health was also made. This includes collating data and literature on the biological, physical and chemical characteristics of the estuary. The available literature and data for the abiotic and biotic components (invertebrates, fish and birds) of the Bushmans Estuary was summarised and presented in the standard Estuarine Health Index tables (DWAF, 2008a). From this preliminary assessment of estuarine health, monitoring indicators for the Bushmans Estuary were identified. The significant impacts and problems in the estuary were taken into consideration so that the most important and integrative indicators were chosen from the existing monitoring protocols in South Africa that would be suitable for the Bushmans Estuary.

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4.2 Materials and Methods

4.2.1 Data Availability

The Resource Directed Measures method for determining the environmental water requirements of estuaries sets out the method for determining the health of an estuary using the Estuarine Health Index (DWAF, 2008a). The accuracy and confidence of the health assessment depends on the available data for the estuary. Tables indicating the data required to conduct an Estuarine Health Assessment and Environmental Water Requirement study are shown in Table 4.1-4.6. These tables also indicate the available data for the Bushmans Estuary. These tables were completed after a thorough review of the available literature and they can be used to identify data gaps.

4.2.2 Estuarine Health Index

The available data on the Bushmans Estuary were summarised and tabulated in the Estuarine Health Index (EHI) tables as described in DWAF (2008a). The Present Ecological Status is a measure of the health of an estuary based on a comparison between the reference (natural condition) and the present state. The reference condition is when the estuary received 100% of the MAR and when there was no development in the catchment or surrounding estuary area. The Estuarine Health Index includes the abiotic and biotic characteristics of the estuary. These include the hydrology, mouth condition, water quality, physical habitat and biotic health of the microalgae, macrophytes, invertebrates, fish and birds.

For each variable the conditions are described as quantitatively as possible as a percentage of the pristine (natural) state. Each variable is then weighted and aggregated using the prescribed rules (DWAF, 2008a). This approach usually includes a multi-disciplinary team of people and therefore for this exercise only the relevant data were included in the Tables (Tables 4.7-4.15) and no quantification was done for the Bushmans Estuary.

4.3 Results

4.3.1 Data Availability

The available data required for the assessment of the health of the Bushmans Estuary is indicated in Tables 4.1-4.6.

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Table 4.1 Data availability on sediment dynamics, hydrodynamics and water quality

DATA REQUIRED AVAILABILITY COMMENT Simulated monthly runoff data (at the head of the Measured flow data from This gauge station occurs close to the town of estuary) for present state, reference condition and a DWA gauge station Alicedale, which is 150 km above the head of the the selected future runoff scenarios over a 50 to 70 (P1H003) Bushmans Estuary, and is too far from the year period estuary to be able to properly represent the freshwater inflow. Modelling studies would be needed to provide some hydrological data. Simulated flood hydrographs for present state, No data reference conditions and future runoff scenarios: 1:1, 1:2, 1:5 floods (influencing aspects such as floodplain inundation) 1:20, 1:50, 1:100, 1:200 year floods (influencing sediment dynamics) Series of sediment core samples for the analysis of Robertson (1984) Robertson (1984) reported on sediment particle size distribution (PSD) and origin (i.e. using Jafta (current study) distribution. Jafta has PSD for the intertidal and microscopic observations) taken every 3 years along subtidal for 2006, 2008 and 2009. the length of an estuary (200 m to 2 km intervals). Series of cross-section profiles (collected at about Reddering and Rossouw (2003) has cross sections along the 500 to 1000 m intervals) taken every 3 years to Esterhuysen (1981), estuary but this was a once off survey. quantify the sediment deposition rate in an estuary. Baird et al (1981) Rossouw et al. (2003) Aerial photographs of estuary (earliest available year Surveys and Mapping, 1942, 1966, 1973, and 2004 aerial photographs as well as most recent) Mowbray, Cape Town available. Measured river inflow data (gauging stations) at the Gauge station P1H003 Available from DWA head of the estuary over a 5-15 year period Continuous water level recordings near mouth of the No data estuary Water level recordings at about 5 locations along the Rossouw et al. (2003). Records of water level along the length of the length of the estuary over a spring and a neap tidal estuary from the mouth to the upper reaches (at cycle (i.e. at least 14 days) about 20 km). Longitudinal salinity and temperature profiles (in Palmer (1980), Bornman and Klages’ and Palmer’s data are only situ) collected over a spring and neap tide during Robertson (1984), for the lower reaches of the estuary (below the high and low tide at: end of low flow season (i.e. Bornman and Klages R72 bridge). period of maximum seawater intrusion) and peak of (2004), Robertson’s data is up to the upper reaches and high flow season (i.e. period of maximum flushing by Jafta (current study, is representative of low flow and high flow river water) 2006, 2008, 2009) periods. Jafta: The 2006 data is from a sampling session after high rainfall and flooding. 2008 and 2009 represent low flow. Water quality measurements (i.e. system variables, Jafta (current study) Data available for high flow (2006) and low flow and nutrients) taken along the length of the estuary (2007 and 2008). (surface and bottom samples) on a spring and neap high tide at the end of low flow season and peak of high flow season Measurements of organic content and toxic Watling and Watling Measurement of trace metals in the sediment substances (e.g. trace metals and hydrocarbons) in (1982) along the estuary which were found to be low at sediments along length of the estuary. that time as there were no industrial inputs. Water quality (e.g. system variables, nutrients and In the current study the estuary was saline 32 km toxic substances) measurements on river water upstream from the mouth where it was no longer entering at the head of the estuary navigable. Measurements were not made for the head of the estuary. Water quality (e.g. system variables, nutrients and Jafta (this study, 2006, Salinity, pH, nutrients, dissolved oxygen, toxic substances) measurements on near-shore 2008, 2009) temperature, electrical conductivity, nutrients seawater

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Table 4.2 Data availability on microalgae

DATA REQUIRED AVAILABILITY COMMENT Chlorophyll-a measurements taken at 5 stations at the surface, 0.5 Jafta (this study, 2006, 2008, Data represents m and 1 m depths. Cell counts of dominant phytoplankton groups 2009) phytoplankton chl-a and i.e. flagellate, dinoflagellates, diatoms and blue-green algae. dominant phytoplankton Measurements must be taken coinciding with typically high and groups for all the sampling low flow conditions. trips. Intertidal and subtidal benthic chlorophyll-a measurements taken Jafta (this study, 2006, 2008, Data includes chl-a and at 5 stations (at least). 2009) diatom identification. No benthic diatom data for May Epipelic diatoms need to be collected for identification. 2008.

These measurements must to be taken coinciding with a typical high and low flow condition (in temporarily closed estuaries measurements must include open as well as closed mouth conditions). Simultaneous measurements of flow, light, salinity, temperature, Jafta (this study, 2006, 2008, Salinity, temperature, nutrients and substrate type (for benthic microalgae) need to be 2009) nutrients, light attenuation, pH taken at the sampling stations during both the phytoplankton and and sediment particle size. benthic microalgal surveys. Flow was not recorded.

Table 4.3 Data availability on macrophytes

DATA REQUIRED AVAILABILITY COMMENT Aerial photographs of the estuary (ideally 1:5000 scale) reflecting Surveys and Mapping, Photos for 1942, 1966, 1973, the present state, as well as the reference condition (if available) Mowbray, Cape Town 1990s and 2004 are available Available orthophoto maps Number of plant community types, identification and total number of Adams (1995) Plant community types and macrophyte species, number of rare or endangered species or Oloefson (1995) dominant species were those with limited populations documented during a field visit. Jafta (current) described for each community. Permanent transects: No data Measurements of percentage plant cover along an elevation gradient Measurements of salinity, water level, sediment moisture content and turbidity

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Table 4.4 Data availability on invertebrates

DATA REQUIRED AVAILABILITY COMMENT Compile a detailed sediment distribution map of the Robertson (1984), Robertson (1984) reported on sediment estuary. Obtain a detailed determination of the Jafta (this study: 2006, distribution. Jafta has PSD for the intertidal and extent and distribution of shallows and tidally 2008, 2009) subtidal for 2006, 2008 and 2009. exposed substrates. During each survey, collect sediment samples for analysis of grain size and organic content at the six benthic sites. Surveys to determine salinity distribution pattern Jafta (this study, 2006, Jafta: Salinity, temperature, nutrients, light along the length of the estuary, as well as other 2008, 2009) attenuation, pH and sediment particle size (2006 system variables (e.g. temperature, pH and Palmer (1980), Bornman sampled after high rainfall and flooding, 2008 dissolved oxygen and turbidity) are required for and Klages (2004), and 2009 represent low flow) different seasons and for different states of the tide Robertson (1984) Bornman and Klages: salinity (only for the lower 3. Seasonal (i.e. quarterly) physico-chemical data reaches of the estuary, below the R72 bridge) are also required for each of the six benthic Palmer: temperature and salinity (only for the sampling sites lower reaches of the estuary, below the R72 bridge). Robertson: temperature and salinity (up to the upper reaches and is representative of low flow and high flow periods). Collect a set of six benthic samples each consisting Hodgson (1986), Forbes The information relates to bait organisms, no of five grabs. Collect two each from sand, mud and (1998) information is available for zooplankton. Forbes interface substrates. If possible, spread sites for (1998) assessed the population density of each between upper and lower reaches of the Upogebia africana, Callianassa kraussi, Solen estuary. One mud sample should be in an spp. and Arenicola spp (quadrat method was organically rich area. Species should be identified used). to the lowest taxon possible and densities (animal/m2) must also be determined. Seasonal (i.e. quarterly) data sets for at least one year are required, preferably collected at spring tides. Collect two sets of beam trawl samples (i.e. mud No data and sand). Lay two sets of five, baited prawn/crab traps overnight, one each in the upper and lower reaches of the estuary. Species should be identified to the lowest taxon possible and densities (animal/m2) must also be determined. Survey as much shoreline as possible for signs of crabs and prawns and record observations. Seasonal (i.e. quarterly) data sets for at least one year are required, preferably collected at spring tides. Additional trip(s) may be required to gather data on No data the occurrence/recruitment and emigration of key species such as Varuna litterata and Upogebia which require a connection to the marine environment at specific times of the year. Collect three zooplankton samples, at night, one No data each from the upper, middle and lower reaches of the estuary. Seasonal (i.e. quarterly) data sets for at least one year are required, preferably collected at spring tides.

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Table 4.5 Data availability on fish

DATA REQUIRED AVAILABILITY COMMENT In a larger estuary (>5km) sampling can either be at fixed intervals (every 2km) or Harrison et al., Provides information on have the upper, middle and lower reaches subdivided into at least a further three 2000 the changes in species sections each. The samples should be representative of the different estuarine richness and abundance habitat types, e.g. Zostera beds, prawn beds, sand flats. At least one of the sample from the reference sets should be in the 0 to 1 PSU reach of the system. Sampling should be condition. representative of small fish (seine nets) and large fish (gill nets).

Sampling should be done during both the low and the high flow season for the full extent of the system (as far as tidal variation) to allow for predictive capabilities.

Table 4.6 Data availability on birds

DATA REQUIRED AVAILABILITY COMMENT Undertake one full count of all water-associated birds, covering Coordinated Waterbird Counts Accessible from the CWAC website. Bird as much of the estuarine area as possible. All birds should be (CWAC). From the mouth to Ghio counts at Bushmans are said to be done identified to species level and the total number of each Bridge (about 15 km from the irregularly; the last count was done on counted. mouth). 07/02/2008. 9 summer and 8 winter counts have been Monthly data sets for at least one year are required. It this is done. not possible, a minimum of four summer months and one winter month will be required (decisions on the extent of effort required will depend largely on the size of the estuary, extent of shallows present, as well as extent of tidally exposed areas).

4.3.2 Estuarine Health Index

The available data on the Bushmans Estuary were summarised and tabulated in the Estuarine Health Index (EHI) tables (Tables 4.7-4.15) as described in DWA (2008a). No calculation of scores was done as a more comprehensive assessment by a multi-disciplinary team is required.

Table 4.7: Hydrological health score

VARIABLE COMMENTS Reddering and Esterhuysen (1981) MAR = 38 mil m3/a. a. Present MAR as a % of MAR in the reference state Modelling studies needed to compare with present state % similarity in frequency of major floods (floods 1:20 year for a particular system) b. OR % similarity in the magnitude of major floods (e.g. Hydrological modelling required to assess these changes. 1:20, 1:50 and 1:100) in comparison with the reference condition

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Table 4.8: Mouth condition score

VARIABLE COMMENTS Change in mean duration of closure, e.g. over a 5 or 10 year period The mouth has always been open.

Table 4.9: Water quality health score

VARIABLE COMMENTS Long-term reductions in freshwater have changed the estuary from that % Change in axial salinity gradient and characterised by salinity gradients to a homogenous marine system vertical salinity stratification (Bornman and Klages, 2004). Scoring guideline: Unmodified = 100; largely 1 OR salinity has always been high since the estuary is in an area with low natural = 80; moderately modified = 60; MAR, dry catchment. Hydrological and hydrodynamic modelling is needed largely modified = 40; seriously modified = to confirm this. Small floods do reset the estuary and created gradients, 20; completely modified = 0. but only for a short while. Due to the reduced freshwater input the estuary is oligotrophic with below Nitrate and phosphate concentrations in detectable concentrations of the nutrients, unless a flood event occurs. estuary Despite the agricultural activity in the catchment nutrient input is low Scoring guideline: Unmodified = 100; 2a because it gets trapped in the intertidal mud and gets assimilated by the reduced = score is estimated % of original fringing macrophytes and benthic microalgae. There are known sewage level; slightly increased = 75; moderately leakages, which are sources of high nutrients but their effect has not been increased = 50; eutrophic = 0. detected in the water column. High TDS has been recorded in the river (which could be the source in the estuary) and is said to be due to salt loading from the Nanaga and Suspended solids in inflowing freshwater Weltvred formations that the river flows through. Doudenski (2004) said Scoring guideline: Unmodified = 100; slightly 2b that this has greatly deteriorated the water quality of the river, with TDS as increased = 75; moderately increased = 50; high as 4000 mg.l-1 compared to 200 mg.l-1 in the head waters. The heavy load = 25; excessive siltation = 0. agricultural and degraded land along the lateral boundaries could also be a source of suspended sediment. Dissolved oxygen in the estuary can reach close to hypoxic conditions (< 2 mg l-1). Firstly, during the low flow periods oxygen concentrations decrease Dissolved oxygen (mg l-1) concentrations in from the low to the upper reaches (reaching 3.5 to 4.5 mg.l-1 at the head) estuary as less mixing and oxygen exchange occurs, due to the reduced Scoring guideline: Unmodified = 100; largely C freshwater input. Also, results obtained after the high flow event in 2006 natural = 80; moderately modified = 60; show that, generally, the dissolved oxygen of the estuary would be low largely modified = 40; seriously modified = (averaging 4.37 mg.l-1 throughout) when sufficient freshwater enters the 20; completely modified = 0. estuary as the microalgal activity is enhanced by the abundant nutrients that enter and thus use up more of the oxygen in the water column. Level of toxins Scoring guideline: Unmodified = 100; largely There might be pesticide and herbicide input through return flow from the D natural = 80; moderately modified = 60; farming and agricultural land adjacent. largely modified = 40; seriously toxic = 20; completely toxic = 0.

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Table 4.10: Physical habitat health score

VARIABLE COMMENTS Resemblance of Intertidal sediment structure 1 and distribution to reference condition Sedimentation has potentially increased the Intertidal area (Reddening and Esterhuysen, 1981). Jetties, retaining walls, 1a % Similarity in Intertidal area exposed parking areas and other developments have disturbed approximately 15% of the intertidal habitat. There is more sand now due to deposition of marine sand into the estuary (Reddering and Esterhuysen, 1981) and mud banks have % Similarity in sand fraction relative to total been displaced by the sedimentation. 1b sand and mud This is only evident in the lower reaches, the middle and upper reaches are characterised by silty sediment, contributed by the catchment. Resemblance of subtidal estuary to reference condition: depth, bed or channel There has been marine sedimentation, which might have made the 2 morphology estuary shallow in the lower reaches, and there is a bridge, which Scoring guideline: No alteration = 0%, Total might have changed the configuration of the benthic community. alteration = 100%. Overall physical habitat health = Weighted mean

Table 4.11: Biotic health score for microalgae

VARIABLE MEASUREMENT COMMENTS Estimated % of original species remaining Since the estuary has become highly saline, it is Scoring guideline: 100% = 100, 90% = 80; a. Species richness expected that some of the original species have been 80% = 65; 70% = 50, 60% = 35; 50% = 25; lost and salinity tolerant species are dominant. 40% = 17; 30% = 10; 20% = 5; 10% = 0 Biomass decrease expected because of low freshwater Estimated % of total biomass remaining of b. Abundance input and thus decreased nutrients. Disturbance of the the original species benthic habitat due to human activities such as boating. The phytoplankton near the mouth is expected to be of Estimated % resemblance to original marine origin, whereas the head might have species of composition. c. Community freshwater origin. The middle reaches of the estuary will Scoring guideline: No change = 100% composition be expected to have more salinity tolerant species. Original community totally displaced by Community expected to be mostly composed of opportunistic spp = 0% flagellates with a potential loss of brackish species. Community health score = minimum score of a, b and c

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Table 4.12: Biotic health score for macrophytes

VARIABLE MEASUREMENT COMMENTS a. Species Estimated % of original Probably most of the original species are still present, but at different richness species remaining abundance. For example, there is more seagrass than before. Currently seagrass extends to the upper reaches of the estuary due to the marine-like condition of the estuary and little disturbance of the sediment from the lack of flooding. Under reference conditions the distribution would have been limited to the lower saline reaches. Estimated % of total Pockets of reeds have been detected in the lower reaches due to point and non- b. Abundance biomass remaining of point sources of pollution into the estuary. Reeds are usually limited to the fresh the original species to brackish upper reaches of the estuary. There has been an increase in salt marsh area in the lower reaches due to the increase in marine sedimentation and stabilization of the sediment due to reduced flooding. Macrophyte distribution has changed greatly from reference conditions. Estimated % c. Community Attributed to seagrass distribution to the upper reaches, pockets of reed resemblance to composition communities at seepage sites in the lower reaches, greater salt marsh cover due original composition. to sediment stabilisation. Community health score = minimum score of a, b and c

Table 4.13: Biotic health score for invertebrates

VARIABLE MEASUREMENT COMMENTS Most of the original species are probably still present but at different abundances, depending on habitat and food. The macrobenthic species diversity might be greater now Estimated % of because of the increase in seagrass cover. Marine sedimentation might increase diversity a. Species original species of sand-dwelling species. richness remaining Bait collection might influence species richness, due to exploitation of certain species (Forbes, 1998). The exploited species were found to be Upogebia africana, Callianassa kraussi, Solen spp. and Arenicola spp. Impacts have increased over time due to an increase in subsistence and recreational use. Estimated % of Zooplankton biomass expected to have decreased due to the low phytoplankton biomass. total biomass b. Abundance However, the possibility of the presence of more sand-dwelling species and macrobenthic remaining of the species might increase biomass. original species

Invertebrate composition depends on the food source, microalgae, and on habitat Estimated % availability, submerged macrophytes and sediment. These have changed in the Bushmans resemblance to c. Community Estuary and might have influenced the invertebrates. There might be less filter feeders due original composition to decreased phytoplankton biomass. Seagrass beds have created more habitat, thus composition. increasing species diversity. Due to sedimentation, there might be more sand-dwelling

species at the expense of a mud community (DWAF, 2008b).

Community health score = minimum score of a, b and c

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Table 4.14: Biotic health score for fish

VARIABLE MEASUREMENT COMMENTS Some of the original estuarine fish species probably died due to a. Species Estimated % of original species the highly saline conditions e.g. river pipefish. Harrison et al. richness remaining (2000) showed that 84 % of the fish assemblage is similar to reference assemblage Estimated % of total biomass 54 % species relative abundance similar to reference (Harrison b. Abundance remaining of the original species et al., 2000)

c. Community Estimated % resemblance to original Community probably composed of more marine species at the composition composition. expense of freshwater and brackish species. Community health score = minimum score of a, b and c

Table 4.15: Biotic health score for birds

VARIABLE MEASUREMENT COMMENTS a. Species Estimated % of original species Greater species richness is expected due to the wider range of richness remaining habitat along the estuary Estimated % of total biomass b. Abundance Recreational use could disturb birds and reduce abudance. remaining of the original species A large percentage of the community should still resemble the reference community because all the different types of habitats are still there but at different cover compared with the reference c. Community Estimated % resemblance to original condition which would change the different types of food composition composition. available for the birds (i.e. more of a certain type of invertebrates or fish than others due to increases and reductions in their food availability and habitat). Community health score = minimum score of a, b and c

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4.4 Discussion

4.4.1 Change in the health of the estuary

This study has collated available data on the Bushmans Estuary and indicated how it can contribute to the assessment of Estuarine Health. This assessment has also highlighted important data gaps such as long-term monitoring to assess sedimentation. The preliminary assessment of the estuary‟s health indicated some problem areas. Prolonged periods of freshwater inflow less than 0.02 m3.s-1 are now common. The reduced freshwater input deprives the estuary of nutrients, which are especially required by the primary producers (i.e. microalgae, macrophytes).

The prediction was that reduced freshwater input results in a homogenous well-mixed condition that lacks salinity gradients and available nutrients. This then reduces phytoplankton diversity and biomass because different phytoplankton species are associated with different salinity conditions and biomass will be greater under stratified, nutrient-rich conditions. The result would then be dominance of the tolerant microalgal groups in the estuary and decline in cells numbers. This was the case in the Bushmans Estuary with the flagellates prevailing over other phytoplankton groups due to their tolerance to oligotrophic conditions.

Under reference conditions the estuary would have horizontal and vertical gradients, which would have promoted phytoplankton species richness and abundance. Now, with the low abundance the filter-feeding invertebrates are expected to have reduced numbers in the estuary. However, a greater distribution of the Z. capensis might have increased species diversity and biomass in the invertebrate group due to increased submerged macrophyte habitat availability.

Thus, the impact of the freshwater input reductions to the estuary has been mostly experienced by the water column biota due to the lack of nutrient supply for productivity and salinity-diverse environments for greater community composition. On the contrary, abundance and productivity has increased in the intertidal area, and possibly increased biodiversity as well. This can be mostly attributed to the lack of freshwater input and flooding, which created stable marine conditions for Z. capensis to colonise the intertidal banks of the estuary. Large beds now extend into the upper reaches of the estuary. Increased marine sediment input and lack of flushing by floods also created the desired environment for the salt marsh. Nutrient-rich freshwater seepage along the lateral margins of the estuary created favourable conditions for the fresh- to brackish reed and sedge communities along the saline banks of the estuary. These areas provide favourable habitat for macroinvertebrates, fish and birds.

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4.4.2 Monitoring indicators for the Bushmans Estuary

A successful monitoring protocol is one that is based on a minimum number of easy to implement indicators that can adequately reflect core issues (Adams and McGwynne, 2004). Thus, based on the current state of the estuary, information from historical data and reports on monitoring indicators for estuaries (e.g. Adams and McGwynne, 2004), monitoring indicators that will best represent the health of the Bushmans Estuary were selected. These are magnitude of freshwater flow into the estuary, salinity gradients, sedimentation in the lower reaches and changes in bathymetry, benthic diatom species composition and macrophyte composition and cover.

These attributes require monitoring because the estuary has been in a freshwater-deprived state for a long time, which has led to a series of changes. Part of the problem has been marine sediment accumulation in the lower reaches of the estuary which the residents of the Ndlambe Municipality want to be removed by dredging. The monitoring indicators were carefully selected to highlight the problem areas of the estuary, in order for the Department of Water Affairs (DWA) to come up with a solution, if possible. The main drivers for the selection of the indicators were based on the sedimentation problem that has been identified by the Department of Water Affairs (DWA), the reduced freshwater input into the estuary from the river and pollution of the estuary from wastewater inputs. The frequency of the monitoring as advised by Adams and McGwynne (2004) has also been identified.

Freshwater inflow drives the physical dynamics of estuaries and is thus an important indicator to monitor. This is a limiting resource in the Bushmans Estuary and measures need to be taken to improve the situation. The only operating flow gauge in the Bushmans River catchment is 150 km away from the head of the estuary and is thus not representative of inputs into the estuary. A closer gauge (at the head of the estuary) is thus required and this is a responsibility of DWA. The installation of this gauge is necessary so that the frequency and duration of episodic events like floods and drought can be monitored. The data would also assist with management decisions about restrictions on abstractions. The estuary is in a dry area; with the MAR approximately 38 x 106 m3 y-1 and it is therefore sensitive to freshwater inflow reduction. Frequent flood events are important as they scour out accumulated sediment and maintain an open mouth in permanently open estuaries. This is especially important for the Bushmans Estuary as it is experiencing marine sedimentation and receives low freshwater inflow. This then brings up the next problem in the estuary that also requires monitoring; sedimentation. In order to be able to quantify the magnitude of sediment accumulation in the estuary, the changes in bathymetry and sedimentation in the mouth should be assessed every three years. Bathymetric surveys (e.g. using a D-GPS and echo- sounder) at pre-selected cross sections along the estuary and topographical surveys (using standard land surveying techniques) should be conducted at least every three years and after major floods (i.e. 1:10, 1:20 year floods). The measurement of bathymetric changes will be an important indicator of long-term marine sedimentation processes. Long-term

145 measurements will show the shifts between sediment deposition from the marine environment and inputs from the catchment.

As a measure of freshwater inflow and the prediction of the health of the phytoplankton in the estuary salinity distribution patterns would be a significant indicator. Salinity is a good indicator of freshwater intrusion and community composition (corresponding to vertical and horizontal gradients). The other water quality constituents that are generally important in estuary monitoring programmes are nutrients; dissolved oxygen; temperature; pH; turbidity and conductivity. The seasonal analysis of these constituents can help to detect perturbations when they arise and before they impact on ecological processes (Adams and McGwynne, 2004). Monitoring just the water column variables will not give effective results as was discovered in the current study. Measurements of the water column variables did not show pollution threats in the estuary while, for a long time, there had been sewage seepage. The expected outcome would have been high nutrient concentrations and phytoplankton blooms in the water column. However these impacts were recorded in the benthic diatom species composition and the macrophyte community. Estuary benthic diatom identification should be included in the monitoring plan to indicate sites of sewage and other inputs. This is a good tool because benthic diatom species composition can reveal the quality of the aquatic environment based on the tolerances or preferences of the species present.

Macrophyte distribution and composition is also an important parameter to include, especially since the productivity of the Bushmans Estuary is driven by macrophytes more than phytoplankton as it is a marine dominant estuary. Monitoring the extent and presence of plant communities is important because these are important habitats for other organisms and they also form the base of the food chain. Pockets of reeds and sedges at certain places in the saline lower reaches were indicators of freshwater seepage. Evaluations can be done every four years through the mapping of the different communities and measurement of changes in their cover and distribution over time. Aerial photographs of estuaries are available from the Department of Surveys and Mapping, Cape Town every four years. According to Adams and McGwynne (2004), at least two mud and two sandbanks should be chosen to serve as reference sites to monitor change in sediment distribution patterns that affect the availability of physical habitat space for benthic microalgae as well as emergent macrophytes.

4.5 Conclusion

A large number of important parameters that need to be incorporated into a monitoring plan were identified by Adams and McGwynne (2004), however to establish the health of the estuary inclusion of all these factors in a monitoring programme would be expensive. Therefore certain issues / indicators need to be identified for monitoring. In the Bushmans Estuary these indicators were chosen based on the prevailing factors of concern in the estuary namely reduced freshwater inflow, lack of salinity gradients, sedimentation and seepage of point and non-point source pollutants.

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APPENDICES

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Appendix A: List of diatom species found in Bushmans Estuary

Table A.1: Relative abundance of the dominant (relative abundance > 10%) benthic diatom species identified in the estuary

2006 2008 2009 Relative Relative Relative abundance abundance abundance Site Species Name (%) Species Name (%) Species Name (%) Hippodonta cf. gremainii Lange-Bertalot, Intertidal 1 Metzeltin & Witkowski 58 Amphora wisei (Salah) Simonsen 13 Tryblionella constricta (Kutzing) Ralfs 26 Amphora wisei (Salah) Simonsen 10 Hippodonta cf. gremainii Lange-Bertalot, Metzeltin & Witkowski 24 Navicula cf. uniseriata Hustedt 9 Subtidal 1 Stauroneis sp. Ehrenberg 6 Intertidal 2 Tryblionella constricta (Kutzing) Ralfs 35 Amphora acutiuscula Kutzing 15 Nitzschia fonticola Grunow 9 Stauroneis sp Ehrenberg 13 Nitzschia frustulum (Kutzing) Grunow 11 Nitzschia frustulum (Kutzing) Grunow 29 Hippodonta cf. gremainii Lange-Bertalot, Fragilaria sopotensis Witkowski & Lange- Metzeltin & Witkowski 8 Bertalot 10 Fragilaria elliptica Schumann 11 Nitzschia compressa (Bailey) Boyer var. Compressa 5 Subtidal 2 Navicula cf pennata Schmidt 32 Amphora acutiuscula Kutzing 17 Navicula salinicola Hustedt 19 Grammatophora oceanica Ehrenberg 8 Cocconeis scutellum Ehrenberg 12 Fallacia sp. Stickle & Mann 19 Delphineis surirelloides (Simonsen) Andrews 9 Tryblionella constricta (Kutzing) Ralfs 6 Intertidal 3 Tryblionella constricta (Kutzing) Ralfs 50 Opephora horstiana Witkowski 18 Seminavis sp. Cox 8 Diploneis bombus Cleve-Euler 13 Nitzschia frustulum (Kutzing) Grunow 14 Nitzschia fonticola Grunow 10 Seminavis sp. 10 Cox 7 Fragilaria elliptica Schumann 10 Subtidal 3 Diploneis smithii (Brebisson) Cleve var. smithii 18 Amphora acutiuscula Kutzing 25 Amphora acutiuscula Kutzing 21 Planothidium delicatulum (Kutzing) Diploneis bombus Cleve-Euler 12 Navicula gregaria Donkin 12 Round & Bukhtiyarova 16 Planothidium delicatulum (Kutzing) Round & Tryblionella constricta (Kutzing) Ralfs 9 Bukhtiyarova 10 Amphora sp. Ehrenberg 9 Planothidium sp. Round & Bukhtiyarova 9

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Cocconeis placentula v euglypta (Ehrenberg) Grunow 6 Navicula cf. cryptotenelloides Lange- Intertidal 4 Navicula sp. Bory 39 Nitzschia closterium (Ehrenberg) Smith 30 Bertalot 15 Seminavis sp.11 Cox 34 Nitzschia frustulum (Kutzing) Grunow 9 Cocconeis placentula v euglypta(Ehrenberg) Planothidium delicatulum (Kutzing) Round & Subtidal 4 Grunow 60 Bukhtiyarova 22 Navicula ammophila Grunow 17 Grammatophora oceanica Ehrenberg 9 Amphora acutiuscula Kutzing 20 Navicula gregaria Donkin 14 Tryblionella constricta (Kutzing) Ralfs 6 Navicula gregaria Donkin 10 Amphora coffeaeformis (Agardh) Kutzing 10 Astartiella punctifera (Hustedt) Witkowski & Lange-Bertalot 10 Intertidal 5 Navicula sp. Bory 75 Amphora subacutiuscula Schoeman 26 Nitzschia flexa Schumann 15 Nitzschia coarctata Grunow 26 Mastogloia exigua Lewis 16 Planothidium delicatulum (Kutzing) Round & Subtidal 5 Diploneis smithii (Brebisson) Cleve var. smithii 17 Bukhtiyarova 26 Amphora acutiuscula Kutzing 16.21622 Navicula ramosissima (Agardh) Navicula sp Bory 11 Amphora acutiuscula Kutzing 20 Grunow 13.51351 Tryblionella cf.littoralis Grunow 10 Ardissonea crystalline (Agardh) Gronow 6 Intertidal 6 Mastogloia exigua Lewis 17 Navicula tenelloides Hustedt 50 Planothidium delicatulum (Kutzing) Round & Subtidal 6 Bukhtiyarova 18 Diploneis elliptica (Kutzing) Cleve 33.33333 Amphora acutiuscula Kutzing 10 Navicula gregaria Donkin 10

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Table A.2: Relative abundance of the identified epiphytic diatom species in the estuary

Relative abundance (%) Species name 2006 2008 2009 Achnanthes brevipes Agardh 0.3 Achnanthes sp. Bory 0.3 Amphora acutiuscula Kutzing 0.8 Amphora coffeaformis (agardh) Kutzing 0.3 Amphora jostesorum Witowski, Metzeltin & Lange-Bertalot 4.3 Amphora sp. Ehrenberg 1.6 1.0 Berkeleya fennica Juhlin-Dannfelt 4.8 Berkeleya sp. Greville 0.2 Campylodiscus sp. Ehrenberg 1.2 Cocconeis engelbrechtii Cholnoky 6.2 Cocconeis placentula v euglyphyta (Ehrenberg) Grunow 27.8 12.9 Cocconeis scutellum v scutellum Ehrenberg 3.8 7 Cocconeis sp. 1 Ehrenberg 1.1 Cocconeis sp. 2 Ehrenberg 1.9 Cocconeis sp. Ehrenberg 9.2 Coscinodiscus sp. Ehrenberg 0.2 Cylindrotheca closterium (Ehrenberg) Reimann & Lewin 0.3 Delphineis sp. Andrews 0.2 Denticula sp. Kutzing 0.2 Diploneis sp. 1 Ehrenberg 0.3 Diploneis sp. 2 Ehrenberg 0.2 Fallacia sp. 1 Stickle & Mann 12.2 Fallacia sp. 2 Stickle & Mann 0.5 Fragilaria investiens (Smith) Clevel-Euler 11.8 6 Fragillaria sp. Lyngbye 4 0.8 Fragillaria sp.3 Lyngbye 1.6 Gomphonema sp. 1 Ehrenberg 0.2 Gomphonema sp. 2 Ehrenberg 0.2 Grammataphora sp. 1 Ehrenberg 1.0 Grammataphora sp. 2 Ehrenberg 0.3 Grammataphora sp. 3 Ehrenberg 3.2 Licmophora gracilis v gracilis (Ehrenberg) Grunow 4.9 Licmophora sp. Agardh 2.8 0.2 Masotgloia sp. Thwaites ex. Smith 0.6 Mastogloia elliptica (Agardh) Cleve 1.6 Mastogloia fimbriata (Brightwell) Grunow 0.3 Navicula cincta (Ehrenberg) Ralfs 0.5 Navicula durrenbergiana Hustedt 1.4 Navicula paeninsulae Cholnoky 6.2 Navicula ramosissima (Agardh) Grunow 11.1 27.6 Navicula sp. 1 Bory 15.7 Navicula sp. 2 Bory 4.1 Navicula sp. 3 Bory 1.7 Navicula sp. 4 Bory 1.9 Navicula sp. 5 Bory 0.3 Navicula sp. 6 Bory 1.7 Navicula sp. 7 Bory 0.8 Navicula sp. 8 Bory 1.3 Navicula sp. 9 Bory 0.6 Navicula sp. 10 Bory 0.2

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Navicula sp. 11 Bory 1.0 Navicula sp. 12 Bory 1.3 Navicula sp. 13 Bory 0.2 Navicula sp. 14 Bory 0.2 Navicula tenelloides Hustedt 0.2 Nitzschia closterium (Ehrenberg) Smith 6.4 Nitzschia frustulum (Kutzing) Grunow 1.9 43.5 7.5 Nitzschia sp. Hassall 0.1 Nitzschia sp.1 Hassall 1.7 Nitzschia sp.2 Hassall 0.3 Odontella sp. Agardh 0.2 Planothidium delicatulum (Kutzing) Round & Bukhtiyarova 0.2 Pleurasigma sp. Smith 0.4 Rhopalodia sp. Muller 0.2 Seminavis sp. Cox 0.7 0.3 Surirella sp. Turpin 0.2 Synedra sp. Ehrenberg 16.2 Tryblionella constricta (Kutzing) Ralfs 4.5 Tryblionella sp. Smith 0.2

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Table A.3: List of the abbreviated epiphytic and benthic diatom species, used in the ordination analysis, and their full names

Abbreviated species names Full name Amp_acut Amphora acuttuiscula Amp_coff Amphora coffaeformis Amp_jost Amphora jostesorum Amp_suba Amphora subacuttuiscula Amp_wise Amphora wisei Amph_sp. Amphora species Ard_crys Ardissonia crystalline Ast_punc Astartiella punctifera Ber_fenn Berkeleya fennica Camp_sp. Campylodiscus species Coc_enge Cocconeis engelbrechtii Coc_plac Cocconeis placentuala v euglyphyta Coc_scut Cocconeis scutellum v scutellum Cocc_sp. Cocconeis species Del_suri Delphineis surirelloieds Dip_bomb Diploneis bombus Dip_elli Diploneis elliptica Dip_smit Diploneis smithii v smithii Fall_sp. Fallacia species Fra_elli Fragillaria elliptica Fra_inve Fragillaria investiens Fra_sopo Fragillaria sopolensis Frag_sp. Fragillaria species Gram_ocea Grammatophora oceanica Gram_sp Grammatophora species Hip_grem Hippodonta cf. gremainii Lic_grac Licmophora gracilis v gracilis Licm_sp. Licmophora species Mas_exig Mastogloia exigua Nav_ammo Navicula ammophila Nav_cryp Navicula cf. cryptotenelloides Nav_greg Navicula gregaria Nav_paen Navicula paeninsulae Nav_penn Navicula cf. pennata Nav_ramo Navicula ramosissima Nav_sali Navicula salinicola Nav_tene Navicula tenelloides Navi_sp Navicula species Nit_clos Nitzschia closterium Nit_coar Nitzschia coarctata Nit_comp Nitzschia compressa v compressa Nit_flex Nitzschia flexa Nit_font Nitzschia fonticola 170

Nit_frus Nitzschia frustulum Ope_hors Opephora horstiana Pla_deli Planothidium delicatum Plan_sp. Planothidium species Semi_sp. Seminavis species Stau_sp. Stauroneis species Syne_sp. Synedra species Try_cons Tryblionella constricta Try_litt Tryblionella cf. littoralis

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Appendix B: Images of the identified diatom species in the Bushmans Estuary

Nitzschia frustulum (Kutzing) Grunow

Fragillaria elliptica Schumann

172

Nitzschia fonticola Grunow

Seminavis sp. Cox

173

Navicula cf. cryptonelloides Lange-Bertalot

Mastogloia exigua Lewis

174

Nitzschia flexa Schumann

Navicula tenelloides Hustedt

175

Fallacia sp. Stickle & Mann

Navicula salinicola Hustedt

176

Amphora coffeaeformis (Agardh) Kutzing

Planothidium delicatulum (Kutzing) Round & Bukhtiyarova

177

Planothidium delicatulum (Kutzing) Round & Bukhtiyarova

Navicula ammophila Grunow

178

Navicula gregaria Donkin

Cocconeis placentula v. euglypta (Ehrenberg) Grunow

179

Nitzschia palea (Kutzing) W. Smith

Amphora micrometra Giffen

180

Berkeleya rutilans (Trentepohl)

Cocconeis placentula v. lineate (Ehrenberg) Van Heurck

181

Mastogloia gleskesii Cholnoky

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Appendix C: Full names of the salt marsh species identified in the Bushmans Estuary

Table C.1: Full salt mash species names for the abbreviations in Figure 3.18

Abreviation Full Name

Sarc Sarcocornia species Bass Bassia diffusa Spar Spartina maritima Disp Disphyma crassifolium

Spor Sporobolus virginicus Sten Stenotaphrum secundatum Limo Limonium linifolium

Tetr Tetragonia decumbens Trig Triglochin species Bulb Bulboschoenus maritimus

Junc Juncus species Phra Phragmites australis Atri Atriplex species

Sene Senecio species

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