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Environmental Management of
Aquaculture Effluent:
Development of
Biological Indicators and Biological
Filters
Adrian B. Jones
Environmental Management of Aquaculture Effluent:
Development of Biological Indicators and Biological Filters
A Thesis
submitted by
Adrian B. Jones B.Sc. (Hons)
The University of Queensland, Australia
to the
Department of Botany
The University of Queensland
AUSTRALIA
in fulfilment of the requirements for
the degree of Doctor of Philosophy within
The University of Queensland
July 1999
STATEMENT
The work presented in this thesis is, to the best of my knowledge and belief, original, except
as acknowledged in the text, and the material has not been submitted, either in whole or in
part, for a degree at this or any other University.
Signed................................................
ACKNOWLEDGMENTS
This thesis was initiated from a Fisheries Research Development Corporation (FRDC) grant
to Dr Nigel Preston (CSIRO) and Moreton Bay Prawn Farm. Additional funding was
provided through grants from the Australian Research Council (ARC) to Dr William C.
Dennison, the CRC for Aquaculture and a University of Queensland Postgraduate Research
Scholarship.
Its hard to believe its finally finished. From the early days at CSIRO and Moreton Bay
Prawn Farm laying concrete slabs and besser brick raceways to cruising Moreton Bay in
thunder storms, the all-nighters at Straddie and then the endless days and nights in front of
the computer. None of it would have been possible without the support and friendship of
many people, mostly from the Marine Botany Group at UQ.
Dr. Bill Dennison, who developed my interest in marine research through his amazing
enthusiasm in the field, for his advice and suggestions, and for his ability to make you realise
that its not all as bad as it seems when you’re in the depths of confusion.
To everyone else in Marine Botany who provided help with field work, help with
interpretation and presentation of results, proof reading of manuscripts, and general
friendship and support. In particular, thanks to Cindy Heil, Michele Burford, Mark
O’Donohue and Joelle Prange who reviewed various sections of the thesis. Also a special
mention to Ros Murrell who has tirelessly helped me to track down a certain person when
times were desperate. You were my link to the outside world during those last six months.
ACKNOWLEDGMENTS viii
Dr. Nigel Preston, CSIRO Cleveland Marine Laboratories, for his constant “encouragement”
regarding my writing abilities, his invaluable help regarding the initial planning of the thesis
topic, his timely pep talks and last minute editing.
Theresa Mitchell, for her help, efficiency, support and advice regarding anything and
everything to do with forms, policies, scholarships and finally the thesis submission
procedures. Life at Botany just isn’t the same without you! Jan Stewart, for conducting the
isotopic analyses, and Gordon Moss for analysing the amino acid samples.
Sabine Roberts who read early drafts of all the chapters, picking my grammar to pieces and
helping get my thoughts on the right track. Thankyou so much for all your help,
encouragement and friendship throughout my PhD.
My parents who put up with me over the last 4 years constantly complaining about how this
didn’t work and that didn’t work, and “no, I don’t know when I will finish it all.” Thankyou
for being there to listen to my complaining, and for your support and most importantly, for
never pressuring me.
Finally to Tracey, who despite my constant rebuttal continued to maintain that somehow I
would finish it before our flight left. Thankyou for your support during my constant state of
stress and panic, especially when I was having serious doubts as to how I would manage to
finish it on time. Thankyou for putting up with no end of complaints and irrational
ACKNOWLEDGMENTS ix
behaviour, and for always being there and for accepting my stream of unfulfilled promises.
Thankyou for your marathon proof reading / printing / collation efforts at the end, especially
amongst the continual printer jams and photocopying nightmares and my accompanying
frustration and rampage. Thankyou for somehow managing to keep me together! Most
importantly, thankyou for your love, companionship, support and understanding.
ABSTRACT
Rapid global expansion of the aquaculture industry has prompted the need for development of techniques for
effective environmental management. In intensively farmed regions, aquaculture effluent has resulted in
environmental degradation of receiving waters. The issues to be addressed include analysis of effluent water
quality, determination of the ecological impact of effluent on the ecosystem, and development of remediation
strategies to reduce these impacts. Physical and chemical water quality analyses can identify elevated
concentrations of suspended solids, chlorophyll a, water column nutrients and other components of aquaculture
effluent, however, additional biological sampling is required to provide meaningful information about the
ecological impacts of effluent discharge on receiving waters.
Analyses of the amino acid composition, tissue nitrogen content and stable isotope ratio of nitrogen (δ15N) in
seagrasses, mangroves and macroalgae were developed as biological indicators to determine the influence of
shrimp farm effluent on a coastal ecosystem. Different responses in these biological parameters revealed that
the impacts of aquaculture effluent on receiving waters were qualitatively different to the impacts of sewage
effluent. The impacts were also spatially more extensive than identified by water quality analyses, which
revealed no elevation in the concentration of water column nutrients, chlorophyll a concentration or total
suspended solids further than 400 m from the mouths of the creeks receiving the sewage and aquaculture
effluent. The maximum δ15N of the mangroves, seagrass and macroalgae associated with the treated sewage
discharge was 19.6‰, which was significantly higher than the influence of the shrimp effluent (7.6‰). A δ15N
value of 4.5‰, which is elevated relative to unimpacted sites, indicated that the impacts extended up to 4 km
from the mouths of the creeks. Differences in the concentrations of the amino acids proline, serine, glutamine
and alanine in the seagrass and macroalgae were suggested to reflect the source (aquaculture or sewage) of the
nutrients taken up by the plants.
To reduce the environmental impacts, effluent treatment techniques using biological filters were investigated.
Filtration by oysters (Saccostrea commercialis) significantly reduced the concentrations of chlorophyll a
(phytoplankton), bacteria, total nitrogen, total phosphorus and total suspended solids to 5%, 32%, 67%, 63%
and 11% of the initial concentrations, respectively. However, oyster excretion increased the concentrations of
the dissolved nutrients, ammonium (from 18 to 51 µM), nitrate / nitrite (from 1.0 to 13 µM), and phosphate
(from 0.5 to 3.3 µM), however macroalgal (Gracilaria edulis) absorption significantly reduced these
concentrations to 2.3%, 2.2% and 4.8%, respectively. The ratio of ammonium to nitrate / nitrite in the effluent
was also significantly reduced, which has positive implications for recycling of wastewater back into shrimp
production ponds, and reducing impacts on receiving waters.
The efficiency and condition of the oysters and macroalgae was reduced by fouling from the high concentration
of suspended particulates in the effluent. Several novel techniques such as dissolved free amino acid
composition, pigment concentrations, PAM fluorescence, tissue nitrogen and δ15N were used to assess the
condition of the macroalgae. It was observed that an intermediate reduction in the concentration of suspended
ABSTRACT xii
particulates resulted in the best growth and condition of the biofilters. The concentration of particulates in this
treatment (11 nephelometric turbidity units) provided sufficient particulates for oysters to filter, and a source for
regeneration of nutrients for macroalgal uptake, as well as reducing the effects of photoinhibition which can
occur in Gracilaria spp. at relatively low light intensities.
The problems associated with fouling were successfully mitigated by incorporating natural sedimentation prior
to oyster filtration, and subsequent macroalgal absorption. This combined system of treatment proved effective
at optimising the performance of the biological filters to improve the water quality of the effluent. Using this
combination of polyculture, it was estimated that up to 18 kg N ha-1 d-1 and 15 kg P ha-1 d-1 could be removed
from commercial shrimp ponds.
The water quality of aquaculture effluent and its impact on the receiving waters will vary due with differing
environmental conditions, as well as the type of aquaculture being conducted. Regardless, this thesis has
demonstrated that filtration / absorption by various marine organisms can be effective tools for monitoring and
reducing the environmental impacts of aquaculture effluent.
TABLE OF CONTENTS
Statement v
Acknowledgments vii
Abstract xi
Table of Contents xiii
List of Tables xviii
List of Figures xx
List of Plates xxiii
CHAPTER 1. INTRODUCTION 1
1.1 Aquaculture 1
1.2 Environmental Impacts 2
1.3 Biological Indicators 4
1.4 Biological Treatment Options 5
1.4.1 Oysters 6
1.4.2 Macroalgae 9
1.5 Polyculture / Integrated Aquaculture 12
1.6 Thesis Aims 13
1.7 Thesis Overview 14
1.7.1 Chapter Outline 16
1.9 Publication Status of Thesis Chapters 18
CHAPTER 2 ASSESSING ECOLOGICAL IMPACTS OF SHRIMP
AND SEWAGE EFFLUENT: BIOLOGICAL INDICATORS WITH
STANDARD WATER QUALITY ANALYSES 19
Abstract 19
2.1 Introduction 20
2.2 Materials and Methods 25
TABLE OF CONTENTS xiv
2.2.1 Study Region 25
2.2.2 Experimental Design 26
2.2.3 Collection 27
2.2.4 Analytical Procedures 27
2.2.5 Statistical Analysis 30
2.3 Results 31
2.3.1 Physical and Chemical Water Quality Analyses 31
2.3.1.1 Salinity 31
2.3.1.2 Nutrients 31
2.3.1.3 Phytoplankton 32
2.3.1.4 Suspended Solids and Secchi Depth 34
2.3.1.5 Sediment Organic Content 34
2.3.2 Bioindicators 35
2.3.2.1 Tissue Nitrogen Content 35
2.3.2.2 δ15N Stable Isotope Ratio of Nitrogen 36
2.3.2.3 Free Amino Acid Composition 40
2.4 Discussion 45
2.4.1Water Quality Parameters 45
2.4.1.1 Effluent Composition 45
2.4.1.2 Phytoplankton Biomass and Productivity 46
2.4.2 Biological Indicators 47
2.4.2.1 Tissue N Content 47
2.4.2.2 δ 15N Isotopic Signature 48
2.4.2.3 Amino Acid Composition 51
2.4.3 Comparison of Impacts 55
2.4.4 Conclusion 57
2.4.5 Application for other types of Aquaculture 58
2.4.6 Remediation Options 59
TABLE OF CONTENTS xv
CHAPTER 3 OYSTER FILTRATION OF SHRIMP FARM EFFLUENT,
THE EFFECTS ON WATER QUALITY 63
Abstract 63
3.1 Introduction 64
3.2 Materials and Methods 67
3.2.1 Experimental Design 67
3.2.2 Analytical Procedures 68
3.3 Results 70
3.3.1 Suspended Solids 70
3.3.2 Organic content 70
3.3.3 Chlorophyll a 72
3.3.4 Bacteria 72
3.3.5 Total Nutrients 72
3.4 Discussion 73
3.4.1 Scaling Up Calculations 75
3.4.2 Summary 75
CHAPTER 4 THE EFFICIENCY AND CONDITION OF OYSTERS AND
MACROALGAE USED AS BIOLOGICAL FILTERS OF SHRIMP POND
EFFLUENT 77
Abstract 77
4.1 Introduction 78
4.2 Materials and Methods 81
4.2.1 Experimental Design 81
4.2.1.1 Filtration Efficiency Experiments 81
4.2.1.2 Biofilter Condition Experiments 83
4.2.2 Analytical Procedures 86
4.3 Results 90
4.3.1 Filtration Efficiency Experiments 90
4.3.1.1 Continual Flow 90
4.3.1.2 Recirculating Experiments 92
TABLE OF CONTENTS xvi
4.3.2 Biofilter Condition Experiments 97
4.4 Discussion 103
4.4.1 Efficiency of Biofilters 103
4.4.2 Condition of Biofilter Organisms 110
4.4.3 Conclusion 115
CHAPTER 5 IMPROVEMENTS IN WATER QUALITY OF AQUACULTURE
EFFLUENT AFTER TREATMENT BY SEDIMENTATION, OYSTER
FILTRATION AND MACROALGAL ABSORPTION 117
Abstract 117
5.1 Introduction 118
5.2 Materials and Methods 121
5.2.1 Experimental Design 121
5.2.2 Analytical Procedures 123
5.3 Results 126
5.3.1 Suspended Solids 126
5.3.2 Organic content 126
5.3.3 Chlorophyll a 128
5.3.4 Bacteria 129
5.3.5 Dissolved Oxygen 131
5.3.6 Total Nitrogen 132
5.3.7 Total Phosphorus 133
5.3.8 Ammonium 134
5.3.9 Nitrate / Nitrite 136
5.3.10 Phosphate 137
5.3.11 Nutrient Uptake Rates and Ratios 137
5.4 Discussion 140
5.4.1 Sedimentation 140
5.4.2 Oyster Filtration 141
5.4.3 Macroalgal Absorption 144
5.4.4 Nutrient Regeneration 147
5.4.5 Conclusions 149
TABLE OF CONTENTS xvii
CHAPTER 6 CONCLUSION 151
6. 1 Downstream Impacts 151
6.2 Efficiency of Biological Filters 152
6.3 Condition of Biofilters 153
6.4 Scaling up for Commercial Treatment 155
6.5 Other Potential Biofilters 156
6.6 Management Implications and Potential Problems with
Biofiltration / Polyculture 157
6.7 Benefits of Polyculture or Integrated Aquaculture 158
6.8 Future Research 159
6.9 Summary 160
BIBLIOGRAPHY 163
APPENDIX 1 FACTORS LIMITING PHYTOPLANKTON BIOMASS IN THE
BRISBANE RIVER AND MORETON BAY 192
APPENDIX 2 PHOTOSYNTHETIC CAPACITY IN CORAL REEF SYSTEMS:
INVESTIGATIONS INTO ECOLOGICAL APPLICATIONS FOR THE
UNDERWATER PAM FLUOROMETER 203
LIST OF TABLES
Table 2.1. Results of traditional water quality monitoring for the creek with shrimp farm
effluent and sewage treatment effluent. DIN = Dissolved Inorganic Nitrogen; DIP =
Dissolved Inorganic Phosphorus; Chl a = Chlorophyll a; Phyto Prod = phytoplankton
productivity; TSS = total suspended solids; VSS = volatile suspended solids; Secchi =
secchi disc depth. Only one replicate measurement was recorded for Secchi disk depth
and salinity (Practical Salinity Scale). 33
Table 2.2. Correlations (r2) between the concentration of phytoplankton (chlorophyll a) and
phytoplankton productivity (14C uptake) and various water quality parameters. DIN =
Dissolved Inorganic Nitrogen; DIP = Dissolved Inorganic Phosphorus; Phyto Prod =
phytoplankton productivity (mg C m-3 h-1); Chl a = Chlorophyll a (µg L-1); TSS = total
suspended solids (mg L-1); VSS = volatile suspended solids (mg L-1); ISS = inorganic
suspended solids (mg L-1); Secchi = secchi disc depth (m). Numbers in bold type
indicate significant correlations (r2 ≥ 0.6). 35
Table 2.3. Results of bioindicator monitoring for the creek with shrimp farm effluent and
sewage treatment effluent. δ15N = Nitrogen stable isotope ratio; %N = Tissue N content;
nd = no data (no plants were present). 38
Table 2.4. Results of bioindicator monitoring for the shrimp and sewage creeks. % refers to
percentage of total free amino acid pool. SER = serine; ALA = alanine; GLN =
glutamine; PRO = proline; Total αα = total concentration of free amino acids
(µmol g wet-1); nd = no data (no plants were present). 42
Table 3.1 Combinations of live and dead oysters (Saccostrea commercialis) used in
experiments to determine the effects of oyster density on the water quality of shrimp
pond effluent. 68
Table 3.2 Concentration of various water quality parameters before and after filtration by
oysters at 3 different densities (see Table 3.1). Values for control (no oysters) and shells
(dead shells only) are also given. Values in brackets are concentrations expressed as a
percentage of the inflow value. Values in italics are standard errors. 71
Table 4.1 Water quality parameters after filtration by oysters under flow through conditions
in raceways. Total N = total Kjeldahl nitrogen; Total P = total phosphorus. 92
LIST OF TABLES xix
Table 4.2 Water quality parameters after filtration by oysters after the first circuit during
recirculating flow in raceways. TSS = total suspended solids; Organic = organic
component of TSS (loss on ignition); Inorganic = inorganic component of TSS. 94
Table 4.3 Changes in the free amino concentration and composition of macroalgae for
various treatments in laboratory settling experiments. % refers to percentage of total
free amino acid pool. CIT = citrulline; GLU = glutamate; ALA = alanine; GLN =
glutamine; PHE = phenylalanine; SER = serine; Total αα = total concentration of free
amino acids (µmol g wet-1). 101
Table 5.1 Percentage of original concentrations of various water quality parameters after
settling, filtration by oysters and filtration by macroalgae. * p ≤ 0.05; ** p ≤ 0.01; *** p ≤ 0.001. Percentage of highest concentration represents the final concentration as a
percentage of the highest recorded concentration after sedimentation and oyster
filtration. The percent of initial concentration represents the final concentration as a
percentage of the initial concentration in the untreated effluent. The only differences
between the two values are for the dissolved inorganic nutrients (NH4+, NO3
-, & PO43-).
128
Table 5.2 Nutrient uptake and release rates for sedimentation, oyster filtration and
macroalgal absorption. Negative symbols represent nutrient uptake, and positive
represent nutrient release. The top value for each treatment is the gross value, the
middle value is the control and the bottom value (in bold type) is the net value after
correction for nutrient changes in the control tanks. The last row of results represent the
rates of macroalgal nutrient uptake over the first hour, when nutrient concentrations
were still saturating uptake kinetics. 136
LIST OF FIGURES
Figure 2.1 Map of study sites in Moreton Bay, including the location of shrimp and sewage
effluent discharges. 27
Figure 2.2. Map showing the values of %N in seagrass (Zostera capricorni), macroalgae
(Catenella nipae), and mangroves (Avicennia marina) at the study sites (see Fig. 2.1 for
site references). 39
Figure 2.3. Map showing the values of δ15N in seagrass (Zostera capricorni), macroalgae
(Catenella nipae), and mangroves (Avicennia marina) at the study sites (see Fig. 2.1 for
site references). 40
Figure 2.4. Map showing the amino acid composition of seagrass (Zostera capricorni) at the
study sites (see Fig. 2.1 for site references). 43
Figure 2.5. Map showing the amino acid composition of macroalgae (Catenella nipae) at the
study sites. Pie graphs have been reduced to quarters for layout purposes. The
remaining three quarters of the pie graphs not represented are a continuation of the
“other” amino acid category (not serine or alanine) (see Fig. 2.1 for site references). 44
Figure 2.6. Conceptual model of the two creeks and the range and type of impacts from the
different effluent sources. 61
Figure 3.1. Location map of Moreton Bay Prawn Farm near Brisbane, Australia. 66
Figure 3.2. Schematic representation of tank and waterflow layout. 67
Figure 4.1 Diagrammatic representation of experimental setup, a) single raceway with
baffles and oyster trays, and b) laboratory settling experiment. NTU = nephelometric
turbidity units. The oysters used in the experiments were Sydney Rock oysters,
Saccostrea commercialis and the macroalgae was Gracilaria edulis. Effluent was from
an intensive Penaeus japonicus shrimp farm. 84
Figure 4.2 Impacts of effluent on biofilters: a) Oyster mortality (%) from upper, middle and
lower trays after 2 weeks at low, medium and high oyster stocking densities in raceways
supplied with unsettled shrimp effluent, and b) change in dissolved nutrient
concentrations after passing effluent through low, medium and high macroalgal stocking
densities in raceways supplied with unsettled shrimp effluent. Positive change
represents an increase, negative change represents a decrease. 91
LIST OF FIGURES xxi
Figure 4.3 Particle size distribution, a) before and after control and oyster treatment
raceways during single continuous flow, b) before and after consecutive circuits through
oyster treatment raceways (linear scale), and c) before and after consecutive circuits
through oyster treatment raceways (log scale). 93
Figure 4.4 Concentrations of water quality components before and after consecutive circuits
through oyster treatment raceways, a) bacterial numbers, b) chlorophyll a concentration,
and c) total suspended solids (TSS). 96
Figure 4.5 Growth of oysters and macroalgae after 8 weeks in tanks supplied with shrimp
effluent pre-settled for 0, 1, 6 & 24 h. a) change in oyster growth rate expressed as
changes in oyster volume (cm3 oyster -1), and b) macroalgal biomass. n.d. = no data. 98
Figure 4.6 Response of macroalgae to 8 weeks in tanks supplied with shrimp effluent pre-
settled for 0, 1, 6 & 24 h. a) macroalgal growth expressed as number of news shoots per
tank, and b) concentration of the photosynthetic pigments, chlorophyll a (CHL) and
phycoerythrin (PE). 99
Figure 4.7 Macroalgal nitrogen content (a) and δ15N (b) after 8 weeks in tanks supplied with
shrimp effluent of different settlement times, a) %N, and b) δ15N. 100
Figure 4.8 The response of electron transport rate (ETR) versus photosynthetically active
radiation (PAR) in macroalgae incubated in seawater (control) or shrimp effluent
(settled 24 h plus oyster filtered for 12 h). 102
Figure 5.1 Design of integrated treatment system stocked with oysters (40 g Saccostrea
commercialis), and macroalgae (Gracilaria edulis). 124
Figure 5.2 Changes in total suspended solids (A) and phytoplankton biomass (chlorophyll a)
(B) from sedimentation, oyster filtration and macroalgal absorption. Standard error bars
have been plotted, but are too small to be visible. 127
Figure 5.3 Concentration of particles settled per litre from sedimentation and oyster
filtration. 129
Figure 5.4 Changes in the organic content of the a) total suspended solids (TSS) and b)
settled particles in the effluent water from sedimentation, oyster filtration and
macroalgal absorption. Standard error bars have been plotted, but are too small to be
visible. 130
Figure 5.5 Changes in bacterial numbers from sedimentation, oyster filtration and macroalgal
absorption. 131
LIST OF FIGURES xxii
Figure 5.6 Changes in water column dissolved oxygen concentrations from sedimentation,
oyster filtration and macroalgal absorption. Standard error bars have been plotted, but
are too small to be visible. 132
Figure 5.7 Changes in water column total N (A) and P (B) concentrations from
sedimentation, oyster filtration and macroalgal absorption. Standard error bars have
been plotted, but are too small to be visible. 133
Figure 5.8 Changes in water column NH4+, NO3
-, PO43- concentrations from sedimentation,
oyster filtration and macroalgal absorption. Standard error bars have been plotted, but
are too small to be visible. 135
Figure 5.9 Changes in water column total N: P ratio (A) and DIN: DIP ratio (B) from
sedimentation, oyster filtration and macroalgal absorption. Standard error bars have
been plotted, but are too small to be visible. 139
Figure 6.1 Diagrammatic design of water flow for typical untreated shrimp farms (left) and a
design to incorporate physical (sedimentation) and biological (oyster filtration and
macroalgal absorption) treatment (right). 162
LIST OF PLATES
Plate 1.1 Ponds at Moreton Bay Prawn Farm, an intensive shrimp farm (Penaeus japonicus)
near Moreton Bay, Queensland, Australia. 2
Plate 1.2 Shrimp Farm plume discharging into Moreton Bay, Queensland, Australia. 3
Plate 1.3 High phytoplankton concentration in plume from shrimp farm discharging into
Moreton Bay, Queensland, Australia. 3
Plate 1.4 Penaeus japonicus from ponds at Moreton Bay Prawn Farm, Queensland,
Australia. 5
Plate 1.5 Sydney Rock Oysters (Saccostrea commercialis) cultured in Moreton Bay,
Queensland, Australia. 7
Plate 1.6 Gracilaria edulis collected from Moreton Bay, Queensland, Australia. 10
Plate 4.1 Raceways constructed at Moreton Bay Prawn Farm, Queensland, Australia. 82
Plate 4.2 Control raceway on the left with no oysters and treatment stocked at “low” density
55 g oysters. Demonstrates changes in water clarity (reduction in suspended solids) with
the oyster tray clearly visible in the raceway stocked with oysters, but not in the control
raceway. 104
Plate 4.3 First chamber (foreground) and second chamber (background) of an oyster
treatment raceway showing the improvement in water clarity (reduction in suspended
solids) within the raceway. 105
Plate 4.4 Fouling of oysters by settling particulates in raceways. 108
Plate 5.1 Experimental setup with sedimentation drum (background) and control, oyster,
macroalgal filtration tanks (foreground). 124
Plate 5.2 Water samples collected: a) before sedimentation; b) after sedimentation;
and c) after biofiltration. 150
CHAPTER 1
INTRODUCTION
1.1 Aquaculture
The UN Food and Agricultural Organisation has estimated that by 2020 more than 50% of
fisheries production will need to come from aquaculture due to human population growth,
continuing demand for seafood, and static or declining natural fish harvests. However, in
many countries aquaculture practices have already resulted in the destruction of coastal
vegetation, salinisation of land, pollution of waterways and massive crop losses (Phillips et
al., 1993). Further expansion using current technologies is simply not justifiable or
sustainable. If the level of demand for seafood is to be met the only alternative is to develop
new technologies that require less space and have minimal adverse environmental impacts.
Penaeid prawn (shrimp) farming has been one of the most economically successful of all
intensive aquaculture industries. In the early days of shrimp farming and other forms of
aquaculture, the perception was that they were completely “clean” industries (Weston, 1991).
Recent reviews of intensive shrimp aquaculture have emphasised the need for more effective
controls on the quality of effluent water discharged into the environment (Phillips et al.,
1993; Primavera, 1994).
Shrimp farming can be separated into extensive, semi intensive and intensive culture systems
(Macintosh & Phillips, 1992). Extensive culture systems have large pond sizes (>5 ha),
relatively low stocking densities (<10 per m2), no aeration, and natural food sources (through
fertilisation). Intensive farming consists of smaller ponds (1 ha), very high stocking densities
(>20 per m2), aeration, and formulated high protein feed pellets (Plate 1.1). Intensive farming
CHAPTER 1 2
is becoming more prominent, increasing the potential for environment impact from shrimp
farming (Phillips et al., 1993).
Plate 1.1 Ponds at Moreton Bay Prawn Farm, an intensive shrimp farm (Penaeus japonicus) near Moreton Bay,
Queensland, Australia.
1.2 Environmental Impacts
Intensive shrimp aquaculture systems rely on high protein feed pellets to produce high rates
of growth, but a large proportion of the pellets are not assimilated by the shrimps (Primavera,
1994). Approximately 10% of the feed is dissolved and 15% remains uneaten. The
remaining 75% is ingested, but 50% is excreted as metabolic waste, producing large amounts
of gaseous, dissolved and particulate waste (Lin et al., 1993). Subsequently, the effluent
contains elevated concentrations of dissolved nutrients (primarily ammonia), plankton and
other suspended solids (Ziemann et al., 1992). The dissolved nutrients and organic material
in shrimp ponds stimulate rapid growth of bacteria, phytoplankton, and zooplankton (Lin et
al., 1993). These untreated wastes are usually discharged directly into the environment,
where they may enhance eutrophication, organic enrichment and turbidity of the local
waterways (Plates 1.2 & 1.3) (Eng et al., 1989; O' Connor et al., 1989; Prakash, 1989).
INTRODUCTION 3
Plate 1.2 Shrimp Farm plume discharging into Moreton Bay, Queensland, Australia.
Plate 1.3 High phytoplankton concentration in plume from shrimp farm discharging into Moreton Bay,
Queensland, Australia.
CHAPTER 1 4
Australia has a small but expanding coastal aquaculture industry. From 1984 to 1998, the
shrimp farming sector rose from 15 to 2 000 t. The industry is well placed to take advantage
of developments in integrated aquaculture systems, such as the use of natural biofilters and
recirculating systems. In southeast Queensland, there are two shrimp species farmed,
Penaeus monodon Fabricius and P. japonicus Bate (Plate 1.4), with stocking of post larvae in
October and harvest the following April to June. With increasing development of the
industry, concerns have risen about the impact of effluent from the farms. The effluent from
shrimp farms discharging into Moreton Bay, Queensland often has concentrations of
dissolved nitrogen and phosphorus which are 60 fold higher than receiving waters,
chlorophyl l a concentrations 200 fold higher, and total suspended solids (TSS) 20 fold higher
(Jones et al., 2001a; Chapter 2). Australian waters are relatively low in nutrients
compared with other coastal waters (State of the Environment Council, 1996), and therefore
impacts may be potentially more acute. In Moreton Bay, background concentrations of water
quality parameters are: NH4+ < 2 µM; NO3
- / NO2- ~ 0.1 µM; PO4
3- ~ 0.2 µM;
chlorophyll a < 1 µg L-1; TSS < 20 mg L-1.
1.3 Biological Indicators
Due to the close proximity of shrimp farm discharges to several other point and non-point
nutrient sources (ie. sewage effluent, agricultural runoff), it can be difficult to determine the
impacts of aquaculture on the environment (Grant et al., 1995). Techniques are needed to
distinguish the effect of each source and its range of impact so that appropriate discharge
limits can be applied.
INTRODUCTION 5
Plate 1.4 Penaeus japonicus from ponds at Moreton Bay Prawn Farm, Queensland, Australia.
Traditional water quality analyses provide little information as to the impact of nutrients on
the biota in the ecosystem (Lyngby, 1990). As a result there is a lack of data on the
ecological impact of aquaculture effluent (Gowen et al., 1990). The use of biological
indicators can provide information as to the nutrient source, the bioavailability of the
nutrients, and the integration of short lived nutrient pulses (Lyngby, 1990; Horrocks et al.,
1995; Jones et al., 1996; Udy & Dennison, 1997b; Jones et al., 1998; Appendix 1).
1.4 Biological Treatment Options
Concerns about the possible adverse impacts of aquaculture discharge have become a risk
factor for the industry (Braaten, 1991). This has prompted efforts to develop cost-effective
methods of effluent treatment. In addition to prohibitive costs, because of the large volume
of effluent, sewage treatment practices have proved inefficient due to the high suspended
solid load (Tetzlaff & Heidinger, 1990) and the high salinity of aquaculture effluent. There
are a number of commercially available bacterial systems to promote nitrification and
subsequent denitrification to remove nitrogen from the effluent. However, the effectiveness
of such systems for treating shrimp pond effluent have yet to be examined rigorously.
CHAPTER 1 6
The use of filter feeding bivalves such as oysters to consume phytoplankton, zooplankton,
and bacteria (Lin et al., 1993), and macroalgae to assimilate the remaining dissolved nutrients
(Haines, 1975) may prove to be an efficient and economically viable alternative for
improving the water quality of shrimp farm discharges (Hopkins et al., 1993a; Lin et al.,
1993). In addition to filtering organic food particles, oysters can also improve the quality of
pond effluent by reducing the concentration of inorganic suspended solids.
1.4.1 Oysters
The oyster industry in Moreton Bay, Queensland is based on the Sydney Rock Oyster
Saccostrea commercialis (Iredale & Roughley) (Plate 1.5), an estuarine species native to Port
Stevens, N.S.W., Australia and found from Victoria, Australia to Thailand (Angell, 1986).
This species is considered marketable between 29 and 40 g (bottle oysters) and 40 to 67 g
(plate oysters) whole weight (Witney et al., 1988).
Culture of oysters on traditional leases from spat to marketable size takes two to three years
(Witney et al., 1988). Local availability of oysters is seasonal, with oysters being fat and
ready for sale in summer, while in winter when phytoplankton concentrations are lower
(Dennison et al., 1993), they are lean and growth is substantially slower. The use of land
based aquaculture systems has been trialed to improve productivity and year-round
marketability, but most attempts have been relatively unsuccessful, primarily due to the high
cost and unreliability of mass algal culture. In an attempt to find alternative sources of
microalgae as food for enhanced oyster production, shrimp farm effluent has been trialed
(Wang & Jakob, 1991; Hopkins et al., 1993a; Jakob et al., 1993; Lin et al., 1993). Very fast
INTRODUCTION 7
rates of growth has been observed for oysters grown under controlled conditions with shrimp
pond effluent (Jakob et al., 1993).
Plate 1.5 Sydney Rock Oysters (Saccostrea commercialis) cultured in Moreton Bay, Queensland, Australia.
Oysters are suspension feeders and use their gills to filter phytoplankton, zooplankton,
bacteria and other microscopic particles. Bivalves can remove phytoplankton from the water
with high efficiency (Jørgensen, 1966), but their filtering ability is affected by a number of
factors including the water flow rate (Walne, 1972), temperature (Loosanoff & Tommers,
1948), salinity (Djangmah, 1979), reproductive effort, and silt concentration (Loosanoff &
Tommers, 1948; Angell, 1986). A temperature of approximately 30°C (Angell, 1986) and
salinity of 35‰ (Nell & Gibbs, 1986) is optimal, although the survival range for
S. commercialis is 15 – 50‰.
Shrimp pond effluent water typically has elevated concentrations of total suspended solids, a
large fraction being small inorganic clay minerals (Smith, 1996). In waters with high
concentrations of silt, oyster pumping and therefore feeding can be greatly inhibited
(Loosanoff & Tommers, 1948), or they may even cease pumping entirely.
CHAPTER 1 8
To overcome the problems of oyster fouling, sedimentation ponds could be used to remove
the larger settleable particles prior to oyster filtration (Wang, 1990). The remaining particles
are either motile, or are small particles (<5 µm) with a specific gravity similar to seawater
and therefore do not easily settle out of suspension (Rubel & Hager Inc., 1979). Oysters can
then be used to filter the organic food particles and convert them to meat and faeces. Oysters
can also combine small inorganic clay particles (which otherwise would not settle out) with
mucous into larger particles and egest them as pseudofaeces (Tenore & Dunstan, 1973),
which readily settle out of suspension.
Phytoplankton can be ingested, excreted as pseudofaeces, or simply settle out of suspension
without being available as food (Scura et al., 1979). High concentrations of plankton can
also inhibit feeding in bivalves (Ali, 1970; Schulte, 1975). At high food concentrations
oysters produce large amounts of pseudofaeces (ie. food filtered but not ingested) and
therefore have high deposition rates, demonstrating an inefficient use of filtered food (Tenore
& Dunstan, 1973).
Although oyster stocking density does not appear to affect survival, high densities may
decrease individual weight due to competition for food (Holliday et al., 1991). Given the
high concentrations of phytoplankton in shrimp ponds and pond effluent, it is more likely to
be the high concentrations of fine inorganic particles that control filtering rates.
A significant portion of the research to date regarding growing oysters in the effluent from
shrimp ponds has focussed on the advantages for growing the oysters. The highest ever
reported sustained growth rates for the American oyster (Crassostrea virginica), were when
oysters were supplied with effluent from a shrimp pond. Growth rates were reported to be
INTRODUCTION 9
0.5 g d-1, from seed size (0.04 g) to market size (55.0 g), in just 4 months (Jakob et al., 1993).
The authors state that “it was clearly shown that undiluted, semi-intensive, marine shrimp
pond water provides all the requirements for the very rapid growth of the American oyster
C. virginica from 0.05 g spat through 78 g adults” (Jakob et al., 1993). Evidence that the
quality of oysters remains high in aquaculture was observed by Lam & Wang (1989) who
used shrimp pond water to produce excellent quality half-shell oysters, grown from 0.1g to
54.2 g in 198 days with 96% survival.
The use of oysters as biofilters can improve the quality of water leaving aquaculture ponds,
and potentially provide a secondary cash crop. After filtration by oysters most of the
nutrients (those bound up in phytoplankton and other suspended solids), are deposited as
faeces and pseudofaeces, while the rest are incorporated into oyster tissue. However, oysters
can also contribute significant amounts of ammonia to the effluent through excretion (Srna &
Baggaley, 1976). Ammonia toxicity to shrimp is one of the primary reasons farmers
undertake water exchange (Kou & Chen, 1991), and therefore it must be removed before
effluent water can be recycled back into production ponds. Consequently, removal of these
deposited nutrients from the system entirely will require either physical removal of the settled
sediment, denitrification, or assimilation of the dissolved nutrients (from the remineralisation
of faeces and pseudofaeces) by macroalgae such as Gracilaria spp. (Funge-Smith & Briggs,
1998).
1.4.2 Macroalgae
Macroalgae can absorb significant quantities of dissolved inorganic and organic nutrients,
usually with a preference for NH4+ (D'Elia & DeBoer, 1978; Haines & Wheeler, 1978;
Hanisak & Harlin, 1978; Harlin, 1978). The ability of macroalgae to rapidly take up nutrients
CHAPTER 1 10
for growth, and store luxury reserves in the form of amino acids and pigments makes them
ideal for stripping nutrients from aquaculture effluent (Haines, 1975). Additionally,
macroalgae are known to absorb and store heavy metals (Burdin & Bird, 1994), which may
be a potential pollutant in shrimp pond effluent.
Removal of nutrients by macroalgae is also efficient as harvesting is relatively simple, and
provides an additional cash crop (Hopkins et al., 1995b). Macroalgae can also assimilate
metabolic wastes from mariculture animals, which is beneficial to shrimp production ponds if
the wastewater is to be recycled (Qian et al., 1996). Macroalgae can be used to ensure
complete removal of inorganic nitrogenous excreta from the bivalves (Mann & Ryther,
1977). In particular, commercial red seaweeds such as species from the genera Chondrus,
Gracilaria, Agardhiella and Hypnea are candidates as a final “polishing” step to leave the
effluent virtually free of inorganic nitrogen (Ryther et al., 1975) (Plate 1.6).
Plate 1.6 Gracilaria edulis collected from Moreton Bay, Queensland, Australia.
Certain species of red macroalgae (Rhodophyta), in particular those from the genera
Gracilaria, Gelidium, and Hypnea are harvested commercially. These species contain
INTRODUCTION 11
sulfated galactan agar and carrageenin which are widely used in the pharmaceutical, cosmetic
and food industries (Raven et al., 1987). Nutrients are generally the limiting factor to
macroalgal growth in natural systems, and attempts have been made to culture them in land
based aquaculture systems. The wastewater from aquaculture effluent contains sufficient
nutrients to sustain the high growth rates required without fertilisation, but the high
concentrations of suspended solids can foul the macroalgae and reduce light availability
(Briggs & Funge-Smith, 1993).
There are potentially considerable economic benefits to be gained from growing macroalgae
in shrimp pond effluent. The growth of Hypnea musciformis in the effluent from a tropical
mariculture system has been estimated as producing a gross harvest value of $107 250 ha-1
annually (Roels et al., 1976). H. musciformis cultured in aquaculture effluent grew at 64.5 g
wet wt d-1, compared to “deep water” growth of 12.1 g wet wt d-1 (Haines, 1975). Percent
carrageenin yields however were lower, ie., 16% dry wt versus 29% for the “deep water”,
however the total production of carrageenin is approximately 3 times greater from the
aquaculture effluent (Haines, 1975).
The use of bivalves and / or macroalgae to treat the effluent from shrimp farms has been
investigated in a number of studies (Wang & Jakob, 1991; Hopkins et al., 1993a; Jakob et al.,
1993; Shpigel et al., 1993b; Jones & Preston, 1999; Chapter 3). Using oysters (to filter
phytoplankton, bacteria and other suspended solids), and macroalgae (to take up dissolved
nutrients) can potentially improve the quality of shrimp pond effluent. In addition to the
environmental benefits for receiving waters, there are also economic gains resulting from the
conversion of high cost uneaten and dissolved feed pellets into two additional marketable
crops (Wang, 1990).
CHAPTER 1 12
1.5 Polyculture / Integrated Aquaculture
Polyculture is defined as the culture of several different organisms in the one culture unit. In
contrast, integrated aquaculture is the co-culture of different organisms, but in discrete culture
units (Chien & Tsai, 1985). These techniques are regarded as being more ecologically sound
methods of aquaculture (Mackay & Lodge, 1983), with a more efficient use of resources, and
a higher resilience against environmental fluctuation (Chien & Liao, 1995).
Despite the advantages of these types of combined aquaculture, there may be some problems
associated with management of several organisms all with differing culture requirements.
Management can be more complex with respect to stocking densities, culture techniques and
associated infrastructure, harvesting procedures, and effluent flow management (Chien &
Liao, 1995). Specific problems for intensive shrimp farming relate to fouling effects from the
high concentrations of suspended solids on secondary crop species (and potential biofilters)
such as oysters and macroalgae (Ziemann et al., 1992; Funge-Smith & Briggs, 1998).
Although the use of these and other biological treatment techniques for facilitating water
recycling are ecologically sound, much research is needed to improve the efficiency of these
systems (Lin, 1995).
Effective management of aquaculture effluent can be separated into identification of
downstream impacts, and effective farm management to reduce these impacts. Identification
of impacts to receiving waters may be accomplished with a combination of water and
sediment water analyses with biological indicators to elucidate ecological impacts. Effective
on-farm management of effluent can probably be accomplished by a combination of physical
and biological treatment techniques.
INTRODUCTION 13
1.6 Thesis Aims
• Characterise the components of shrimp pond effluent, and their concentrations relative to
the receiving waters,
• Develop the use of various marine plants as bioindicators to determine the effects of
prawn farm effluent on receiving waters,
• Determine the viability of oysters and macroalgae as biological treatment organisms for
shrimp pond effluent,
• Determine the differences in biological filter performance with changes in density, size,
and water flow regimes,
• Identify problems associated with maintaining oysters and macroalgae in the high
suspended solids environment and optimise techniques to minimise the impact on their
growth, condition, and effectiveness as biofilters,
• Design an integrated system to produce the greatest improvements in water quality, while
maintaining the condition of the biological filter organisms.
CHAPTER 1 14
1.7 Thesis Overview
Despite several reported cases of large scale environmental degradation linked to aquaculture
effluent, there has been no successful determination of the ecological impacts, and certainly
no techniques to distinguish these downstream impacts in relation to other nutrient inputs.
Techniques to improve effluent discharge water quality, including the use of bivalves to filter
aquaculture effluent have been undertaken on a small scale by the industry in other regions of
the world. However, there has been a distinct lack of quantitative data to determine the most
efficient use of these techniques and the ecophysiological responses of the biofilter
organisms. This thesis has addressed these shortcomings.
Bioindicator techniques were developed to investigate the ecological impacts of aquaculture
effluent and biofilter organisms were employed, not only to mitigate these impacts, but also
to provide an efficient use of resources by producing secondary crops from aquaculture farm
effluent. The research is this thesis has been conducted using techniques to look at
ecophysiological responses of organisms and ecological changes in the system. This
contrasts much of the published material in this area, which has been conducted purely at an
applied level.
Evaluation of ecological impacts from shrimp farming was conducted using biological
indicator techniques (tissue nitrogen content, δ15N isotopic signatures, amino acid
composition, phytoplankton productivity) in conjunction with more traditional water quality
parameters (nutrient concentrations, suspended solids, chlorophyll a) to determine
ecophysiological changes in the biota in the receiving waters. Rates of isotopic fractionation
of nitrogen in the effluent and the changes in the dissolved free amino acid composition of
the macroalgae incubated in shrimp effluent under controlled laboratory conditions provided
INTRODUCTION 15
some of the background responses used for determining the spatial range of impacts of
shrimp effluent in receiving waters. These biological or ecological health indicators provided
direct measures of the influence of aquaculture discharge.
The effects of different sizes and stocking densities of oysters and different densities of
macroalgae on the water quality (total and dissolved nutrients, chlorophyll a, bacteria, total
suspended solids, organic versus inorganic particulates, and particle size distribution) of
shrimp effluent was determined for a variety of effluent flow regimes and during different
stages of the shrimp growout season. In particular, analysis of the particle size distribution of
the effluent provided information into the mechanisms by which oysters remove particulates
from the water column, especially the small inorganic clay particles that are difficult to
remove by sedimentation or mechanical filtration.
The effects of different concentrations of suspended solids from shrimp effluent on oyster
and macroalgal condition was determined by physiological responses in the organisms. This
information facilitated estimates of the optimum concentration of suspended particulates for
efficient filtration performance by oysters and macroalgae, while minimising sedimentation
time and / or mechanical filtration costs. A variety of novel techniques such as dissolved free
amino acid composition, pigment concentrations, PAM fluorescence, tissue nitrogen and
δ15N were used to assess the condition of the macroalgae. These techniques also provide
information about the bioavailability of the nutrient profile from shrimp effluent, i.e. whether
it is suited for uptake by biofilter organisms (e.g. oysters and macroalgae).
Higher effluent flow rates are likely to improve biofilter condition, but may reduce the
filtration performance. In an attempt to improve the condition and performance of the
CHAPTER 1 16
biofilters, experiments were conducted to recirculate the effluent though the biofilter
organisms several times to test the possibility of increasing the effluent flow rate, without
sacrificing improvements in water quality.
The combined efficiency of sedimentation, followed by oyster filtration of particulates and
macroalgal absorption of dissolved nutrients proved to be an effective technique for
improving shrimp pond effluent water quality. After treatment in this polyculture system, the
effluent proved suitable for reuse in shrimp production ponds. The rates of nutrient
regeneration from settled particulates, oyster excretion rates, nutrient uptake rates (bacteria,
phytoplankton and macroalgae) and loss of N to the atmosphere via volatilisation and
denitrification were determined directly, or inferred by difference.
1.7.1 Chapter Outline
Chapter 2
Investigations were conducted to determine the impact of effluent from a local shrimp farm
on the biota and integrity of the receiving waters in Moreton Bay. Results were compared
with data from other unimpacted regions in Moreton Bay and with a nearby sewage treatment
plant. Several bioindicator techniques were utilised to characterise the impacts of the
effluent.
Chapter 3
Experiments were conducted to determine if oysters would be successful at improving the
water quality of shrimp pond effluent, and to assess the optimal stocking density of the
oysters to produce the greatest improvements in water quality.
INTRODUCTION 17
Chapter 4
The filtering efficiency of macroalgae and different sized oysters in raceways with flow
through effluent supply, and recirculating supply were conducted. Issues regarding fouling of
oysters and macroalgae were investigated to determine the maximum concentration of
suspended solids that the oysters and macroalgae could tolerate without adversely impacting
their health, survival and filtration efficiency.
Chapter 5
The overall efficiency of a polyculture treatment system was tested using sedimentation
followed by oyster filtration and macroalgal absorption.
Chapter 6
The conclusions of the study and areas of potential future research and comparisons with the
results of other studies are discussed.
CHAPTER 1 18
1.9 Publication Status of Thesis Chapters
Chapter 2
Jones, A.B., O'Donohue, M.J., Udy, J. & Dennison, W.C. (2001) Assessing ecological impacts of
shrimp and sewage effluent: Biological indicators with standard water quality
analyses. Estuarine, Coastal and Shelf Science 52, 91–109.
Presented at the Australian Marine Science Association annual conference,
Adelaide, Australia, July 1998.
Chapter 3
Jones, A.B. & N.P. Preston (1999) Oyster filtration of shrimp farm effluent, the effects on
water quality. Aquaculture Research 30, 51-57.
Chapter 4
Jones, A.B., N.P. Preston & W.C. Dennison (2002) The efficiency and condition of oysters
and macroalgae used as biological filters of shrimp pond effluent. Aquaculture
Research 33, 1-19.
Chapter 5
Jones, A.B., Dennison, W.C. & Preston, N.P. (2001) Integrated treatment of shrimp effluent
by sedimentation, oyster filtration and macroalgal absorption: a laboratory scale study.
Aquaculture 193 (1-2), 155-178.
Presented at the World Aquaculture Society Meeting, Sydney, Australia, May 1999.
CHAPTER 2
ASSESSING ECOLOGICAL IMPACTS OF SHRIMP AND SEWAGE EFFLUENT:
BIOLOGICAL INDICATORS WITH STANDARD WATER QUALITY ANALYSES
Abstract
Despite evidence linking shrimp farming to several cases of environmental degradation, there remains a lack of
ecologically meaningful information about the impacts of effluent on receiving waters. The aim of this study
was to determine the biological impact of shrimp farm effluent, and to compare and distinguish its impacts from
a nearby treated sewage discharge. Assessment of impacts was conducted using both water quality / sediment
analyses and biological indicators. Water quality and sediment parameters measured included chlorophyll a,
total suspended solids, volatile suspended solids, dissolved nutrients, salinity, and sediment organic content.
Biological indicator monitoring consisted of analysis of amino acid composition, tissue nitrogen (N) content and
stable isotope ratio of nitrogen (δ15N) in seagrasses, mangroves and macroalgae. The study area consisted of
two tidal creeks, one receiving effluent from a sewage treatment plant (sewage creek) and the other receiving
effluent from an intensive shrimp farm (shrimp creek). The creeks discharged into Moreton Bay, a sub tropical
coastal embayment on the east coast of Australia. Water quality in both creeks was significantly modified, but
changes were indistinguishable from unimpacted eastern Moreton Bay levels further than 750 m from the c reek
mouths. Biological indicators, however, detected significant impacts up to 4 km beyond the creek mouths. The
shrimp creek was more turbid due to clay minerals with a relatively high dissolved NH4+ (3.8 µM)
concentration, whereas the sewage creek had a higher percentage of organic material (35%) and dissolved
nutrient concentrations were higher, particularly NO3- / NO2
- (65 µM) and PO43- (31 µM). The sewage creek did
not support high phytoplankton productivity (18-20 mg C m-3 h-1), in spite of high nutrient concentrations.
Mangroves and macroalgae in the sewage creek were highly enriched with sewage nitrogen (indicated by high
δ15N), as was seagrass at the creek mouth. The δ15N of seagrasses, mangroves and macroalgae ranged from
10.4-19.6‰ at the site of sewage discharge to 2.9-4.5‰ at the reference site, 4 km from the creek mouths. The
δ15N values of seagrass (4.5‰) and mangroves (3.4‰) at the reference site were higher than values reported for
oligotrophic areas of Moreton Bay, but the δ15N of macroalgae (2.9‰) was close to unimpacted eastern
Moreton Bay values. Macroalgae derive nutrients from the water column, whereas seagrass and mangroves take
up nutrients from the sediment. Therefore, deposition of effluent derived N into the sediments is implicated in
the elevated δ15N values of the seagrass and mangroves at the reference site. The free amino acid concentration
and composition of seagrass and macroalgae was used to distinguish uptake of sewage and shrimp derived N.
Proline (seagrass) and serine (macroalgae) were high in sewage impacted plants and glutamine (seagrass) and
alanine (macroalgae) were high in plants impacted by shrimp effluent. The δ15N and amino acid composition
indicated sewage N extended further from the creek mouths than shrimp N. This analysis of physical / chemical
and biological indicators was able to distinguish the composition and subsequent impacts of aquaculture on the
receiving waters.
CHAPTER 2 20
2.1 Introduction
Aquaculture is a rapidly expanding industry, and its effluent can be a major source of
pollution in marine ecosystems (Chua et al., 1989; Twilley, 1989; Gowen et al., 1990;
Braaten, 1991; Holmer, 1991; Phillips et al., 1991; Macintosh & Phillips, 1992; Pruder, 1992;
Raa & Liltved, 1992; Wu et al., 1994; Samocha & Lawrence, 1997; Hargreaves, 1998).
Environmental studies into the effects of shrimp aquaculture are limited and have mostly
focussed on in-pond water quality, with little research conducted into the ecological impacts
of wastewater on receiving waters (Pillay, 1992).
The monitoring of traditional water quality parameters has identified that downstream
impacts of shrimp effluent, and other forms of aquaculture, are only measurable in close
proximity to the discharge point. Hensey (1991) observed that environmental monitoring of
aquaculture effluent using water quality sampling techniques showed no impacts. Samocha
& Lawrence (1997) observed large diurnal fluctuations in water quality parameters measured
downstream of shrimp farm discharge points, and that no increase in total suspended solids
(TSS) or nutrient concentrations could be measured further than 400 m from the farm’s
discharge. It is possible, however, that sediment impacts such as increased organic matter
and anoxia, may extend further (up to 1 km) than water column impacts (Wu et al., 1994).
With the current projected expansion of shrimp farming in most coastal areas of the world,
large scale increases in nutrients and suspended solids in the receiving waters are likely.
Elevated loadings of particulate material to receiving waters have immediate effects on the
receiving environment such as reduced light penetration and smothering of benthic fauna and
flora (Abal et al., 1994). In addition, particle loading may also contribute to longer term
ASSESSING ECOLOGICAL IMPACTS 21
changes through initial downstream settling, with resuspension into the water column at a
later time.
Increases in the concentration of NH4+ in receiving waters from shrimp farms have been
observed by many researchers, and as a result nutrient enrichment of poorly flushed
embayments may occur (Gowen & Rosenthal, 1993). In some instances the level of impact
has been sufficient to result in feedback which affects the aquaculture operation itself
(Gowen et al., 1990). Evidence suggests that serious shrimp farm production losses resulting
from the outbreak of disease in Asia and Latin America, are due to the environmental impacts
of shrimp culture (Phillips et al., 1993). In addition to impacts on the aquaculture operation
itself, shrimp farming has been linked to several cases of environmental degradation,
however, despite this type of evidence, there is still a lack of quantitative data on the
ecological impacts to receiving waters (Phillips et al., 1993).
The need for data on the ecological impact of aquaculture effluent has been identified
(Gowen et al., 1990), and it has been shown that physical and chemical water quality
monitoring techniques cannot provide this information. Bioindicators have long been used to
determine ecological impacts of point source discharges (Worf, 1980; Kramer, 1994). For
example, marine macrophytes can be used to provide insights into the ecological impacts of
nutrients and suspended particulates by measuring changes in plant distributions,
morphology, pigment concentrations and total tissue N (Lyngby, 1990; Alamoudi, 1994;
Horrocks et al., 1995; Abal & Dennison, 1996; Udy & Dennison, 1997b). The morphology
of seagrasses can change with reduced light availability, as a consequence of elevated
concentrations of suspended solids in the water column (Abal et al., 1994). Seagrass
CHAPTER 2 22
distribution and depth penetration are also reduced as a consequence of reduced light
availability in shallow estuarine systems (Dennison et al., 1993; Abal & Dennison, 1996).
Recently, marine plants have been used to detect and integrate the long term effects of small
and /or pulsed nutrient inputs in well flushed oceanic systems (Costanzo, 1996), and elucidate
the possible sources of the nutrient inputs (Jones et al., 1996). Macrophyte amino acid
concentrations and composition have been shown to change with various N sources in both
controlled laboratory experiments (Nasr et al., 1968; Di Martino Rigano et al., 1992; Jones et
al., 1996) and field surveys (Udy & Dennison, 1997b). In particular, accumulation of the
amino acids alanine, glutamine, proline and serine in plants (both terrestrial and marine) has
been associated with N uptake, with different amino acids responding to different N sources
(Steward and Pollard, 1962; Silveira et al., 1985; Lawlor et al., 1987; Kiladze et al., 1989; Di
Martino Rigano et al., 1992; Vona et al., 1992; Heuer & Feigin, 1993). Stable isotope ratios
of nitrogen (δ15N) have been used widely in marine systems as tracers of discharged nitrogen
from point and diffuse sources, including sewage effluent (Rau et al., 1981; Wada et al.,
1987; Van Dover et al., 1992; Macko & Ostrom, 1994; Cifuentes et al., 1996; McClelland &
Valiela, 1998). Elevated δ15N signatures in seagrass, mangroves and macroalgae have been
attributed to plant assimilation of N from treated sewage effluent (Wada et al., 1987; Grice et
al., 1996; Udy & Dennison, 1997b; Abal et al., 1998). This study, however, appears to be the
first to use all these techniques to study the extent of impacts from aquaculture effluent.
In coastal marine systems a variety of point source inputs from aquaculture ponds, sewage
treatment plants, fertiliser plants, agriculture and urban runoff can make it difficult to
determine responsibility for ecological impacts (Grant et al., 1995). With increasing conflict
ASSESSING ECOLOGICAL IMPACTS 23
between users of coastal resources, it has become essential to determine the specific influence
of each source (Teichert-Coddington, 1995).
To assess the potential impact of aquaculture effluent, comparisons between the water
volumes and water quality parameters of aquaculture effluent and treated sewage effluent
have been conducted (Bergheim & Selmer-Olsen, 1982; Muir, 1982; Solbe, 1982; Macintosh
& Phillips, 1992; Paez Osuna et al., 1997). However, there are very few studies comparing
the impacts on the receiving waters (Pearson & Rosenberg, 1978; Cifuentes et al., 1996).
Despite the differences in the two forms of waste, Pearson & Rosenberg (1978) hypothesised
that the impacts of these two sources on receiving sediments would be similar.
Shrimp pond effluent has higher concentrations of suspended solids and phytoplankton
(Ziemann et al., 1992), but lower concentrations of nutrients than sewage effluent (Muir,
1982). Dissolved nutrients in shrimp effluent are predominantly NH4+, whereas sewage
effluent is proportionally higher in NO3-, and PO4
3- (Macintosh & Phillips, 1992). Shrimp
effluent is typically produced in large volumes (Macintosh & Phillips, 1992), which can
equate to up to 40% of the total inputs of N and P in some localised areas (Bergheim &
Selmer-Olsen, 1982). Sewage is freshwater, whereas the salinity of shrimp effluent is
typically 35-36 on the practical salinity scale. These differences may have a considerable
impact on the fate of organisms in the receiving waters when effluent is released into shallow
tidal estuaries. Both sewage and aquaculture effluent can be discharged intermittently,
resulting in large diel fluctuations in water quality. Difficulties in monitoring these variable
discharges can be overcome by the use of biological indicators, which integrate the impacts
of these effluents over time (Costanzo, 1996). Unlike traditional chemical analyses of water
CHAPTER 2 24
column nutrients, these biological indicators reflect the availability of biologically available
nutrients (Lyngby, 1990) which provides more ecologically meaningful information.
The aims of this study were to assess the influences on the receiving environment of
wastewater discharges to a shallow estuarine system. Changes in receiving water and
sediment quality analyses were compared with biological impacts measured as a consequence
of shrimp farm and sewage effluent discharges. The region of influence of these two
pollutant sources is defined, and mechanisms are suggested which may aid in discerning the
relative impacts of these two discharges on a common receiving environment.
ASSESSING ECOLOGICAL IMPACTS 25
2.2 Materials and Methods
2.2.1 Study Region
Moreton Bay is a shallow coastal embayment on the east coast of Australia. The western side
of the bay receives a variety point and non point source inputs including agricultural runoff,
sewage and aquaculture effluent. The eastern bay is well flushed and influenced by oceanic
waters. In eastern Moreton Bay, background concentrations of water quality parameters are:
NH4+ < 2 µM; NO3
- ~ 0.1 µM; PO43- ~ 0.2 µM; chlorophyll a < 1 µg L-1; TSS < 20 mg L-1,
and typical δ15N values for mangroves, seagrass and macroalgae in the eastern bay are 2-3‰
(Abal et al., 1998).
Two tidal creeks in close proximity (1.5 km apart) were studied in Moreton Bay, Australia
(Fig. 2.1). One creek received discharge (18000 m3 d-1 containing 2.0 mg N L-1 and
0.2 mg P L-1, which equates to 36 kg N d-1 and 3.6 kg P d-1) from a shrimp farm (Jones,
unpub. data). The creek was 2-3 m deep, approximately 1 km in length, and the shrimp farm
discharge was 500 m from the mouth of the creek. The intensive shrimp farm (6 ha of ponds)
was stocked with Penaeus japonicus (35 animals m-2). Ponds were routinely flushed (~20%
per day) and water discharged into the creek on low tide. Except during the effluent
discharge, the creek runs dry at low tide. The other creek (Eprapah Creek) received
discharge (2400 m3 d-1 containing 4.5 mg N L-1 and 8.0 mg P L-1, which equates to
10.8 kg N d-1 and 19.2 kg P d-1) from a sewage treatment plant (Redland Shire Council, pers.
comm.). The creek was approximately 2-5 m deep, 15 km in length, has a standing body of
water at low tide, and the sewage discharge point was 2 km from the mouth. The sewage
treatment plant serviced approximately 14 000 people and utilised secondary (activated
sludge) treatment techniques. Both creeks were tidally flushed, and had virtually no
freshwater flow during the study period (Autumn, 1997).
CHAPTER 2 26
2.2.2 Experimental Design
Three sites were chosen in each creek, the first at the nutrient source (discharge site), the
second approximately mid way between the nutrient source and the mouth (middle site), and
the final at the mouth of the creek (mouth site) (Fig. 2.1). A site was positioned midway
between both creeks and in close proximity to the shore (midway site). Several more sites
were selected in Moreton Bay in a radiating pattern out from the creek mouths (Oyster Point,
Sewage Plume and Cox Bank), including a reference site located approximately 4 km from
the creek mouths. The creek banks at low tide extended the mouth of the sewage creek as far
as the sewage plume site. At eight of the sites, traditional water quality parameters (dissolved
N & P, total suspended solids, volatile suspended solids, sediment organic content, secchi
depth, chlorophyll a, and physico-chemical parameters) were determined.
Brisbane
MoretonBay
•
⊗
⊗
0 0.5 1.0
kilometres
N
Eprapah Ck
ReferenceSite
SewageTreatment
Plant
ShrimpFarm
OysterPoint
DischargeSite
DischargeSite
MiddleSite
MiddleSite
Mouth Site
Mouth Site
MidwaySite
Cox BankSite
SewagePlumeSite
Oyster PointSite
VictoriaPoint
PointHalloran
Coochie-mudloIsland
⊗ Nutrient Source
Sampling Site
ASSESSING ECOLOGICAL IMPACTS 27
Figure 2.1 Map of study sites in Moreton Bay, including the location of shrimp and sewage effluent discharges.
Bioindicators were utilised at eleven sites with macroalgae and phytoplankton at all sites,
mangroves at the creek sites, and seagrass at the bay sites. Amino acid composition was
determined for macroalgae and seagrass, and the δ15N signature and total tissue N was
determined for all bioindicator species.
2.2.3 Collection
Samples of seagrass (Zostera capricorni), mangrove (Avicennia marina), and macroalgae
(Catenella nipae) were collected, placed on ice and returned to the laboratory and prepared
for analysis of %N, δ15N and amino acids. In the case of the seagrass and mangroves, the
second youngest leaves were chosen, and for the macroalgae a single mangrove
pneumatophore covered in macroalgae was collected for each replicate. Three replicates for
each plant type were collected at each site.
2.2.4 Analytical Procedures
Salinity was measured with a Horiba U-10 water quality meter (California, U.S.A.) and
expressed on the Practical Salinity Scale.
Chlorophyll a concentration was determined by filtering a known volume of water sample
through Whatman GF/F filters, which were immediately frozen. Acetone extraction and
calculation of chlorophyll a concentration was performed using the methods of Clesceri et al.
(1989), and Parsons et al. (1984).
Light-saturated phytoplankton productivity (potential productivity in mg C m-3 h-1) was
determined in the laboratory using the 14C-bicarbonate incorporation method
CHAPTER 2 28
(Parsons et al., 1984). One hundred millilitres of water from each site was dispensed to three
120 ml polycarbonate bottles. A common dark control was established for each site by
combining 33 ml of sample from each replicate into a fourth bottle wrapped in foil. Aqueous
14C sodium bicarbonate (4 µCi) was added and bottles were incubated at a light intensity of 1100
to 1200 µE m-2 s-1. A recirculating water bath and perspex heat shields maintained temperatures
at ambient levels. After approximately two hours, water samples were filtered through 0.4 µm
polycarbonate filters (Poretics). The filters were placed into 5 ml scintillation vials and two
drops of 5N HCl were added to each vial to drive off any remaining 14CO2. Four millilitres of
scintillation fluid was added to each vial, and radioactivity as disintegrations per minute (DPM)
determined using a scintillation counter (Packard Tricarb 1600TR, Meriden, Connecticut,
U.S.A.). Total CO2 concentration in samples was determined from carbonate alkalinity using
the method of Parsons et al. (1984).
Total suspended solids concentrations were determined using the methods of
Clesceri et al. (1989). A known volume of water was filtered onto a pre-weighed and pre-
dried (110 ºC; 24 h) Whatman GF/C glass fibre filter. The filter was then oven dried at 60 ºC
for 24 h and total suspended solids calculated by comparing the initial and final weights.
Volatile suspended solids were determined as loss on ignition by combusting samples in a
muffle furnace for 12 h at 525 ºC (Clesceri et al., 1989). The organic content of the sediment
was determined from 10 cm deep core samples collected using 50 mL cut-off syringes. The
sediment sample was combusted in a muffle furnace at 525 °C for 12 h and the proportion of
organic material determined by loss on ignition (Clesceri et al., 1989).
Dissolved inorganic nutrients (NH4+, NO3
-/NO2-, and PO4
3-) were determined by filtering
water samples through Whatman GF/F glass fibre filters and freezing them immediately on
ASSESSING ECOLOGICAL IMPACTS 29
dry ice. Samples were analysed within two weeks by the NATA accredited Queensland
Health Analytical Services Laboratory in accordance with the methods of Clesceri et al.
(1989) using a Skalar autoanalyser (Norcross, Georgia, U.S.A.).
For analysis of plant total tissue N, δ15N and amino acids, tissue was rinsed in distilled water
to remove nutrients and sediment from the thallus surface, and then prepared for analysis.
For calculation of total tissue N content and the δ15N isotopic signature, samples were oven
dried to constant weight (24 h at 60 °C), ground and three sub-samples were oxidised in a
Roboprep CN Biological Sample Converter (Europa Tracermass, Crewe, U.K.). The
resultant N2 was analysed by a continuous flow isotope ratio mass spectrometer (Europa
Tracermass, Crewe, U.K.). Total %N of the sample was determined, and the ratio of 15N to
14N was expressed as the relative difference between the sample and a standard (N2 in air)
using the following equation (Peterson & Fry, 1987):
δ15N = (15N/14N (sample) / 15N/14N (standard) – 1) x 1000 (‰)
For amino acid analysis, approximately 1.0 g wet weight of plant tissue was weighed and
placed in 5 mL of 100% methanol (analytical reagent grade) for 24 h to extract amino acids.
The methanol extract was filtered through Millipore Millex - HV13 (0.45 µm) filters and
injected into a post column derivatisation HPLC amino acid analyser (Beckman System
6300, Fullerton, California, U.S.A.), for detection of ninhydrin positive free amino acid
groups at 570 nm. Results were calculated and expressed as µmol g-1 wet weight. As well as
detecting free amino acids, this technique also measures the concentration of free NH4+ in
plant tissue. Changes in amino acid composition were used to infer nutrient source (either
shrimp or sewage effluent). This technique was based on responses observed under ambient
field, as well as controlled laboratory conditions using artificial nutrient additions (Jones et
al., 1996; Udy & Dennison, 1997a).
CHAPTER 2 30
2.2.5 Statistical Analysis
For all sampling techniques, three replicates were analysed and means and standard errors
were calculated. Differences between treatments were tested for significance using one way
analysis of variance (ANOVA) and Tukey's Test for multiple comparison of means at a
significance level of 0.05 using Minitab 12.1 software (State College, Pennsylvania, U.S.A.).
ASSESSING ECOLOGICAL IMPACTS 31
2.3 Results
2.3.1 Physical and Chemical Water Quality Analyses
The concentrations of dissolved nutrients, chlorophyll a, phytoplankton productivity, total
suspended solids, volatile suspended solids, and sediment organic content were different
between the sewage and shrimp creek discharge sites. However, these parameters failed to
detect an impact at the midway site (750 m beyond the mouths of the creeks), with values not
significantly higher than at the reference site (4 km from the creek mouths) (Table 2.1).
2.3.1.1 Salinity
All sites in the shrimp creek and in the bay were close to full salinity seawater (35-36). At
the sewage creek discharge, the salinity was 29 as a consequence of freshwater inputs from
the sewage effluent. Salinity increased downstream to 35 at the mouth site.
2.3.1.2 Nutrients
The concentrations of dissolved nutrients (NH4+, NO3
- / NO2- and PO4
3-) for each creek were
highest at the discharge sites, and declined rapidly towards the mouth site. In particular,
NH4+ concentration decreased to near eastern Moreton Bay concentrations at both creek
mouth sites. The concentration of dissolved nutrients at the midway site (750 m from the
creek mouths) was not significantly different (p > 0.05) from the reference site (~4 km from
the creek mouths) (Table 2.1). The dissolved nutrient ratios were significantly different
(p < 0.05) between the two discharge sites. At the sewage creek discharge site, NH4+ : NO3
- /
NO2- was 0.45, compared to 3.8 for the shrimp creek discharge site. The ratio of DIN (NH4
+
+ NO3- / NO2
-) to DIP (PO43-) at the sewage creek discharge site was 3.0, compared to 24 at
the shrimp creek discharge site. The concentrations of dissolved nutrients at the midway and
CHAPTER 2 32
reference sites were at eastern Moreton Bay levels, and the relative ratios of the dissolved
nutrients were similar to the shrimp creek.
2.3.1.3 Phytoplankton
Chlorophyll a concentration was not significantly different (p > 0.05) between the discharge
site in the shrimp creek (10.8 µg L-1) and the discharge site in the sewage creek (11.1 µg L-1).
However, at the creek mouth sites the concentration in the sewage creek (5.2 µg L-1) was
significantly lower (p < 0.001) than in the shrimp creek (17.9 µg L-1). The concentration of
chlorophyll a at the midway site (2.5 µg L-1) was not significantly higher (p > 0.05) than the
reference site (1.8 µg L-1) (Table 2.1).
Despite the relatively low NH4+ concentration at the shrimp creek discharge site, the
phytoplankton productivity (212 mg C m-3 h-1) was significantly higher (p < 0.001) than at
the sewage creek discharge site (20 mg C m-3 h-1). However, the high productivity at the
shrimp creek discharge site did not result in a significantly higher (p > 0.05) chlorophyll a
concentration. The concentration of chlorophyll a in the shrimp creek increased from
10.8 µg L-1 at the discharge site to 17.9 µg L-1 at the mouth site, probably due to increased
light availability resulting from the reduction in the concentration of inorganic and other
suspended solids. Phytoplankton productivity at the midway site (9 mg C m-3 h-1) was not
significantly different (p > 0.05) to the reference site (11 mg C m-3 h-1) (Table 2.1).
Table 2.1. Results of traditional water quality monitoring for the creek with shrimp farm effluent and sewage treatment effluent. DIN = Dissolved Inorganic Nitrogen;
DIP = Dissolved Inorganic Phosphorus; Chl a = Chlorophyll a; Phyto Prod = phytoplankton productivity; TSS = total suspended solids; VSS = volatile suspended solids;
Secchi = secchi disc depth. Only one replicate measurement was recorded for Secchi disk depth and salinity (Practical Salinity Scale).
Sampling
Site
Salinity
NH4+
(µM)
NO3-/NO2
--
(µM)
PO43-
(µM)
DIN: DIP
Ratio
Chl a
(µg L-1)
Phyto Prod
(mg C m-3 h-1)
TSS
(mg L-1)
VSS
(% of TSS)
Secchi
(m)
Sediment
%Organic
Shrimp Discharge 35 3.8a 1a 0.2a 24c 10.8a 212d 63.5a 19a 0.5 6.0a
Shrimp Middle 35.5 1.6a 0.4a 0.3a 7ab 11.9ab 87c 51.5ab 21a 0.7 6.1a
Shrimp Mouth 36 2.3a 0.3a 0.2a 13bc 17.9b 157b 40.2b 26a 0.5 8.3ab
Midway 36 0.8a 0.2a 0.4a 3a 2.5c 9a 18.2d 25a 1.0+ 7.1a
Sewage Discharge 29 29b 65b 31b 3a 11.1ab 20a 44.3b 35b 1.1 13.5c
Sewage Middle 33 5.4a 8a 5.5a 2a 9.2ac 20a 32.9bc 33b 1.0 7.5ab
Sewage Mouth 35 2.4a 2.9a 2.1a 3a 5.2ac 18a 32.5bc 28ab 1.1 5.6a
Reference 36 1.2a 0.5a 0.3a 6ab 1.8c 11a 20.2d 28ab 1.9 11bc
F Value 40*** 15*** 34*** 13*** 15*** 88*** 35*** 14*** 15***
* p < 0.05; ** p < 0.01; *** p < 0.001. abc Means with different letters are significantly different at p < 0.05.
CHAPTER 2 34
2.3.1.4 Suspended Solids and Secchi Depth
The concentration of total suspended solids (TSS) at the shrimp creek discharge site
(63 mg L-1) was significantly higher than the sewage creek discharge site (44 mg L-1). The
concentrations at the midway (18 mg L-1) and reference sites (22 mg L-1) were not
significantly different (p > 0.05) from each other, but were significantly (p < 0.05) lower than
both creek mouth sites indicating significant sedimentation or dilution (Table 2.1).
The organic fraction of the suspended solids (volatile suspended solids) was significantly
higher (p < 0.001) at the sewage discharge site (35%) compared to the shrimp discharge site
(19%). In the shrimp creek the concentration of organic particles increased towards the
mouth in proportion with the increasing chlorophyll a concentration (r2 = 0.60). In
comparison, the concentration of organic particles in the sewage creek decreased in
proportion with the concentration of chlorophyll a (r2 = 0.8) (Table 2.2).
Secchi disk depths did not vary along the length of either creek, from discharge site to mouth
site. The mean secchi depth in the sewage creek (~1.0 m) was approximately double the
depth in the shrimp creek (~0.6 m), but only half that of the reference site (1.9 m). The
secchi depth at the midway site was greater than 1 m, but water depth was too shallow to
obtain a measurement (Table 2.1).
2.3.1.5 Sediment Organic Content
The organic content of the sediment (loss on ignition) in the shrimp creek increased from
6.0% at the discharge site to 8.3% at the mouth site, probably due to sedimentation. In the
sewage creek, the organic content declined significantly (p < 0.001) from the discharge site
ASSESSING ECOLOGICAL IMPACTS 35
(13.5%) to the mouth site (5.6%), probably due to senescence and subsequent sedimentation
of phytoplankton and other organic particulates near the discharge site (Table 2.1).
Table 2.2. Correlations (r2) between the concentration of phytoplankton (chlorophyll a) and phytoplankton
productivity (14C uptake) and various water quality parameters. DIN = Dissolved Inorganic Nitrogen; DIP =
Dissolved Inorganic Phosphorus; Phyto Prod = phytoplankton productivity (mg C m-3 h-1); Chl a =
Chlorophyll a (µg L-1); TSS = total suspended solids (mg L-1); VSS = volatile suspended solids (mg L-1); ISS =
inorganic suspended solids (mg L-1); Secchi = secchi disc depth (m). Numbers in bold type indicate significant
correlations (r2 ≥ 0.6).
TSS
(mg L-1)
VSS
(mg L-1)
ISS
(mg L-1)
NH4+
(µM)
NO3-/NO2
-
(µM)
PO43-
(µM)
DIN:
DIP
Salinity
Shrimp Creek
Chl a - 0.85 - 0.60 - 0.87 0.12 0.51 0.14 0.09 0.86
Phyto Prod 0.21 0.48 0.19 0.93 0.56 - 0.81 0.94 0.19
Sewage Creek
Chl a 0.60 0.80 0.37 0.66 0.63 0.66 0.04 - 0.85
Phyto Prod 0.27 0.51 0.11 0.34 0.32 0.35 0.25 - 0.60
2.3.2 Bioindicators
In contrast to the water quality parameters, the responses of the bioindicator parameters at
sites beyond the creek mouths were elevated compared to the reference site. For some of the
parameters, the reference site appeared to be influenced by nutrients from the discharges.
2.3.2.1 Tissue Nitrogen Content
The %N of the macroalgae was responsive to the nutrient sources, with the highest value at
the sewage creek discharge site (3.1%), which was significantly higher (p < 0.05) than the
shrimp discharge site (1.9%) (Table 2.3; Fig. 2.2). There was no decrease in the %N of the
macroalgae with distance from the discharge site in the shrimp creek. However, in the
sewage creek the macroalgae at the mouth site had a %N of 1.6%, which was not
CHAPTER 2 36
significantly elevated (p > 0.05) above the reference site (1.5%). The %N of the macroalgae
at the midway site (2.3%) was significantly (p < 0.05) elevated above values in the
macroalgae at both of the creek mouth sites (Table 2.3; Fig. 2.2).
The %N of seagrass leaves was significantly higher (p < 0.05) at the sewage creek mouth site
(2.7%) compared with the shrimp creek mouth site (2.3%). The next three sites distant from
the creek mouths (midway, Oyster Point, and Sewage Plume) were not significantly lower
than the shrimp creek mouth site (2.3%). The seagrass %N at the next most distant site (Cox
Bank) was 2.0%, which was not significantly higher (p < 0.05) than at the reference site
(1.7%) (Table 2.3; Fig. 2.2).
The %N of the mangrove leaves appears less sensitive to nutrient inputs, with none of the
mangroves at the other sites being significantly higher (p < 0.05) than the reference site
(1.7%) (Table 2.3; Fig. 2.2).
2.3.2.2 δ15N Stable Isotope Ratio of Nitrogen
The δ15N isotopic signatures of the seagrass, macroalgae and mangroves were significantly
different (p < 0.001) between sites (Table 2.3; Fig. 2.3). The highest δ15N was in the
macroalgae at the sewage creek discharge site (19.6‰), and the lowest in the macroalgae at
the reference site (2.9‰). δ15N in the macroalgae in the sewage creek decreased with
distance away from the source. The value at the discharge site in the shrimp creek was 7.1‰,
with no significant difference (p > 0.05) along the length of the shrimp creek to the mouth
(7.9‰). The δ15N at the midway site (6.4‰) was not significantly lower (p > 0.05) than the
ASSESSING ECOLOGICAL IMPACTS 37
creek mouth sites, indicating influence of nutrients from the discharges. The δ15N at the
reference site (2.9‰) was significantly lower (p < 0.001) than all other sites.
The δ15N of seagrass leaves at the sewage plume site (8.0‰) was significantly higher
(p < 0.05) than all other sites. The δ15N was not significantly different (p > 0.05) between the
two creek mouths (7.1‰ and 6.8‰), but both were significantly higher (p < 0.05) than the
midway (5.8‰), Oyster Point (4.7‰) and Cox Bank (5.5‰) sites, and the reference site
(4.5‰) (Table 2.3; Fig. 2.3).
The highest δ15N of mangrove leaves was 10.4‰ at the sewage discharge site, compared with
7.7‰ at the shrimp discharge site. Despite the significant differences (p < 0.05) at the
source, the δ15N values at the creek mouths were not significantly different (p > 0.05) from
each other (4.9‰ and 4.6‰). The midway site is a small intertidal sand bank, and as such is
a site of sediment deposition. The δ15N of the mangroves (7.7‰) at the midway site was
significantly higher (p < 0.05) than at both of the creek mouth sites, possibly due to
deposition of δ15N enriched particulates from the creeks. The δ15N of the mangroves at the
reference site (3.4‰) was significantly lower (p < 0.05) than all other sites (Table 2.3;
Fig. 2.3).
CHAPTER 2 38
Table 2.3. Results of bioindicator monitoring for the creek with shrimp farm effluent and sewage treatment
effluent. δ15N = Nitrogen stable isotope ratio; %N = Tissue N content; nd = no data (no plants were present).
Sampling
Site
Mangrove
(Avicennia marina)
Macroalgae
(Catenella nipae)
Seagrass
(Zostera capricorni)
δ15N (‰) %N δ15N (‰) %N δ15N (‰) %N
Shrimp Discharge 7.7d 1.4b 7.1b 1.9ab nd nd
Shrimp Middle 6.2c 1.7a 8.6b 1.8ab nd nd
Shrimp Mouth 4.6ab 1.6a 7.9b 1.9b 7.1de 2.3c
Midway 7.7d 1.6a 6.4b 2.3c 5.8c 2.3c
Sewage Discharge 10.4e 1.7a 19.6d 3.1d nd nd
Sewage Middle 9.4e 1.4b 16.3c 1.7ab nd nd
Sewage Mouth 4.9bc 1.3b 6.4b 1.6ab 6.8d 2.7d
Sewage Plume nd nd nd nd 8.0e 2.1bc
Cox Bank nd nd nd nd 5.5bc 2.0ab
Oyster Point nd nd nd nd 4.7ab 2.2bc
Reference 3.4a 1.7a 2.9a 1.5a 4.5a 1.7a
F Value 73.5*** 12.7*** 70.5*** 49*** 33.3*** 19.6***
*p < 0.05; **p < 0.01; ***p < 0.001. abc Means with different letters are significantly different at p < 0.05.
ASSESSING ECOLOGICAL IMPACTS 39
Brisbane
MoretonBay
•
⊗
⊗
0 0.5 1.0
kilometres
N
Eprapah Ck SewageTreatment
Plant
ShrimpFarm
OysterPoint
VictoriaPoint
PointHalloran
Coochie-mudloIsland
2.2
2.0
2.1
1.9
1.7
3.1
1.5
2.3
1.9
1.6
1.7
2.3
2.3
2.7
1.8
1.4
1.4
1.7
1.7
1.6
1.6
1.3
1.7
Seagrass
Macroalgae
Mangrove
Figure 2.2. Map showing the values of %N in seagrass (Zostera capricorni), macroalgae (Catenella nipae), and
mangroves (Avicennia marina) at the study sites (see Fig. 2.1 for site references).
n Sampling Site ⊗ Nutrient Source
CHAPTER 2 40
Brisbane
MoretonBay
•
⊗
⊗
0 0.5 1.0
kilometres
N
Eprapah Ck SewageTreatment
Plant
ShrimpFarm
OysterPoint
VictoriaPoint
PointHalloran
Coochie-mudloIsland
4.7
5.5
8.0
7.1
16.3
19.6
2.9
6.4
7.9
6.4
4.5
5.8
7.1
6.8
8.6
7.7
9.4
10.4
3.4
7.7
4.6
4.9
6.2
Seagrass
Macroalgae
Mangrove
Figure 2.3. Map showing the values of δ15N in seagrass (Zostera capricorni), macroalgae (Catenella nipae),
and mangroves (Avicennia marina) at the study sites (see Fig. 2.1 for site references).
2.3.2.3 Free Amino Acid Composition
In the macroalgae, the total dissolved free amino acid concentration ranged from 1.7 to
6.5 µmol g wet-1), with the highest concentrations at the shrimp (3.6 µmol g wet
-1) and sewage
(4.1 µmol g wet-1) creek mouth sites and the midway site (6.5 µmol g wet
-1). However, there
was no direct correlation between the total amino acid concentration and dissolved nutrient
availability, with the lowest concentration of amino acids being in the macroalgae at the
sewage discharge site (1.7 µmol g wet-1). The concentration at the reference site
n Sampling Site ⊗ Nutrient Source
ASSESSING ECOLOGICAL IMPACTS 41
(3.7 µmol g wet-1) was only significantly lower (p < 0.05) than the midway site (Table 2.4;
Fig. 2.5).
The total free amino acid pool of the seagrass leaves proved to be more responsive to nutrient
sources than the macroalgae, ranging from 5.3 to 16.7 µmol g wet-1 (Table 2.4; Fig. 2.4). The
total free amino acid concentration in the seagrass at the shrimp (16.7 µmol g wet-1) and
sewage (10.2 µmol g wet-1) creek mouth sites and the midway site (10.4 µmol g wet
-1) were
significantly higher (p < 0.05) than at the reference site (5.3 µmol g wet-1).
The amino acid composition in the plants close to the nutrient sources was used to infer the
source of nutrients being taken up by plants at more distant sites. The percentages of
particular amino acids in the total free amino acid pool of the seagrass (glutamine and
proline) and macroalgae (alanine and serine) were correlated to either the shrimp or sewage
discharge source (Table 2.4; Fig. 2.4; Fig. 2.5). No other amino acids were significantly
different (p > 0.05) between the creek discharge sites for the macroalgae and the creek
mouths sites for the seagrass.
The %alanine (4.5%) in the macroalgae at the shrimp creek discharge site was significantly
higher (p < 0.05) than the macroalgae at the sewage creek discharge site (2.7%). At the
sewage discharge site, the %serine (10.3%) in the macroalgae was significantly higher
(p < 0.05) than at the shrimp discharge site (6.6%). The %serine at the midway site (9.0%)
was not significantly lower (p > 0.05) than the sewage discharge or sewage mouth sites.
However, %serine at the reference site (6.8%) was significantly lower (p < 0.05) than at the
sewage discharge site. In contrast, the %alanine at all sites, except at sites influenced directly
by sewage (the sewage discharge and mouth sites) was not significantly lower (p > 0.05) than
CHAPTER 2 42
the shrimp discharge site, indicating that macroalgae at the reference site may have been
influenced by the shrimp effluent (Table 2.4; Fig. 2.5).
In the seagrass at the shrimp creek mouth site, the %glutamine (30%) was significantly
higher (p < 0.05) than at the sewage creek mouth (15%). On the other hand, the %proline
was significantly higher (p < 0.05) in the seagrass at the sewage creek mouth site (59%)
compared to the shrimp creek mouth site (44%). The %glutamine at the midway site (23%)
was not significantly lower (p > 0.05) than at the shrimp mouth site, however all other sites
were lower, with a minimum of 14% at the reference site. The %proline at all other sites,
except the oyster point site (37%) were not significantly lower (p > 0.05) than at the sewage
mouth site.
Table 2.4. Results of bioindicator monitoring for the shrimp and sewage creeks. % refers to percentage of total
free amino acid pool. SER = serine; ALA = alanine; GLN = glutamine; PRO = proline; Total αα = total
concentration of free amino acids (µmol g wet-1); nd = no data (no plants were present).
Sampling Site Macroalgae (Catenella nipae) Seagrass (Zostera capricorni)
%SER %ALA Total αα
µmol g wet-1
%GLN %PRO Total αα
µmol g wet-1
Shrimp Discharge 6.6b 4.5a 3.5bc nd nd nd
Shrimp Mouth 7.9ab 3.0ab 3.6bc 30a 44b 16.7a
Midway 9.0ab 3.8ab 6.5a 23ab 50ab 10.4b
Sewage Discharge 10.3a 2.7b 1.7c nd nd nd
Sewage Mouth 6.5ab 2.2b 4.1b 15b 59a 10.2b
Sewage Plume nd nd nd 14b 61a 13.5ab
Cox Bank nd nd nd 18b 51ab 7.7bc
Oyster Point nd nd nd 17b 37b 6.2bc
Reference 6.8b 2.9ab 3.7bc 14b 46ab 5.3c
F Value 6.5* 5.9** 9.5** 7.9*** 5.0** 16.4***
*p < 0.05; **p < 0.01; ***p < 0.001. abc Means with different letters are significantly different at p < 0.05.
ASSESSING ECOLOGICAL IMPACTS 43
Brisbane
MoretonBay
•
⊗
⊗
0 0.5 1.0
kilometres
N
Eprapah Ck SewageTreatment
Plant
ShrimpFarm
OysterPoint
VictoriaPoint
PointHalloran
Coochie-mudloIsland
Proline
Glutamine
Other
15000
10000
5000
nmol g-1 wet wt
Figure 2.4. Map showing the amino acid composition of seagrass (Zostera capricorni) at the study sites (see
Fig. 2.1 for site references).
n Sampling Site ⊗ Nutrient Source
CHAPTER 2 44
Brisbane
MoretonBay
•
⊗
⊗
0 0.5 1.0
kilometres
N
Eprapah Ck SewageTreatment
Plant
ShrimpFarm
OysterPoint
VictoriaPoint
PointHalloran
Coochie-mudloIsland
SerineAlanineOther
5000
3000
1000
nmol g-1 wet wt
Figure 2.5. Map showing the amino acid composition of macroalgae (Catenella nipae) at the study sites. Pie
graphs have been reduced to quarters for layout purposes. The remaining three quarters of the pie graphs not
represented are a continuation of the “other” amino acid category (not serine or alanine) (see Fig. 2.1 for site
references).
n Sampling Site ⊗ Nutrient Source
ASSESSING ECOLOGICAL IMPACTS 45
2.4 Discussion
Impacts from the shrimp and sewage effluent discharged into the two creeks were
significantly different, both within the creeks and in the receiving waters. In addition to
differences in the nature of the effluent, the creeks were physically different, with the sewage
creek being longer, wider and deeper. The sewage creek had a longer residence time,
allowing for greater assimilation of nutrients and deposition of sediments within the creek
itself before being released into Moreton Bay. The potential reduction in nutrients and
suspended solids within the creek may have a marked effect on the impacts from the sewage
creek compared with the shrimp creek which has a low residence time and high suspended
solid load.
2.4.1Water Quality Parameters
2.4.1.1 Effluent Composition
Although shrimp farm and treated sewage effluents have similar components, the relative
concentrations and proportions can be significantly different (Macintosh & Phillips, 1992).
The sewage discharge site had significantly higher concentrations of nutrients, but a lower
N: P ratio than at the shrimp discharge site. The ratios of NH4+ to NO3
- / NO2- and dissolved
inorganic nitrogen (DIN) to dissolved inorganic phosphorus (DIP) at the sites in the shrimp
creek were higher than in the sewage creek. These differences in nutrient composition may
have implications regarding the phytoplankton community structure in receiving waters.
Cyanobacteria and green flagellates in shrimp ponds have been observed to dominate over
diatoms in conditions of lower light, higher NH4+ concentrations, and higher DIN: DIP
(Burford, 1997).
CHAPTER 2 46
2.4.1.2 Phytoplankton Biomass and Productivity
Ammonium is generally considered the most biologically available form of N for marine
phytoplankton (McCarthy et al., 1977). This is consistent with the positive correlation of
phytoplankton productivity to NH4+ concentration (r2 = 0.93) but not to NO3
- / NO2-
(r2 = 0.56) observed in the shrimp creek (Table 2.2).
In contrast, phytoplankton productivity in the sewage creek was low (not significantly higher
(p > 0.05) than at the reference site), and not significantly correlated (r2 = 0.34) to NH4+
concentration. The low phytoplankton productivity and diminishing chlorophyll a
concentration (from 11.1 µg L-1 to 5.2 µg L-1) in the sewage creek was probably not a result
of nutrient or light limitation, but rather due to senescence of freshwater phytoplankton with
increasing salinity downstream (Ahel et al., 1996). This is evidenced by the strong negative
correlation between chlorophyll a and salinity (r2 = 0.85).
Partitioning of nutrients within the creeks may be influenced by chlorophyll a concentrations,
TSS and salinity gradients. The concentration of chlorophyll a in the sewage creek was
positively correlated to TSS (r2 = 0.60), suggesting that the phytoplankton make up a
considerable portion of the suspended particulates (Table 2.2). In comparison, 81% of the
TSS at the shrimp discharge site was inorganic. Phytoplankton can take up significant
amounts of NH4+ from the water column, whereas inorganic particulates usually have a high
affinity for sorbing P (Pomeroy et al., 1965).
ASSESSING ECOLOGICAL IMPACTS 47
2.4.2 Biological Indicators
2.4.2.1 Tissue N Content
The tissue N content (%N) of marine plants is a potential indicator of biologically available
nutrient concentrations (Gerloff & Krombholz, 1966; Duarte, 1990), especially in macroalgae
(Horrocks et al., 1995) which have the ability to store large reserves of “luxury” nitrogen for
metabolism during times of nutrient stress. All the %N values of the seagrass and
macroalgae were shown to be higher at sites in close proximity to either the sewage or shrimp
discharges, and at many of the sites within the bay the %N was elevated above the reference
site.
In the shrimp creek, there was no decrease in %N of the macroalgae down the creek away
from the discharge site. In contrast, the %N of the macroalgae at the sewage mouth site was
significantly lower than at the discharge site, indicating significant assimilation of nutrients
by the system within the length of the creek. These differences are probably due to the
greater length of the sewage creek allowing for much greater assimilation and deposition of
nutrients (Cifuentes et al., 1996).
The %N content in the mangrove leaves had no correlation with nutrient availability, and
were relatively unchanged between all sites, despite significant (p < 0.05) changes in δ15N.
This is consistent with another study which found no significant difference between the %N
of mangroves throughout the Moreton Bay, thereby making them less sensitive as indicators
of nutrient inputs than seagrasses and macroalgae (Rogers, 1998).
CHAPTER 2 48
2.4.2.2 δ 15N Isotopic Signature
The δ15N of raw sewage and treated sewage particulates discharging into Moreton Bay is
around 5.1‰ and 9.2‰ respectively (Loneragan et al., in prep.). The elevated δ15N signature
subsequent to treatment of the sewage effluent is a result of isotopic fractionation during
ammonia volatilisation, nitrification and denitrification (McClelland & Valiela, 1998). The
δ15N of particulates in the ponds of the shrimp farm discharging into the shrimp creek is
around 6‰ (Preston, unpub. data). This is considerably higher than the value of 0.9‰
reported for suspended particulate matter in shrimp effluent in Ecuador (Cifuentes et al.,
1996). The farm in the present study uses artificial pelleted feeds made predominantly of
shrimp meal, compared with the farm studied by Cifuentes et al. (1996) in Ecuador which
relies predominantly on natural feed (phytoplankton). The markedly greater δ15N of the
effluent from the farm in the present study may be a result of the animal derived N in the
pelleted feeds versus the plant derived N at the farm in Ecuador.
The δ15N of seagrass (Udy & Dennison, 1997b), macroalgae (Abal et al., 1998) and
mangroves (Rogers, 1998) in Moreton Bay has been shown to be positively correlated to
proximity to nutrient sources. The δ15N of seagrass at sewage impacted sites is
approximately 10‰ (Udy & Dennison, 1997b), which closely reflects the δ15N of treated
sewage effluent (9.2‰) discharged into Moreton Bay (Loneragan et al., in review). In
comparison, the δ15N of seagrasses at unimpacted eastern bay sites ranges from 2 to 3‰ (Udy
& Dennison, 1997b). In the present study, the δ15N stable isotope signature of seagrasses,
mangroves and macroalgae ranged from 2.9‰ to 19.6‰.
ASSESSING ECOLOGICAL IMPACTS 49
Variations in δ15N may be more greatly influenced by isotopic fractionation than by changes
in the relative contributions of the N sources (Fogel & Cifuentes, 1993). The enrichment of
the δ15N of NH4+ in estuaries is mediated predominantly by nitrification of NH4
+ (Mariotti et
al., 1984; Cifuentes et al., 1989; Fogel & Cifuentes, 1993). The δ15N values recorded for the
macroalgae at the sewage creek discharge site (19.6‰) are among the highest reported in the
literature (Owens, 1987). Given the high concentrations of NO3- / NO2
- in the sewage creek,
nitrification probably accounts for the high δ15N observed in the mangroves and macroalgae,
which would predominantly take up the isotopically heavy NH4+, in preference to NO3
- /
NO2- (Hanisak, 1983). The longer residence times of the sewage creek (compared with the
shorter shrimp creek) may also increase the δ15N of the organic N, NH4+ and NO3
- / NO2- due
to further isotopic segregation. The plants may also be taking up NO3- / NO2
- or NH4+ or
organic N, including dissolved free amino acids (Hanisak, 1983). Uptake of organic forms of
N such as urea require more metabolic energy than uptake of NH4+ (Wheeler, 1983),
however, direct uptake of DFAAs may be possible and more energy efficient (Jørgensen,
1982).
The lower δ15N of the mangroves and macroalgae in the shrimp creek reflects the lower
initial δ15N of the shrimp effluent (Preston, unpub. data). The apparent lack of isotopic
fractionation by bacteria within the shrimp creek may be due to less NH4+ being available for
nitrification because the NH4+ in shrimp ponds is typically taken up by phytoplankton and
bacteria, rather than oxidised by nitrifying bacteria (Hargreaves, 1998). Therefore, there is
usually less enrichment of the δ15N isotopic signature in shrimp effluent than in sewage
effluent, even though the initial values may be similar. The sewage effluent also contained a
much greater proportion of organic material than the shrimp farm effluent. This organic
material has the potential to become more greatly enriched in 15N than NH4+ or NO3
- due to
CHAPTER 2 50
isotopic fractionation at each step in the microbial conversion of organic N to NH4+,
subsequently to NO3- and finally to N2 gas via denitrification (Handley & Raven, 1992). The
high concentrations of NO3- / NO2
- resulting from nitrification in the sewage effluent may
also act as the substrate for denitrification, thereby elevating the δ15N of the remaining NO3- /
NO2-.
Strong isotopic enrichment was observed in the macroalgae compared to the mangroves at
the sewage discharge site. The relatively low concentrations of suspended particulates in the
sewage creek may have reduced the transfer of nutrients to the sediments via sedimentation
of material with incorporated or adsorbed nutrients. Mangroves obtain their nutrients from
the sediment, and as such may have sufficient nutrients available so that they can
preferentially take up isotopically light 14N, whereas the macroalgae may be N limited and
will therefore take up the both 14N and the heavier 15N isotope (Wada, 1980). Uptake of
nutrients from nitrogen fixation associated with decomposing leaves may also provide high
concentrations of N with a much less enriched signature (Hicks & Silvester, 1985).
In contrast to the sewage creek discharge site, the mangrove and macroalgal δ15N at the
shrimp creek discharge site were not significantly different. This may be due to
sedimentation of high concentrations of N rich particulates from the shrimp effluent, thereby
making these nutrients available to the mangroves. The δ15N of the mangroves and
macroalgae in the shrimp creek are similar at the discharge site, but diverge downstream with
the mangroves having a lower δ15N, while the macroalgal signature remains unchanged. This
rapid drop in the δ15N of the mangroves downstream is probably a result of significant initial
concentrated sedimentation of N rich particulates near the discharge. The next major
ASSESSING ECOLOGICAL IMPACTS 51
depositional zone appears to be at the sewage plume site (the mouth of the creek at low tide)
with a seagrass δ15N of 8.0‰.
The values for δ15N in the mangroves (3.4‰) and seagrass (4.5‰) at the reference site were
higher than those reported for sites in the eastern bay and other western bay sites which are
not impacted by point source nutrient discharges (2-3‰) (Udy & Dennison, 1997b;
Abal et al., 1998; Rogers, 1998). These elevated values indicate that the seagrass (and
possibly the mangroves) at the reference site are receiving nutrients from point source
discharges. Seagrass and mangroves obtain most of their nutrients from the sediment (Iizumi
& Hattori, 1982; Short & McRoy, 1984; Ziemann et al., 1984), suggesting that biologically
available nutrients from the effluent sources are being transported to the sediment. However,
the values of δ15N in the seagrass at the two creek mouth sites (6.8 and 7.1‰) were similar,
and so δ15N values alone cannot be used to distinguish between shrimp and sewage effluent
sources.
2.4.2.3 Amino Acid Composition
The seagrass at the shrimp creek mouth site had a significantly higher total free amino acid
pool than at the sewage creek mouth site. Increases in the free amino acid pool of both
terrestrial and marine plants have been linked to elevated nutrient availability (Nasholm et
al., 1994; Ohlson et al., 1995; Jones et al., 1996; Udy & Dennison, 1997b; Udy & Dennison,
1997a). However, there is evidence to suggest that increases in amino acids may indicate
growth limitation by factors such as light (Vergara & Niell, 1995) or nutrients (Di Martino
Rigano et al., 1992). Limitation of growth may thereby prevent metabolism of free amino
acids into proteins. Seagrass at the shrimp creek mouth site have lower growth rates and
biomass than several other sites in Moreton Bay, including those close and distant from
CHAPTER 2 52
nutrient sources (Udy & Dennison, 1997b). The comparatively shallow secchi depth at the
shrimp creek mouth (50% of the sewage creek mouth site) indicates that the reduced seagrass
growth rate may be due to light limitation, thereby resulting in the storage of excess nutrients
as amino acids (Udy & Dennison, 1997a).
The macroalgae at the sewage discharge site had the lowest amino acid concentration, but the
highest %N content. It has been shown in macroalgae that uptake of NO3- can represent a
temporary N storage pool which contributes significantly more to the cellular N content than
the dissolved free amino acid pool (Naldi & Wheeler, 1999). The high NO3- / NO2
- to NH4+
ratio of dissolved inorganic N in the sewage creek may have stimulated uptake of NO3- by the
biota. Although most species usually prefer NH4+ because it does not need to be reduced
(D'Elia and DeBoer, 1978), macroalgae can take up NO3- and NH4
+ simultaneously
(Thomas et al., 1987).
The specific free amino acid composition in the seagrass leaves and macroalgae near to the
discharges was shown to reflect the different effluent sources. Previous studies have linked
the source of N (NH4+, NO3
- / NO2- or organic N) with the amino acid composition of various
plants, including macroalgae (Nasr et al., 1968; Di Martino Rigano et al., 1992; Sarimento,
1992; Jones et al., 1996; Flynn et al., 1997). In particular, glutamine concentration is known
to be associated with NH4+ uptake and can inhibit uptake of NO3
- (Nasr et al., 1968). The
presence of elevated proportions of glutamine in the seagrass at the shrimp creek mouth site
was reflective of the high NH4+ (relative to the concentration of NO3
- / NO2-). The presence
of alanine and glutamine in plant tissue is associated with N metabolism, and increases have
been observed in response to an increase in N availability (Steward and Pollard, 1962; Vona
et al., 1992), particularly with NH4+ and urea (Nasr et al., 1968).
ASSESSING E COLOGICAL IMPACTS 53
The concentration of dissolved free amino acids in marine plants has also been linked directly
with the concentration of amino acids in the water column, due to direct uptake from the
water column (Jørgensen, 1982; Hanisak, 1983). It was also determined that the
concentrations of these amino acids was independent of the water column NH4+ or NO3
- /
NO2- concentrations. In particular, direct uptake of glutamine, alanine and serine was
observed (Jørgensen, 1982).
Under controlled laboratory conditions, macroalgae incubated in shrimp farm effluent has
been shown to store both alanine and glutamine (Jones et al., 2002; Chapter 4). The
relative proportions (as a percentage of the total dissolved free amino pool) correlated
directly (r2 = 0.66 for alanine; r2 = 0.60 for glutamine) with the concentration of particulates
and nutrients in the effluent (Jones et al., 2002; Chapter 4). Alanine is used by shrimp for
osmoregulation and is found in high concentrations in shrimp heads (Teerasuntonwat &
Raksakulthai, 1995) which is one of the major components in pelleted shrimp feeds. Up to
25% of the feed pellets applied to the ponds are uneaten and become dissolved in the effluent
(Lin et al., 1993), thereby enriching the effluent in dissolved free amino acids such as
alanine. Glutamine (which occurred in elevated concentrations in seagrass at the shrimp
creek mouth site) is one of the main amino acids in shrimp muscle tissue (Lee et al., 1989). It
is also stored in large concentrations in the haemolymph to overcome ammonia toxicity,
which can often occur in the pond environment (Chen et al., 1994). Alanine and glutamine
are two of the dominant amino acids in shrimp cuticles (Horst, 1989). After moulting the
cuticles are broken down, thereby increasing the concentrations of these amino acids in the
effluent.
CHAPTER 2 54
The relationships described above provide support for the use of alanine and glutamine as
indicators of uptake of shrimp effluent derived nutrients by biota in the receiving waters. The
elevated proportions of alanine (%alanine) in the macroalgae at the shrimp creek discharge
site may be reflecting the relatively high availability of NH4+, or may be related to the
presence of elevated concentrations of alanine in the water column. The magnitude of the
responses in the %alanine observed between the shrimp and sewage discharge sites are
similar to the responses to NH4+ versus NO3
- supply under controlled laboratory conditions
(Nasr et al., 1968).
The %serine in the macroalgae was correlated with the proximity to the sewage discharge
site, which contained predominantly NO3- / NO2
- as the main form dissolved inorganic N.
The marine alga, Cyanidium caldarium has been shown to accumulate more serine grown in
NO3- than in NH4
+ (Di Martino Rigano et al., 1992). Serine has been shown to accumulate in
some terrestrial plants in response to increasing NO3- (Silveira et al., 1985; Lawlor et al.,
1987; Kiladze et al., 1989), but not with NH4+ supply. In addition to synthesis of serine,
some marine plants can take it up directly from the water column (Jørgensen, 1982;
Alamoudi, 1988). These relationships suggest that the elevated concentrations of serine in
the macroalgae near the sewage discharge were related to the high NO3- concentrations, or
perhaps the presence of high concentrations of dissolved serine in the effluent.
The %proline of the seagrass in close proximity to the sewage creek was significantly higher
relative to other sites. Proline accumulation in plants as a response to heavy metal toxicity is
widespread (Sharma et al., 1998). Plants incubated in sewage effluent can have elevated
proline concentrations in response to high cadmium and lead levels found in sewage effluent
(Mohan & Hosetti, 1998). In terrestrial plants, proline has been observed to increase in
ASSESSING ECOLOGICAL IMPACTS 55
response to NO3- uptake (Heuer & Feigin, 1993), and proline has been shown to stimulate
nitrate reductase activity in aquatic plants (Salonen & Simola, 1989), probably by facilitating
iron transport which is essential for nitrate reductase (Wilkinson, 1994). Accumulation of
proline in response to either heavy metals or nitrate in the sewage effluent may indicate its
potential as an indicator in seagrasses.
At the reference site the %proline in the seagrass was elevated above concentrations
measured in at unimpacted eastern Moreton Bay sites (Udy & Dennison, 1997b), and was
indicative of the response in the seagrass at the sewage creek mouth site. This is consistent
with the high δ15N value in the seagrass relative to the macroalgae at this site. This reflects
the availability of effluent derived N in the sediment, but not in the water column.
Tracing changes in the amino acid composition of seagrass and macroalgae at various
distances from the discharges indicated that the N from both sewage and shrimp discharges
may be influencing these macrophytes.
2.4.3 Comparison of Impacts
Sewage effluent is freshwater, relatively low in suspended solids, and high in NO3- and PO4
3-.
In contrast, shrimp effluent is saltwater, contains a high concentration of highly productive
phytoplankton, total suspended solids, NH4+ and organic nitrogen (both of which are very
biologically available forms of nitrogen) (Macintosh & Phillips, 1992; Ziemann et al., 1992).
The high concentrations of biologically available nutrients in the effluent stimulate high rates
of primary productivity, resulting in production of phytoplankton, thereby reducing light
availability for seagrasses (Abal & Dennison, 1996) and subsequently increasing
sedimentation (Frid & Mercer, 1989).
CHAPTER 2 56
Both adsorbed and interstitial nitrogen are higher at the sewage creek mouth than the shrimp
creek (Udy & Dennison, 1997b), and are elevated in comparison to the rest of Moreton Bay
(Abal et al., 1998). Because N is typically the limiting nutrient in Moreton Bay (Horrocks et
al., 1995; O'Donohue & Dennison, 1997; Udy & Dennison, 1997a; Jones et al., 1998;
Appendix 1), and NH4+ is the preferred N source for seagrass (Iizumi & Hattori, 1982), the
high concentrations of NH4+ in the sediments at the sewage creek mouth site resulted in
higher seagrass shoot density and growth rate (Udy & Dennison, 1997b). Low sediment
NH4+ concentrations at the shrimp creek mouth may be a result of the high concentrations of
phytoplankton limiting the movement of NH4+ to the sediment (Hargreaves, 1998). NH4
+ can
also be sorbed to clay particles, but is more likely to be associated with organic particles in
the sediment (Rosenfeld, 1979). The high concentration of particulates at the shrimp creek
mouth resulted in reduced light availability. The seagrass canopy height at the shrimp creek
mouth is higher than at the sewage creek mouth (Udy & Dennison, 1997b), which is
consistent with responses observed for light limitation (Abal et al., 1994). Despite the results
of some studies regarding the similarity of impacts on sediment (Pearson & Rosenberg,
1978), the results from the present study reveal that there may be quite different impacts from
shrimp versus sewage waste.
Results from the seagrass amino acid composition (%proline) and the δ15N signature suggest
that impacts of the sewage nutrients may be experienced further from the discharge than
shrimp impacts. The greater response by the seagrass at the reference site is postulated as
being due to transport of nutrients to the sediments. This correlates to observations that
impacts from waste discharges on the sediments can extend further than water column
impacts (Wu, 1995). See Figure 2.6 for a conceptual representation of these interpretations.
ASSESSING ECOLOGICAL IMPACTS 57
2.4.4 Conclusion
Physical and chemical water quality analyses alone would have concluded that there was no
influence by either the shrimp or sewage effluent at the midway site, which is approximately
750 m from each creek mouth. This is consistent with the findings of Samocha & Lawrence
(1997) who found that shrimp effluent could not be detected at a distance of 400 m from the
source, and Hensey (1991) who observed no impacts from fish farming using water quality
sampling techniques. This indicates that the nutrients are being taken up by the system
(sediments and biota), reduced by dilution, or being lost to the atmosphere through
volatilisation and denitrification.
Several workers have suggested that aquaculture and other wastewater inputs should be
restricted so that their discharge is within the assimilative ability of the ecosystem (Omori et
al., 1994). In the present study, the concentrations of dissolved nutrients, chlorophyll a, and
total suspended solids were at eastern Moreton Bay concentrations 750 m from the creek
mouths, and the biota had increased %N, δ15N and free amino pool. This suggests that
system was assimilating the nutrients being discharged. However, it can be argued that the
assimilation of nutrients by the system does not indicate an absence of ecological impact
(Welch, 1992). This is particularly important given recent interest in using mangrove forests
as filters for shrimp effluent (Robertson & Phillips, 1995).
Although amino acid composition, δ15N signature and other bioindicator parameters do not
directly elucidate environmental impact, these physiological changes may help to identify
incipient environmental degradation before large scale changes occur in water quality and
community structure, which are the usual signs of environmental degradation.
CHAPTER 2 58
Application of bioindicators was successful at identifying the separate impacts from shrimp
and sewage effluent, and in contrast to chemical and physical water quality analyses,
bioindicators identified spatially broader impacts on ecosystem biota. The δ15N isotopic
signature and amino acid composition of the macroalgae appeared to be most sensitive to
dissolved nutrients. The same parameters in the seagrass appeared to be sensitive to sediment
nutrients, thereby elucidating depositional zones of point source derived nutrients.
This study represents one of the first attempts at detecting the influence of shrimp effluent on
receiving waters using marine plants as bioindicators. It also attempted to distinguish this
influence from sewage, another common point source input. To further develop these
techniques for broad application will require more controlled laboratory experiments to
enable the elimination of confounding factors to isolate the bioindicator responses from
undiluted nutrient sources.
2.4.5 Application for other types of Aquaculture
The impacts of aquaculture effluent can vary considerably depending primarily on the quality
of water supplying the ponds (Boyd & Musig, 1992; Pruder, 1992), the farm management
practices (feeding rates, water exchange), pond soil characteristics, proximity to other farms,
and the flushing rates of the receiving water body (Gowen et al., 1990).
Shrimp farming in Australia is still a small but burgeoning industry, with production
increasing from 15 t to 2,000 t over the last 14 years (Preston, pers. comm.). The relatively
large geographical spread of existing farms means that overcrowding is not currently a
problem, in contrast to certain regions in Asia and Latin America (Phillips et al., 1993).
ASSESSING ECOLOGICAL IMPACTS 59
The two creeks studied are small, and the receiving waters are well flushed and low in
nutrients relative to many of the world’s sites of aquaculture discharge where the quality of
the intake / receiving waters can be so poor as to impact on the aquaculture operation itself
(Chua et al., 1989; Aitken, 1990; Wong et al., 1992). The farm discharging into the “shrimp”
creek consisted of only 6 ha of ponds, and so represents a minor input relative to more
densely farmed regions, where the effluent from several farms are discharging into the same
water body. In some regions, with a dense distribution of farms, the effluent from one farm
can become another’s intake (New, 1990 cited in Macintosh & Phillips, 1992). The inputs of
sewage effluent were also relatively small, with the treatment facility servicing only 14 000
people. Monitoring of small scale discharges like these can provide an indication of how
small inputs can still have a significant impact on receiving waters.
Studies into the impacts on receiving waters for shrimp farming have been very poorly
quantified in comparison to other forms of aquaculture (Phillips et al., 1993). In comparing
the impacts between different operations, there are several differences that need to be
considered. Intensive culture of shrimp has the highest water volume requirements of all
forms of aquaculture (Phillips et al., 1991). Sedimentation which works well for some forms
of aquaculture is typically not as effective for shrimp because of clay particles and high
concentrations of phytoplankton which do not easily settle out (Macintosh & Phillips, 1992).
2.4.6 Remediation Options
To reduce environmental impacts, reductions in nutrient and sediment discharge can be
achieved with changes to management practices (reviewed in Allan et al., 1995; Hopkins et
al., 1995b). Options include more regular feeding to help reduce wastage (Villalon, 1991
CHAPTER 2 60
cited in Samocha & Lawrence, 1997), proper design of settlement ponds to better promote
sedimentation (Samocha & Lawrence, 1997), reduced water exchange (Hopkins et al.,
1993b), sludge removal (Hopkins et al., 1994) and biological filtration using oysters and
macroalgae (Wang & Jakob, 1991; Hopkins et al., 1993a; Shpigel et al., 1993b; Samocha &
Lawrence, 1997; Jones et al., 2002; Chapter 4; Jones & Preston, 1999; Chapter 3; Jones
et al., 2001b ; Chapter 5). In particular, polyculture with oysters and macroalgae is
regarded as having considerable environmental and economic benefits (Chandrkrachang et
al., 1991; Macintosh & Phillips, 1992), but further research and development is needed
(Macintosh & Phillips, 1992).
Figure 2.6. Conceptual model of the two creeks and the range and type of impacts from the different effluent sources.
CHAPTER 3
OYSTER FILTRATION OF SHRIMP FARM EFFLUENT,
THE EFFECTS ON WATER QUALITY
Abstract
Shrimp pond effluent water can contain higher concentrations of dissolved nutrients and suspended particulates
than the influent water. Consequently, there are concerns about adverse environmental impacts on coastal
waters due to eutrophication and increased turbidity. One potential method of improving effluent water quality,
prior to discharge or recirculation, is to use bivalves to filter the effluent. This study examined effects of the
Sydney Rock Oyster, Saccostrea commercialis (Iredale and Roughley) on the water quality of shrimp pond
effluent. Effluent from a shrimp farm stocked with Penaeus japonicus (Bate) was pumped directly into tanks
(34 L) stocked with different densities of oysters. Combinations of live and dead oysters were used to test the
effects of three different densities of live oysters (24, 16 and 8 live oysters per tank). The concentrations of total
suspended solids, proportion of organic and inorganic matter, total nitrogen, total phosphorous, chlorophyll a
and the total number of bacteria in the pond effluent water were determined before and after filtration by
oysters. The oysters significantly reduced the concentration of all the parameters examined, with the highest
oyster density having the greatest effect. Shrimp pond effluent contained a higher proportion of inorganic
matter (72%) than organic matter (28%). The organic component appeared to be mainly detritus, with
chlorophyll a comprising only a minor proportion. Filtration by the high density of oysters reduced the effluent
total suspended solids to 49% of the initial level, the bacterial numbers to 58%, total nitrogen to 80% and total P
to 67%. The combined effects of settlement and oyster filtration reduced the concentration of chlorophyll a to
8% of the initial effluent value.
CHAPTER 3 64
3.1 Introduction
In shrimp farming, the addition of feeds and action of pond aerators can result in increased
nutrient and sediment loads in pond effluent compared to the influent water (Phillips et al.,
1993; Briggs & Funge-Smith, 1994). Consequently, there are concerns about adverse
environmental impacts on coastal waters due to increased turbidity and eutrophication
(Ziemann et al., 1992; Hopkins et al., 1993a). There is a need to develop more effective
controls of the quality of effluent water prior to discharge into the environment or
recirculation (Phillips et al., 1993; Primavera, 1994).
One potential method of reducing adverse environmental impacts, and recapturing otherwise
wasted nutrients, is to use bivalves to filter the pond effluent (Lin et al., 1993). Pond effluent
contains organic matter including bacteria, phytoplankton, and detritus (Ziemann et al., 1992)
that could provide food for bivalves such as oysters (Hopkins et al., 1993a). Pond effluent
can also contain a high proportion small inorganic particles (Hopkins et al., 1995b) that could
be removed from suspension by oyster filtration and subsequently expelled as larger, more
settleable, particles in the form of pseudofaeces (Tenore & Dunstan, 1973).
This study investigated the filtration effects of oysters, Saccostrea commercialis on the water
quality of shrimp farm effluent. The research was conducted in Moreton Bay, Australia
(Fig. 3.1), a region that supports shrimp farming and traditional oyster cultivation. In
Moreton Bay shrimp farming and oyster cultivation are both seasonal activities. In winter
(June to August) shrimp ponds are empty and oysters are generally not harvested from the
bay due to their poor condition. In spring (September to November) shrimp ponds are
fertilised to promote phytoplankton growth and ponds are stocked with shrimp post-larvae.
Oyster growth and conditioning occurs during spring and summer when phytoplankton
EFFECTS OF OYSTER FILTRATION ON WATER QUALITY 65
densities in the bay increase (Dennison et al., 1993). Oyster growth to marketable size takes
two to three years (Witney et al., 1988).
This study used a system of tanks supplied with water pumped from the effluent canal of a
commercial shrimp farm stocked with P. japonicus. The farm’s management practices can
be considered intensive (25 animals per m2), with a regime of periodic water exchange and no
recirculation. The objective of the study was to determine the effects of oyster filtration on
pond effluent water quality.
CHAPTER 3 66
Figure 3.1. Location map of Moreton Bay Prawn Farm near Brisbane, Australia.
EFFECTS OF OYSTER FILTRATION ON WATER QUALITY 67
3.2 Materials and Methods
3.2.1 Experimental Design
The effects of oyster filtration on shrimp pond effluent were determined by monitoring the
inflow and outflow water in each of 15 outdoor plastic tanks (60 cm × 40 cm × 25 cm)
stocked with oysters (Fig. 3.2). Each tank was supplied with 34 L of water pumped directly
from the effluent channel of a commercial shrimp farm. The effluent channel received water
from 6 x 1 ha ponds with a stocking density of approximately 25 shrimp m2. At the time of
the experiments (Summer 1995 / 1996) the mean wet weight of the shrimp was
approximately 10 g.
Wat
er I
nflo
w
Water Outflow
Inflow
Outflow
Tank Layout
Single Tank with oyster tray
60 cm
25 cm
40 cm
Figure 3.2. Schematic representation of tank and waterflow layout.
The 15 tanks were stocked with oysters (S. commercialis), the mean wet weight of individual
oysters was 55 g. In the experimental design, combinations of live and dead oysters (empty
shells) were used to produce 3 different densities of live oysters; low, medium and high plus
CHAPTER 3 68
one treatment of dead oysters (Table 3.1). There were 3 replicates for each density of
oysters, and a control treatment of no oysters. The dead oysters were included to elucidate
any changes in water circulation and settling characteristics effected by the physical
characteristics of the oyster shells.
Table 3.1 Combinations of live and dead oysters (Saccostrea commercialis) used in experiments to determine
the effects of oyster density on the water quality of shrimp pond effluent.
Treatment Description Live Oyster Density (m2)
Control No oysters 0
Shells 24 dead oysters 0
Low Density 8 live oysters and 16 dead 33
Medium Density 16 live oysters and 8 dead 67
High Density 24 live oysters 100
The inflow water samples were collected at the start of the experiment, and the outflow
samples, after a period of two hours during which the oysters were filtering the effluent (no
water flow to simulate conditions in a shrimp farm treatment pond). Sampling consisted of
collecting three replicate one litre containers of water from the inflow pipe and from the
outflow point of each of the 15 tanks. The water samples were returned the laboratory where
they were filtered for chlorophyll a extraction and TSS calculation. Sub-samples were taken
for total N and P, bacterial numbers and determination of the organic/inorganic ratio.
3.2.2 Analytical Procedures
Chlorophyll a was determined by filtering a known volume of water sample through
Whatman GF/C filters, which were immediately frozen. Acetone extraction and calculation
EFFECTS OF OYSTER FILTRATION ON WATER QUALITY 69
of chlorophyll a concentration was performed using the methods of Clesceri et al. (1989), and
Parsons et al. (1984).
Total suspended solids concentrations were determined using the methods of
Clesceri et al. (1989). A known volume of water was filtered onto a pre-weighed and pre-
dried (110 ºC; 24 h) Whatman GF/C glass fibre filter. The filter was then oven dried at 60 ºC
for 24 h and total suspended solids calculated by comparing the initial and final weights.
Volatile suspended solids were determined as loss on ignition by combusting samples in a
muffle furnace for 12 h at 525 ºC (Clesceri et al., 1989).
Unfiltered samples for nutrient analysis (total Kjeldahl nitrogen and total phosphorus) were
collected in 120 mL polycarbonate immediately frozen. They were subsequently analysed
within two weeks by the NATA accredited Queensland Health Analytical Services
Laboratory in accordance with the methods of Clesceri et al. (1989) using a Skalar
autoanalyser (Norcross, Georgia, U.S.A.).
Bacteria samples were preserved with 2% formalin and kept at 4°C until analysis. A known
volume (0.5 mL - 1 mL) of sample was stained with acridine orange, filtered onto a stained
(Irgalan Black) 2µm poretics filter and mounted on a slide. Bacteria were counted using
epifluorescence microscopy (Hobbie, 1977).
CHAPTER 3 70
3.3 Results
3.3.1 Suspended Solids
The mean concentrations of TSS (and organic/inorganic ratio), total nitrogen, total
phosphorous, chlorophyll a and the total number of bacteria in the pond effluent water before
and after filtration by oysters are summarised in Table 3.2.
Oysters were effective in removing total suspended solids (TSS) from pond effluent with
significant reductions by the medium and high oyster densities and no significant reduction in
the low density oysters and control treatments. The reduction in effluent TSS varied with
oyster density. At the high density the concentration of TSS was reduced to 49% of the
initial level in the pond effluent, a significantly greater reduction than at the low density
treatment (80%). The level of reduction at the medium density (64%) was not significantly
different from the high or low densities.
3.3.2 Organic content
Combustion of pond effluent filtrate showed that the effluent contained a higher proportion of
inorganic matter (72%) than organic matter (28%). Comparison between the inflow and the
control treatments showed that there was no significant settling of total organic or total
inorganic matter during the experiment. The oysters removed both organic and inorganic
material from pond effluent approximately in proportion to the concentrations initially
present.
Table 3.2 Concentration of various water quality parameters before and after filtration by oysters at 3 different densities (see Table 3.1). Values for control (no oysters) and
shells (dead shells only) are also given. Values in brackets are concentrations expressed as a percentage of the inflow value. Values in italics are standard errors.
Treatment TSS
g L-1 (%)
Organic
g L-1 (%)
Inorganic
g L-1 (%)
Total N
mg L-1 (%)
Total P
mg L-1 (%)
Bacteria
no. × 106 mL (%)
Chlorophyll a
µg L-1 (%)
Inflow 0.13a
0.0052
0.035a
0.0007
0.091a
0.004
1.40ab
0.00
0.15ab
0.00
22.8a
1.35
44.1a
4.1
Control 0.13a (100)
0.0019
0.038a (110)
0.004
0.089a (97)
0.003
1.46a (104)
0.04
0.16ab (107)
0.008
18.9a (83)
1.01
16.9b (38)
1.03
Shells 0.13a (100)
0.002
0.036a (104)
0.0006
0.091a (99)
0.002
1.53a (110)
0.04
0.17a (116)
0.006
19.5a (85)
0.73
20.8b (47)
0.51
Low 0.10ab (80)
0.007
0.033a (95)
0.004
0.068b (75)
0.004
1.24ab (89)
0.15
0.15ac (101)
0.02
20.9a (92)
1.09
13.6bc (31)
0.64
Medium 0.08bc (64)
0.002
0.025ab (71)
0.0006
0.056bc (62)
0.001
1.23ab (88)
0.04
0.12bcd (83)
0.004
19.2a (84)
0.8
8.3cd (19)
0.53
High 0.06c (49)
0.0009
0.018b (52)
0.001
0.043c (47)
0.001
1.12b (80)
0.06
0.10d (67)
0.006
13.2b (58)
1.05
3.6d (8)
0.39
F-Value 21.8*** 6.2.** 33.8*** 5.1* 8.3** 7.5** 54.4***
* p ≤ 0.05; ** p ≤ 0.01; *** p ≤ 0.001. abc means with different letters are significantly different at p < 0.05.
CHAPTER 3 72
3.3.3 Chlorophyll a
Organisms containing chlorophyll a were a minor component of the total organic matter in
pond effluent. There was significant settlement of chlorophyll a in the controls containing no
oysters or only dead oysters. In addition to settlement in the controls, the concentration of
chlorophyll a was significantly reduced by the high and medium density oyster treatments but
there was no significant difference between the low density treatment and the controls. The
combined effects of settlement and oyster filtration reduced the concentration of
chlorophyll a to 8% of the initial effluent value in the high density and 19% in the medium
density of oysters.
3.3.4 Bacteria
The mean number of total bacteria in pond effluent was 22.8 x 106 mL-1. The controls and the
medium and low density oyster treatments had no significant effects on bacterial
concentration. The high density of oysters significantly reduced the total bacterial numbers to
58% of the initial effluent concentration.
3.3.5 Total Nutrients
The medium and low densities of oysters had no significant effects on the concentration of
total N in the effluent, but the high density oyster treatment significantly reduced the total N
concentration to 80% of the initial level in the pond effluent. The total P concentration in
high and medium density oyster treatments was significantly lower than in the influent water,
but the low density treatment had no significant effect. The high and medium density oyster
treatment reduced the effluent total P concentration to 67% and 83% of the initial level,
respectively.
EFFECTS OF OYSTER FILTRATION ON WATER QUALITY 73
3.4 Discussion
The sediment and nutrient levels in the effluent from the P. japonicus shrimp farm were
comparable to those reported in other studies of shrimp farm pond water and effluent
(Ziemann et al., 1992; Burford, 1997; Samocha & Lawrence, 1997), but the concentrations of
TKN and TP were an order of magnitude lower than those reported by Hopkins et al.
(1993a). The level of chlorophyll a in the effluent from the P. japonicus ponds was
indicative of phytoplankton productivity in shrimp ponds (Burford, 1997). However,
organisms containing chlorophyll a were a minor component of the organic matter (the rest
consisting of faecal material, undissolved feed pellets and detritus) and approximately 60% of
the chlorophyll a was removed by settlement alone. Filtration by oysters removed a
significant proportion of the remaining suspended phytoplankton together with other organic
matter, including bacteria, the level of reduction depending on the oyster density. The
control treatments of no oysters or only oyster shells showed that most of the organic matter
remained in suspension during the two hour experiment. This organic matter probably
consisted of fine detritus derived from shrimp feed (Briggs & Funge-Smith, 1994). The high
density of oysters effectively halved the total organic load by removing this organic matter
from the effluent.
The net effect of oyster filtration on effluent nutrient loads reflects the balance between
uptake, excretion and remineralisation of nutrients from settled faeces. In this study, the
proportion of dissolved or particulate fractions in the total phosphorous or nitrogen load were
not determined. The net reduction of 23% of the total phosphorus at the highest density of
oysters was probably due to the removal of phosphorous bound to organic or inorganic
particulates. The net reduction of total nitrogen was less effective, possibly due to removal
being balanced by nitrogen excreted by the oysters. It is possible that assimilation of
CHAPTER 3 74
dissolved nutrients by macroalgae could be used to counter the excretion by oysters (Shpigel
et al., 1993).
The high proportion (72%) of inorganic matter in the effluent from the P. japonicus farm is
characteristic of effluent from unlined earthen shrimp ponds (Ziemann et al., 1992). The lack
of any significant settlement of the inorganic component during the two hour period of the
experiments indicates that most of the inorganic matter was in the form of very fine particles.
Filtration by oysters was effective in removing this suspended sediment, with the high density
of oysters approximately halving the effluent load. During filtration, oysters sort particles by
size and weight; rejected material (pseudofaeces) is accumulated and then expelled through
the inhalant opening (Barnes, 1994). Oysters may, therefore, serve a useful role in removing
small inorganic particles from effluent. However, high sediment loads can reduce or even
arrest oyster filtration (Loosanoff & Tommers, 1948). Further studies are needed to
determine the effects of high inorganic sediment loads on oyster growth and survival. There
is also a need to determine the effectiveness of reducing silt loads in sedimentation ponds,
prior to oyster filtration. High organic loads can also reduce oyster filtration efficiency (Ali,
1970), resulting in the production of large amounts of pseudofaeces (i.e., food filtered but not
ingested) and a concomitant increase in deposition rates due to an inefficient use of filtered
food (Tenore & Dunstan, 1973). Although oysters are inefficient at ingesting phytoplankton
and other food particles at very high concentrations, their ability to convert this material to
pseudofaeces still results in a net removal from the system. However, if concentrations are
too high, recycling of the effluent through the oysters several times may improve the
efficiency of the system (Jones et al., 2002; Chapter 4).
EFFECTS OF OYSTER FILTRATION ON WATER QUALITY 75
3.4.1 Scaling Up Calculations
Despite these limitations, it is possible to calculate an estimate of the number of oysters
required to obtain the observed improvements in water quality and the pond area that would
need to be devoted to them. Based on 20% water exchange per day, a 1 ha pond would need
120,000 oysters. At the same stocking density used in the high density treatment in this study
(1 oyster per 0.01 m2), the oysters would occupy 0.12 ha. Therefore 12% of the area of ponds
must be set aside for oyster filtration. If greater improvements in water quality were required
the number of oysters (and hence pond area) could be increased. These calculations are
comparable to those from Wang (1990), who suggested 150, 000 adult sized oysters per
hectare of shrimp pond would produce reductions in suspended solids similar to those
observed in the present study. If commercial production of oysters is to be achieved
efficiently, a full range of oyster sizes must be stocked in treatment ponds to ensure a
constant supply of commercial sized oysters. Estimates by Wang (1990) have shown that
360, 000 oysters of varying size would be needed to produce 12, 000 oysters per week from
the effluent from a 1 ha shrimp pond.
3.4.2 Summary
In summary, the results of this short-term study demonstrated that oysters can significantly
improve the water quality of shrimp pond effluent by the filtration and retention of suspended
organic and inorganic matter. Longer term studies are needed to determine the effectiveness
of oyster filtration at other stages of the of the shrimp growout season and the effects of
shrimp effluent on the growth, filtration rates and survival of the oysters.
CHAPTER 4
THE EFFICIENCY AND CONDITION OF OYSTERS AND MACROALGAE USED
AS BIOLOGICAL FILTERS OF SHRIMP POND EFFLUENT
Abstract
Current shrimp pond management practices generally result in elevated concentrations of nutrients, suspended
solids, bacteria, and phytoplankton compared to the influent water. Concerns about adverse environmental
impacts due to discharging pond effluent directly into adjacent waterways have prompted the search for cost-
effective methods of effluent treatment. One potential method of effluent treatment is the use of ponds or
raceways stocked with plants or animals that act as natural biofilters by removing waste nutrients. In addition to
improving effluent water quality prior to discharge, the use of natural biofilters provides a method for capturing
otherwise wasted nutrients. This study examined the potential of the native oyster, Saccostrea commercialis
(Iredale and Roughley) and macroalgae, Gracilaria edulis (Gmelin) Silva to improve effluent water quality from
a commercial Penaeus japonicus (Bate) shrimp farm. A system of raceways was constructed to permit
recirculation of the effluent through the oysters to maximise the filtration of bacteria, phytoplankton and total
suspended solids. A series of experiments were conducted to test the ability of oysters and macroalgae to
improve effluent water quality in a flow-through system compared to a recirculation system. In the flow-
through system oysters reduced the concentration of bacteria to 35% of the initial concentration, chlorophyll a to
39%, total particles (2.28 µm - 35.2 µm) to 29%, total nitrogen to 66% and total phosphorus to 56%. Under the
recirculating flow regime, the ability of the oysters to improve water quality was significantly enhanced. After
four circuits, total bacterial numbers were reduced to 12%, chlorophyll a to 4%, and total suspended solids to
16%. Efforts to increase biofiltration by adding additional layers of oyster trays and macroalgal mesh bags
resulted in fouling of the lower layers causing the death of oysters and senescence of macroalgae.
Supplementary laboratory experiments were designed to examine the effects of high effluent concentrations of
suspended particulates on the growth and condition of oysters and macroalgae. The results demonstrated that
high concentrations of particulates inhibited growth and reduced the condition of oysters and macroalgae.
Allowing the effluent to settle before biofiltration improved growth and reduced signs of stress in the oysters
and macroalgae. A settling time of 6 h reduced particulates to a level that prevented fouling of the oysters and
macroalgae.
CHAPTER 4 78
4.1 Introduction
In many countries, including Australia, effluent from aquaculture ponds is released into
receiving waterways without treatment (Eng et al., 1989). Consequently, there are concerns
about potential adverse environmental impacts on coastal waters due to eutrophication and
increased turbidity (Ziemann et al., 1992; Primavera, 1994). These concerns have become a
risk factor for the aquaculture industry (Braaten, 1991). With the expansion of pond
aquaculture sites in Australia, both industry and resource managers are becoming
increasingly aware that the industry may require stricter environmental controls, such as
those currently required for sewage discharge (Samocha & Lawrence, 1997). This has
prompted efforts to develop cost-effective methods of effluent treatment. In addition to the
environmental aspects, recovery of the nutrients associated with the degraded pelleted feeds
could be of considerable economic benefit.
Due to the use of paddle wheel aerators that scour soil from the sides and bottom of earthen
farm ponds, suspended inorganic solids are one of the principal components of pond effluent.
The removal of these particulates from the pond effluent could probably be promoted by
structural changes to effluent channels such as incorporation of baffles. The effluent also
contains significant amounts of suspended organic particulates, predominantly
phytoplankton, protists and bacteria. Due to the small size of the suspended organic and
inorganic particles in pond effluent, removal by mechanical filtration is often very difficult
and prohibitively expensive (Hopkins et al., 1995b). The smaller particles have a similar
specific gravity to seawater and therefore do not readily settle out of suspension (Rubel &
Hager Inc., 1979). Flocculating agents could be used to enhance sedimentation, but they are
considered prohibitively expensive for application in aquaculture (Norris 1994 in Samocha &
Lawrence, 1997).
EFFICIENCY AND CONDITION OF BIOFILT ERS 79
The use of filter feeding bivalves such as oysters, mussels or clams as natural biofilters could
be effective in removing the small particles from the effluent (Wang, 1990; Hopkins et al.,
1993a). After the larger particles have been settled out of suspension (to prevent fouling of
the oysters), bivalves will feed on the smaller, non-settleable particles such as phytoplankton,
bacteria and other organic material. Small inorganic particles are also filtered, coagulated
into larger, more settleable particles and egested as pseudofaeces (Tenore & Dunstan, 1973).
The organic particles ingested by the oysters are incorporated into tissue, thereby capturing
wasted nutrients and converting them into a secondary cash crop.
To ensure success of oysters cultured in shrimp pond effluent, adequate water flow and
circulation must be maintained such that sufficient oxygen is supplied, the buildup of soluble
metabolites is minimised and the fouling by settling particles reduced (Thielker, 1981). This
circulation also ensures an even distribution of food, which is important to prevent
differential growth in the oysters (Scura et al., 1979).
Water flow rate can have a significant effect on the ability of oysters to filter particulates
(Walne, 1972). If flow rates cannot be optimised for oyster filtration, treatment may require
recirculation through the oyster pond several times, or have a larger treatment area. A large
treatment area is economically less viable due to loss of production area for production of the
primary high value crop (Chien & Liao, 1995), and consequently recirculation may be more
viable depending on pumping costs.
Previous studies have shown that oysters can significantly improve the water quality of
effluent from shrimp ponds (Wang & Jakob, 1991; Hopkins et al., 1993a; Jones & Preston,
1999; Chapter 3; Jones et al., 2001b; Chapter 5). The aims of the present study were the
CHAPTER 4 80
following: a) characterise changes in the biological and chemical composition of shrimp pond
effluent following biofiltration by oysters and macroalgae in raceways; b) determine the
differences in water quality improvements under flow-through and recirculating flow; and c)
determine the effects of the high suspended solids in shrimp pond effluent on the growth and
condition of oysters and macroalgae.
EFFICIENCY AND CONDITION OF BIOFILT ERS 81
4.2 Materials and Methods
4.2.1 Experimental Design
A series of field and laboratory experiments were conducted to test the efficiency of
biofiltration of shrimp pond effluent, to improve discharge water quality or enable reuse of
water in production ponds. Different densities of oysters and macroalgae, and different flow
regimes (continuous flow versus recirculating) were tested for the ability to improve the
water quality of shrimp farm effluent, and the impacts of the effluent on the growth and
condition of the biofilter organisms.
4.2.1.1 Filtration Efficiency Experiments
Twelve 1500 L concrete raceways (Plate 4.1), based on the design by Scura et al. (1979),
were constructed and supplied with water from the effluent canal of a commercial Penaeus
japonicus (Bate) shrimp farm which discharges into Moreton Bay, Queensland, Australia.
Each raceway was divided into chambers by baffles to ensure good mixing and circulation of
the water through the oysters (Fig. 4.1A). The raceways were stocked with locally collected
Sydney Rock oysters, Saccostrea commercialis (Iredale and Roughley) and the red
macroalga, Gracilaria edulis (Gmelin) Silva.
A series of experiments were conducted, with both a continuous flow through and
recirculating flow regimes investigated. Before all experiments, oysters were maintained in
the raceways for a period of two weeks, and macroalgae for two days prior to study initiation.
The first continuous flow experiment was conducted with 55 g oysters stocked at three
different densities in plastic oyster trays stacked three layers deep. The low density treatment
was 18 oysters per 100 L (270 per raceway), the medium density was 36 oysters per 100 L
CHAPTER 4 82
(540 per raceway) and the high density 72 oysters per 100 L (1080 per raceway). Three
replicate raceways for each density treatment were stocked with oysters, and three replicate
raceways containing empty oyster trays were maintained as controls. Water quality
parameters were sampled at the inflow and outflow point of each raceway every 12 h for
72 h. The data presented is an average of all six sampling times. Three replicate one litre
containers of water were collected from the inflow and outflow point of each raceway for
determination of particle size, bacteria, phytoplankton (chlorophyll a), total nitrogen (N) and
total phosphorus (P).
Plate 4.1 Raceways constructed at Moreton Bay Prawn Farm, Queensland, Australia.
The second continuous flow experiment used three different densities of macroalgae in mesh
bags (100 cm × 50 cm) constructed using plastic mesh (15 mm mesh size) folded in half and
cable tied together. Three replicate one litre containers of water were collected from the
inflow and outflow point of each raceway for determination of dissolved nutrients (NH4+,
NO3- / NO2
-, and PO43-).
EFFICIENCY AND CONDITION OF BIOFILT ERS 83
The recirculating flow experiment was conducted with 35 g oysters stocked at 10.8 per 100 L
(162 per raceway). The lower density was chosen to ensure that, based on estimated oyster
pumping rates (Wang, 1990), not all the volume of water in raceways could be filtered within
the 2 h residence time. Measurements were made after the first circuit and compared to the
control raceways. The effluent from all twelve raceways flowed into a holding tank and then
pumped back into the raceways. Therefore, after the first circuit all twelve raceways were
functioning as a combined treatment unit. Water quality sampling was conducted at two
hourly intervals (the time required for full water exchange through the raceways) for four
circuits to determine the increasing improvement in water quality with successive filtering by
the oysters. Three replicate one litre containers of water were collected from the inflow and
outflow point of each raceway for determination of total suspended solids (including organic
content), particle size, bacteria, and phytoplankton (chlorophyll a).
4.2.1.2 Biofilter Condition Experiments
Controlled laboratory experiments to determine the impacts of high loadings of
suspended particulates on the growth and condition of oysters and macroalgae were
conducted over a period of eight weeks with effluent replaced weekly (Fig. 4.1B).
Effluent from the discharge channel at Moreton Bay Prawn Farm was collected and
transported to the laboratory in 30 L plastic drums.
In the laboratory, any settled particles in the sample were resuspended and a suite of physico-
chemical parameters (temperature, pH, salinity, dissolved oxygen, and turbidity) were
determined with a Horiba U-10 water quality meter (California, U.S.A.). Three drums of
effluent were then settled for 1 h, 6 h, and 24 h, respectively. A fourth drum was settled for
24 h and then treated by oyster filtration for another 24 h. The fifth drum was stirred just
CHAPTER 4 84
prior to transfer to the oyster and macroalgal tanks. A control treatment utilised sand filtered
seawater from Moreton Bay. This procedure was repeated weekly for eight weeks. The pre-
treated effluent was transferred into four separate tanks (containing oysters or macroalgae)
for each sedimentation treatment.
Figure 4.1 Diagrammatic representation of experimental setup, a) single raceway with baffles and oyster trays,
and b) laboratory settling experiment. NTU = nephelometric turbidity units. The oysters used in the
experiments were Sydney Rock oysters, Saccostrea commercialis and the macroalgae was Gracilaria edulis.
Effluent was from an intensive Penaeus japonicus shrimp farm.
A
B
EFFICIENCY AND CONDITION OF BIOFILT ERS 85
Oysters and macroalgae were incubated in 11 L tanks and were raised off the bottom by a
plastic mesh tray. Three macroalgal tanks (50 g of macroalgae each) and one oyster tank
(8 oysters) were maintained for each sedimentation treatment. Tanks were maintained at
20 − 23°C, and exposed to light on a 12:12 h light / dark cycle using daylight fluorescent
tubes which provided approximately 250 µmol quanta m-2 s-1. Tanks were aerated vigorously
to maintain water movement and dissolved oxygen concentrations.
Water was changed weekly, with water quality analyses conducted prior to changing the
incubation water. Physico-chemical parameters (turbidity, salinity, dissolved oxygen,
temperature, and pH) were measured, as well as oyster biomass and volume, macroalgal
biomass (wet wt), pigment content (chlorophyll a and phycoerythrin), %N, δ15N and
amino acid content.
Photosynthetic capacity of the macroalgae was also measured as an indicator of macroalgal
condition. Experiments were conducted using the same experimental tanks described for the
previous biofilter condition experiment. The effluent was settled for 24 h and then filtered by
oysters for 12 h prior to incubation with macroalgae. The experiment was conducted under
solar radiation, with tanks shaded to 50% of incident light. Three replicate treatment and
seawater control tanks were maintained. Effluent was replaced every twelve hours for 7 d.
Daily measurements of photosynthetic capacity (electron transport rate) and
photosynthetically active radiation (PAR) were taken using a pulse amplitude modulated
fluorometer, DIVINGPAM (Walz GmbH. Effeltrich, Germany). Three replicate
measurements were taken for each tank (a total of nine seawater control and nine shrimp
effluent measurements).
CHAPTER 4 86
4.2.2 Analytical Procedures
Chlorophyll a was determined by filtering a known volume of water sample through
Whatman GF/F filters, which were immediately frozen. Acetone extraction and calculation
of chlorophyll a concentration was performed using the methods of Clesceri et al. (1989), and
Parsons et al. (1984).
The filtrate collected from filtering for chlorophyll a analysis was collected in 120 mL
polycarbonate containers and immediately frozen. NH4+ and NO3
-/NO2- and PO4
3- were
determined by the Queensland Health Analytical Services (NATA accredited) using a Skalar
autoanalyser.
Unfiltered samples for nutrient analysis (total Kjeldahl nitrogen and total phosphorus) were
collected in 120 mL polycarbonate immediately frozen. They were subsequently analysed
within two weeks by the NATA accredited Queensland Health Analytical Services
Laboratory in accordance with the methods of Clesceri et al. (1989) using a Skalar
autoanalyser (Norcross, Georgia, U.S.A.).
Total suspended solids concentrations were determined using the methods of
Clesceri et al. (1989). A known volume of water was filtered onto a pre-weighed and pre-
dried (110 ºC; 24 h) Whatman GF/C glass fibre filter. The filter was then oven dried at 60 ºC
for 24 h and total suspended solids calculated by comparing the initial and final weights.
Volatile suspended solids were determined as loss on ignition by combusting samples in a
muffle furnace for 12 h at 525 ºC (Clesceri et al., 1989).
EFFICIENCY AND CONDITION OF BIOFILT ERS 87
For analysis of particle size number and distribution, a subsample of 30 mL was stored in a
polystyrene container, and fixed with 5% formalin and refrigerated for subsequent analysis.
The sample was later placed into a Coulter counter, which counted the number of particles in
the range from 2.282 µm to 35.2 µm inclusive.
Bacteria samples were preserved with 2% formalin and kept at 4°C until analysis. A known
volume (0.5 mL - 1 mL) of sample was stained with acridine orange, filtered onto a stained
(Irgalan Black) 2 µm poretics filter and mounted on a slide. Bacteria were counted using
epifluorescence microscopy (Hobbie, 1977).
For analysis of plant pigments (phycoerythrin and chlorophyll a), total tissue N, δ15N and
amino acids, tissue was removed, rinsed in distilled water to remove any nutrients and
sediment from the thallus surface, and then prepared for analysis.
For calculation of total tissue N content and the δ15N isotopic signature, samples were oven
dried to constant weight (24 h at 60° C), ground and three sub-samples were oxidised (Dumas
combustion) in a Roboprep CN Biological Sample Converter (Europa Tracermass, Crewe,
U.K.). The resultant N2 was analysed by a continuous flow isotope ratio mass
spectrophotometer (Europa Tracermass, Crewe, U.K.). Total %N of the sample was
determined, and the ratio of 15N to 14N was expressed as the relative difference between the
sample and a standard (N2 in air) using the following equation (Peterson & Fry, 1987):
δ15N = (15N/14N (sample) / 15N/14N (standard) – 1) x 1000 (‰)
For amino acid analysis, approximately 1.0 g wet weight of algal tissue was weighed and
placed in 5 mL of 100% methanol (analytical reagent grade) for 24 h to extract amino acids.
CHAPTER 4 88
The methanol extract was filtered through Millipore Millex - HV13 (0.45 µm) filters and
injected into a Beckman System 6300 post column derivatisation HPLC amino acid analyser,
for detection of ninhydrin positive free amino acid groups at 570 nm. Results were calculated
and expressed as µmol g-1 wet weight. As well as detecting free amino acids, this technique
also measures the concentration of free NH4+ in plant tissue.
For pigment analysis, approximately 0.5 g fresh weight of G. edulis tissue was separated,
blotted dry and weighed (wet wt). The tissue was ground with a mortar and pestle in a
phosphate buffer solution (pH 6.5) to disrupt the cells. The extract was then poured into a
glass graduated centrifuge tube, made up to 10 mL, and centrifuged (20 min at 2500 rpm) to
produce a supernatant containing phycoerythrin and a pellet with the remaining tissue. The
supernatant was transferred to a cuvette and absorption was determined on a
spectrophotometer at 565 nm and 710 nm for phycoerythrin and turbidity blank, respectively.
The pellet was resuspended in 5 mL of 80% acetone (analytical reagent grade) and disrupted
with a tissue homogeniser to extract chlorophyll. Samples were re-centrifuged (20 min at
2500 rpm), and absorbance was determined at 664 nm and 710 nm for chlorophyll a and
turbidity blank, respectively. Pigment concentrations as mg g -1 dry wt were calculated with
specific formulas for phycoerythrin (Rowan, 1989) and chlorophyll a (Parsons et al., 1984).
Photosynthetic capacity as electron transport rate (ETR) was determined at a range of light
intensities by generation of rapid light curves using the pulse amplitude modulated (PAM)
fluorometer. The rapid light curves were generated over a 90 sec period with 10 sec periods
of actinic irradiance with 1 sec saturating pulses for measurement of quantum yield (White &
Critchley, 1999). Data was stored in the diving PAM for subsequent download to computer.
EFFICIENCY AND CONDITION OF BIOFILT ERS 89
Placement of the fibre optic cable was standardised by attaching it to a leaf clip, which was
located near the growing tip of the thallus.
For all sampling techniques, three replicates were analysed and means and standard errors
were calculated. Differences between treatments were tested for significance using one way
analysis of variance (ANOVA) and Tukey's Test for multiple comparison of means.
CHAPTER 4 90
4.3 Results
4.3.1 Filtration Efficiency Experiments
4.3.1.1 Continual Flow
On completion of sampling, oyster mortality was determined in each tray (Fig. 4.2A). For
each density treatment, the percent mortality increased from the upper tray to the lower tray.
Total mortality for all trays was significantly greater in the high density treatment. This
mortality reduced the number of live organisms available to filter the effluent, and resulted in
the low density treatment being more effective at improving the effluent water quality.
Therefore, only the data from the low density treatment is presented.
During the macroalgal filtration experiment, the total concentration of dissolved nutrients
decreased in the control raceway, but increased in all three macroalgal density treatments
(Fig. 4.2B). NH4+ was the predominant dissolved nutrient and increased significantly
(p < 0.001) after flowing through the macroalgal raceways. This increase was probably due
to decay resulting from fouling of the thallus surface by settling material.
Oysters were effective at reducing bacterial numbers from the shrimp pond effluent,
significantly (p < 0.01) decreasing the concentration to 39% of the initial level in the effluent,
with no significant (p > 0.05) change in the control raceways (Table 4.1). In addition to
significant (p < 0.05) settling in the control raceways (to 76% of the initial concentration),
oyster filtration significantly (p < 0.001) reduced the concentration of chlorophyll a to 30%
of the initial concentration. Oyster filtration significantly (p < 0.001) reduced the
concentrations of total N and total P to 64% of initial concentration and to 55% of initial
concentration, respectively. There was also some natural sedimentation of particles in the
control raceways, with the concentration of total N decreasing significantly (p < 0.05) to
EFFICIENCY AND CONDITION OF BIOFILT ERS 91
89%, and total P decreasing significantly (p < 0.05) to 84% of the initial concentrations
(Table 4.1).
0 10 20 30 40 50 60 70 80 90 1 0 0
Lower Tray
Middle Tray
Upper Tray
Oyster Mortality (% of initial)
High Medium Low
-4
-2
0
2
4
6
8
10
12
Macroalgal DensityCh
ange
in
Dis
solv
ed N
utr
ien
t C
once
ntr
atio
n
(µM
)
N H 4+
N O3-
PO 43 -
Control Low Medium High
Figure 4.2 Impacts of effluent on biofilters: a) Oyster mortality (%) from upper, middle and lower trays after 2
weeks at low, medium and high oyster stocking densities in raceways supplied with unsettled shrimp effluent,
and b) change in dissolved nutrient concentrations after passing effluent through low, medium and high
macroalgal stocking densities in raceways supplied with unsettled shrimp effluent. Positive change represents
an increase, negative change represents a decrease.
A
B
CHAPTER 4 92
Table 4.1 Water quality parameters after filtration by oysters under flow through conditions in raceways.
Total N = total Kjeldahl nitrogen; Total P = total phosphorus.
Treatment Bacteria
(no. × 109 L-1)
Chlorophyll a
(µg L-1)
Total N
(µM)
Total P
(µM)
Inflow 12.3a 35.6a 134a 5.8a
Control Outflow 13.3a 27.1b 119b 4.9b
Oyster Outflow 4.8b 10.7c 86c 3.2c
F Value 11** 93*** 66*** 190***
* p < 0.05;
** p < 0.01;
*** p < 0.001.
abc means with different letters are significantly different at p < 0.05.
Ninety five percent of the suspended particles in the effluent were approximately 2-4 µm in
diameter (Fig. 4.3A). Filtration by the oysters significantly decreased (p < 0.001) the
concentration of particulates to 36% of initial concentration. However, the number of
particles in the control raceways increased to 119% of the initial concentration, due to
breaking up of aggregated particles. This was evidenced by a reduction in the mean particle
size.
4.3.1.2 Recirculating Experiments
After the first circuit of the effluent through the raceways, comparisons were made between
the water quality parameters in treatment and control raceways (Table 4.2). In addition to
significant (p < 0.05) reductions in bacteria (77% of the initial concentration) after passing
through the control raceways, oyster filtration further reduced (p < 0.001) the bacterial
numbers to 45% of the initial concentration. There was no significant settling of
phytoplankton (chlorophyll a) in the control raceway, but oyster filtration decreased the
concentration significantly (p < 0.01) to 70% of the initial level.
EFFICIENCY AND CONDITION OF BIOFILT ERS 93
0
5
1 0
1 5
2 0
2 5
3 0
3 5
0 2 4 6 8 10 12 14
Part i c l e S i ze (µm)
Par
ticl
e C
once
ntr
atio
n
(no.
× 1
06
L-1
) Inflow
Control Outflow
Oyster Outflow
Detect ion
Limit
0
5
1 0
1 5
2 0
2 5
0 2 4 6 8 1 0 1 2 1 4
Part i c l e S i ze (µm)
Par
ticl
e C
once
ntr
atio
n
(no.
× 1
06 L
-1)
Initial
Control
First
Second
Third
Fourth
Detect ion
Limit
1
10
100
1000
10000
100000
0 5 10 15 20 25 30 35 40
Partic le Size (µm)
Par
ticl
e C
once
ntr
atio
n
(no.
× 1
06 L
-1)
Initial
Control
First
Second
Third
Fourth
Detect ion
Limit
Figure 4.3 Particle size distribution, a) before and after control and oyster treatment raceways during single
continuous flow, b) before and after consecutive circuits through oyster treatment raceways (linear scale), and c)
before and after consecutive circuits through oyster treatment raceways (log scale).
A
B
C
CHAPTER 4 94
Table 4.2 Water quality parameters after filtration by oysters after the first circuit during recirculating flow in
raceways. TSS = total suspended solids; Organic = organic component of TSS (loss on ignition); Inorganic =
inorganic component of TSS.
Treatment Bacteria
(no. × 109 L-1)
Chlorophyll a
(µg L-1)
TSS
(mg L-1)
Organic
(mg L-1)
Inorganic
(mg L-1)
Inflow 26.5a 26.3a 140a 41 99a
Control Outflow 20.5b 20.8ab 130ab 41 89ab
Oyster Outflow 11.8c 18.5b 110b 32 78b
F Value 65*** 5.4** 4.9** 2.9 4.7**
* p < 0.05;
** p < 0.01;
*** p < 0.001.
abc means with different letters are significantly different at p < 0.05.
Comparison of total suspended solids concentrations in the inflow and outflow of the
raceways showed that oyster filtration was effective in reducing the concentration of both
organic and inorganic particulates (Table 4.2). In the control raceways, settling reduced the
concentration of total suspended solids to 93% of initial, with all the reduction being
inorganic particles (88% of the initial concentration). However, in the oyster raceways the
concentration of inorganic particles decreased significantly (p < 0.01) to 76% and organic
particles to 78% of the initial concentration. The number of particles decreased to 80% of the
initial after the first circuit (Fig. 4.3B).
The results from the first circuit of the recirculating flow experiment revealed that the smaller
35 g oysters (Table 4.2) were considerably less effective at improving the effluent water than
the larger 55 g oysters (Table 4.1) used in the continuous flow experiment. The net
reductions (relative to the control raceway) in bacteria, chlorophyll a and particle
concentration by the 55 g oysters were 50%, 130% and 120% greater than by the 35 g
oysters, respectively.
EFFICIENCY AND CONDITION OF BIOFILT ERS 95
After four circuits through the raceways, the concentrations of all water quality parameters
were significantly (p < 0.001) lower than after the first circuit and after treatment by the
larger oysters in the continuous flow experiment (Fig. 4.4). The number of bacteria was
reduced to 12% of the initial concentration, chlorophyll a concentration to 20%, and TSS to
19%. The slope of the curve over the consequent circuits through the raceways varied
between parameters. Bacteria and chlorophyll a both exhibited curvilinear relationships,
asymptotically approaching zero. In contrast, the relationship between total suspended solids
and number of circuits was linear. The first circuit through the oysters reduced the
concentration of bacteria by 47% of the initial concentration, 23% in the second circuit, but
only 11% and 8% in the third and fourth circuits, respectively. A similar trend was observed
with reductions in chlorophyll a, with a 32% reduction in the first, 55% in the second, but
only 9% in the third circuit, and a small increase in the fourth circuit. In contrast, total
suspended solids was reduced by 23% of the initial concentration, 10% in the second circuit,
33% in the third and 14% in the fourth circuit.
The total number of particles was reduced to 13% of the initial concentration after four
circuits. Particle concentration expressed on a log scale revealed an increase in the number of
the larger particles with successive passes through the raceways, probably due to coagulation
of smaller particles by oysters into larger faeces and pseudofaeces (Fig. 4.3C).
CHAPTER 4 96
0
5
10
15
20
25
30B
act
eria
(n
o.
x1
09
L-1
)
0
5
10
15
20
25
30
Ch
loro
ph
yll
a (
µg
L-1
)
0
20
40
60
80
100
120
140
Initial First Second Third Fourth
Number of Circuits
TS
S (
mg
L-1
)
Figure 4.4 Concentrations of water quality components before and after consecutive circuits through oyster
treatment raceways, a) bacterial numbers, b) chlorophyll a concentration, and c) total suspended solids (TSS).
A
B
C
EFFICIENCY AND CONDITION OF BIOFILT ERS 97
4.3.2 Biofilter Condition Experiments
The mean turbidity of the oyster and macroalgal treatment tanks over the eight week
experimental period ranged from 61 nephelometric turbidity units (NTU) (raw effluent)
to 8 NTU (24 h settled) (Fig. 4.1B). The turbidity of the filtered seawater control was
0 NTU, and the 24 hr settled + oyster filtered treatment was 2 NTU.
The mean growth rate of oysters was greatest in effluent that had been pre-settled for 6
or 24 h (Fig. 4.5A). However, there was a high level of variation in growth rates and no
significant difference in relation to the duration of pre-settlement of effluent.
Macroalgal biomass was significantly lower in the seawater control and unsettled
effluent, otherwise there was no significant difference in relation to the duration of pre-
settlement of effluent (Fig. 4.5B).
Macroalgal growth (as number of new shoots) and pigment content were maximal in the 24 h
plus oyster filtration treatment (Fig. 4.6). There were new shoots present in all treatments
except the seawater control. The number of new shoots increased with increased settling
time.
The responses in new shoot growth and pigment concentration was correlated with the
responses in %N and δ15N (Fig. 4.7). Tissue N content increased from the initial (0.7%)
in all but the seawater control, in which a decrease to 0.28% was observed. The 0 h and
24 h + oyster treatment had the two highest tissue N values (0.8 and 0.9%, respectively).
The δ15N of the macroalgae increased from the initial (6.9‰) in all treatments except
the seawater control (unchanged), or the 24 h plus oyster treatment which decreased to
5.0‰. The highest value was recorded in the 0 h treatment (10.9‰).
CHAPTER 4 98
0
2
4
6
8O
yst
er G
row
th R
ate
(cm
3 o
yst
er-1
)
0
20
40
60
(0 h)
SeawaterControl
0 1 6 24 24 h +
OysterFiltration
Effluent Settling Time (h)
Mac
roal
gal
Bio
mas
s (g
tan
k-1
) Init ial Biomass
n.d.
Figure 4.5 Growth of oysters and macroalgae after 8 weeks in tanks supplied with shrimp effluent pre-settled
for 0, 1, 6 & 24 h. a) change in oyster growth rate expressed as changes in oyster volume (cm3 oyster -1), and
b) macroalgal biomass. n.d. = no data.
A
B
EFFICIENCY AND CONDITION OF BIOFILT ERS 99
0
10
20
30
40M
acr
oa
lga
l G
row
th
(no
. o
f n
ew s
ho
ots
ta
nk
-1)
0
0.2
0.4
0.6
0.8
1
1.2
1.4
1.6
(0 h)
Seawater
Control
0 1 6 24 24 h + Oyster
FiltrationEffluent Settling Time (h)
Pig
men
t C
on
cen
tra
tio
n
(mg
g-1
dry
wt) PE CHL
Figure 4.6 Response of macroalgae to 8 weeks in tanks supplied with shrimp effluent pre-settled for 0, 1, 6 &
24 h. a) macroalgal growth expressed as number of news shoots per tank, and b) concentration of the
photosynthetic pigments, chlorophyll a (CHL) and phycoerythrin (PE).
A
B
CHAPTER 4 100
0
2
4
6
8
10
12
(0 h)
Seawater
Control
0 1 6 24 24 h + Oyster
FiltrationEff luent Sett l ing Time (h)
δδ 1
5N
0
0.2
0.4
0.6
0.8
1
1.2
1.4
1.6
1.8
2T
issu
e N
itro
gen
(%
N)
Initial
Initial
Figure 4.7 Macroalgal nitrogen content (a) and δ15N (b) after 8 weeks in tanks supplied with shrimp effluent of
different settlement times, a) %N, and b) δ15N.
The total concentration of free amino acids in the macroalgal tissue was highly variable,
ranging from 0.65 µmol g wet-1 in the seawater control to 4.1 µmol g wet
-1 in the 24 h
settling treatment (Table 4.3). These changes in the total free amino acid pools are
probably due to an interplay between light availability (both light limitation and light
inhibition) and nutrient availability. Changes in the amino acid composition were also
A
B
EFFICIENCY AND CONDITION OF BIOFILT ERS 101
observed (Table 4.3). Citrulline was present only in the 6 h and 24 h treatments, but in
these treatments citrulline constituted 32% and 48%, respectively. Other amino acids
such as glutamine, phenylalanine and serine showed significant increases in the 24 h
plus oyster treatment, whereas alanine was highest in the 0 h treatment. The proportion
of glutamate ranged from 21% to 45% of the total free amino acid pool, and like
alanine, was highest in the 0 h treatment.
Table 4.3 Changes in the free amino concentration and composition of macroalgae for various treatments in
laboratory settling experiments. % refers to percentage of total free amino acid pool. CIT = citrulline; GLU =
glutamate; ALA = alanine; GLN = glutamine; PHE = phenylalanine; SER = serine; Total αα = total
concentration of free amino acids (µmol g wet-1).
Treatment CIT
(%)
GLU
(%)
ALA
(%)
GLN
(%)
PHE
(%)
SER
(%)
Total αα
(µmol g wet-1)
(0 h) Seawater Control 0 40 2.4 7.7 0 4.4 0.65
Effluent settled for 0 h 0 45 8.5 5.1 9.9 8.8 3.3
Effluent settled for 1 h 0 30 7.7 3.0 9.4 7 1.8
Effluent settled for 6 h 32 24 7.2 2.2 5.6 5.1 3.5
Effluent settled for 24 h 48 21 2.6 2.4 6.3 3.8 4.1
24 h + Oyster Filtration 0 35 5.2 11 12 16 2.5
The maximum photosynthetic capacity (as electron transport rate) of the macroalgae
incubated in the settled and oyster filtered shrimp effluent was 49 µmol e- m-2 s-1 versus
28 µmol e- m-2 s-1 in the macroalgae incubated in the seawater control (Fig. 4.8). Slight
photoinhibition was evident from 1000 µE m-2 s-1 in macroalgae incubated in the
seawater control, but not in the shrimp effluent treatment. In both treatments the
maximum electron transport rate was observed at 300-500 µE m-2 s-1.
CHAPTER 4 102
0
1 0
2 0
3 0
4 0
5 0
6 0
0 200 400 600 800 1 0 0 0 1 2 0 0 1 4 0 0 1 6 0 0 1 8 0 0
P A R ( µ m o l m-2
s-1
)
ET
R (
µm
ol e
- m-2
s-1
)
Cont ro l Set t led Shr imp Eff luent
Figure 4.8 The response of electron transport rate (ETR) versus photosynthetically active radiation (PAR) in
macroalgae incubated in seawater (control) or shrimp effluent (settled 24 h plus oyster filtered for 12 h).
EFFICIENCY AND CONDITION OF BIOFILT ERS 103
4.4 Discussion
Previous studies of the use of biofilters in aqua culture include investigation of the
effectiveness of bivalves (Mann & Ryther, 1977; Shpigel & Fridman, 1990; Shpigel &
Blaycock, 1991; Shpigel et al., 1993a) and macroalgae (Harlin, 1978; Harlin et al., 1979;
Vandermeulen & Gordin, 1990; Cohen & Neori, 1991; Neori et al., 1991; Haglund &
Pedersen, 1993; Subander et al., 1993). Shrimp farms present a particular problem for the
application of biofilters due to the high concentrations of phytoplankton and clay minerals in
pond effluent (Ziemann et al., 1992) which do not settle out as easily as the particulates in
other forms of aquaculture (Macintosh & Phillips, 1992). Application of biofilters to shrimp
pond effluent has been trialed (Wang & Jakob, 1991; Hopkins et al., 1993a; Jakob et al.,
1993; Lin et al., 1993; Funge -Smith & Briggs, in prep.), but not widely adopted on a
commercial scale due primarily to the problems with fouling from the high concentrations of
particulates (Huguenin, 1976; Funge -Smith & Briggs, 1998). The present study is one in a
series designed to remedy this problem through a better understanding of the effects of the
stocking densities of biofilters, water flow regimes and integrati ng sedimentation prior to
biofiltration (Jones & Preston, 1999; Chapter 3; Jones et al., 2001b; Chapter 5).
4.4.1 Efficiency of Biofilters
Previous studies have demonstrated that sedimentation, in conjunction with nutrient uptake
by phytoplankton and bacteria, can improve the water quality of effluent from shrimp ponds
(Cripps, 1994). Sedimentation alone will not remove phytoplankton, bacteria or small
inorganic particles from pond effluent (Henderson & Bromage, 1988). This study has
confirmed previous observations that these components can be efficiently filtered from the
effluent by oysters (Plates 4.2 & 4.3) (Lam & Wang, 1989; Hopkins et al., 1993a; Jones &
Preston, 1999; Chapter 3; Jones et al., 2001b; Chapter 5). In addition, the present study
CHAPTER 4 104
has highlighted the need to consider the flow regime adopted, the potential benefits of
recirculation, the size of oysters used, the negative impacts of vertical stocking and the
benefits of pre-settlement of effluent prior to treatment by oysters and macroalgae.
Plate 4.2 Control raceway on the left with no oysters and treatment stocked at “low” density 55 g oysters.
Demonstrates changes in water clarity (reduction in suspended solids) with the oyster tray clearly visible in the
raceway stocked with oysters, but not in the control raceway.
In a previous study, Jones & Preston (1999) (Chapter 3) examined the effectiveness of
oysters in filtering effluent in still water (no flow-through). Comparisons with the results of
the present study indicate that oysters may be more effective in a flow-through system. The
relative efficiency between the two flow regimes can be compared in terms of the percentage
reduction of effluent particulates by the oysters, corrected for the differences in oyster
stocking density. Under a flow-through system (the present study), oysters were
approximately 7, 6, and 1.3 times more effective at removing bacteria, suspended particulates
and chlorophyll a respectively, than under the still water system examined in the previous
study. The relatively minor difference in removal of chlorophyll a between the two studies
may have been due to some settling of phytoplankton in the still water system.
EFFICIENCY AND CONDITION OF BIOFILT ERS 105
High water flow rates can enhance oyster filtration (Loosanoff & Tommers, 1948; Walne,
1972), but low flow rates enable more effective settling of particulates (Piper et al., 1982;
Henderson & Bromage, 1988). The current study is the first to examine the potential of
effluent recirculation through raceways stocked with biofilters in order to maximise the
benefits of both sedimentation and biofiltration.
Plate 4.3 First chamber (foreground) and second chamber (background) of an oyster treatment raceway
showing the improvement in water clarity (reduction in suspended solids) within the raceway.
The results showed that, during the first circuit, sedimentation alone in the control
raceways did not reduce the concentration of suspended organic particulates. However,
there was a reduction in the inorganic particulates, which are not a viable food source
for oysters and hence are most likely to foul oysters (Loosanoff & Tommers, 1948).
Consistent with lack of settlement of organic particulates recorded for the first circuit of
the recirculating experiments, phytoplankton did not settle out of suspension in this
CHAPTER 4 106
experiment, which is in contrast to experiments conducted under still water conditions
(Jones & Preston, 1999; Chapter 3; Jones et al., 2001b; Chapter 5). The water flow
may be keeping non-motile forms in suspension and hence more available to oysters as
a food source. Flowing effluent versus the non-flow regime of other studies (Jones &
Preston, 1999; Chapter 3; Jones et al., 2001b; Chapter 5) may also be responsible for
the increase in the number of particles observed in the control raceways in the present
study. Based on the reduction in the mean particle size in the control, the increase was
due to breaking up of aggregated particles, which is consistent with the observations of
Walne (1972).
Biofiltration by oysters resulted in significant reductions in the concentration of both
organic and inorganic particulates. The production of faeces and psuedofaeces by
oysters was shown to increase the mean particle size in the present study. These
particles can more easily settle out of suspension, and therefore do not contribute to the
particle load leaving the system (Haven & Morales-Alamo, 1970). The increase in the
mean size of suspended particulates in the effluent after oyster filtration has been
observed previously (Tenore & Dunstan, 1973).
The dual processes of sedimentation and oyster filtration result in the removal of nutrients
through the filtration and settlement of particulates and the addition of nutrients by
remineralisation (Blackburn et al., 1988) and oyster excretion (Hammen et al., 1966). In the
experimental conditions of this study there was a net reduction in total nutrients. This net
reduction occurred despite the increase in dissolved nutrients (particularly NH4+) from oyster
excretion (Shpigel et al., 1993b; Jones et al., 2001b; Chapter 5).
EFFICIENCY AND CONDITION OF BIOFILT ERS 107
Most of the material filtered by oysters is deposited as faeces and pseudofaeces, while the rest
is incorporated into the oyster tissue. To remove these nutrients from the system entirely will
require harvesting of the oysters, and either removal of the sediment, bacterial denitrification,
or uptake of the remineralised nutrients by macroalgae (Funge -Smith & Briggs, in prep.),
which is subsequently harvested (Hopkins et al., 1995b).
The macroalgae in the present study were shown to be net contributors of dissolved nutrients
due to senescence from fouling by settling particulates. An increase in nutrients from
Gracilaria spp. incubated in shrimp fa rm effluent was also observed by Funge -Smith &
Briggs (in prep.). Given proper sedimentation, and oyster filtration to sufficiently reduce the
concentration of suspended solids, macroalgae have been shown to rapidly reduce the
concentrati ons of dissolved nutrients from shrimp pond effluent (Jones et al., 2001b;
Chapter 5).
By the end of the raceway experiment there was considerable mortality on the bottom
two oyster trays, presumably due to fouling from the increased sedimentation from the
oysters on the tray above (Plate 4.4). This is consistent with high mortality (71%)
observed in oysters placed directly on the bottom of a commercial shrimp pond
(Hopkins et al., 1993a), although this contrasts with the success of Jakob et al. (1993) in
growing oysters in seven layers in tanks fed with shrimp pond effluent. However, their
design enabled much higher flow rates and the effluent water was from a semi intensive
farm (compared with intensive in the present study) with lower concentrations of
particulates. Other studies which have had success growing oysters in multi layer
systems have used cultured phytoplankton and not shrimp pond effluent (Scura et al.,
1979), which eliminates the high inorganic solids loading. A higher flow regime may
CHAPTER 4 108
help to reduce mortality, although if flow is too high it may inhibit the settling of
particulates, or simply result in insufficient residence time for the oysters to effectively
filter the effluent. This could be overcome by either a) recirculating the effluent through
the oysters several times, with a sedimentation regime in between to facilitate settlement
of the larger particles produced by the oysters or, b) simply stacking the trays only one
layer deep. In commercial ponds without baffles, one layer of oysters may result in a
situation of laminar flow where the oysters only have access to the top portion of the
water column, with the rest flowing underneath, thereby reducing filtration performance
(Scura et al., 1979).
Plate 4.4 Fouling of oysters by settling particulates in raceways.
Kinne et al. (1997) found that reducing flow rates (i.e., longer retention time of effluent in the
raceways) increased the efficiency of oysters to improve the quality of the effluent water. An
alternative is recirculation of the effluent through the oysters several times. The second
option would probably be more efficient, as reducing the flow rate is likely to cause a buildup
of metabolites and a decrease in dissolved oxygen (Thielker, 1981). Slowing the flow rate
down in a system with such a high particulate load may increase fouling. A slower flow rate
may also fail to distribute the food evenly amongst the oysters (Scura et al., 1979), and may
EFFICIENCY AND CONDITION OF BIOFILT ERS 109
slow down the filtration rate of the oysters, which is positively correlated to water flow rate
(Walne, 1972). However, Haven & Morales-Alamo (1970) found that slowing the flow rate
increased the total removal of particulates, presumably due to giving the oysters more time to
filter the particles. A recirculating treatment system can facilitate a high effluent flow rate to
enhance the oyster filtration rate, while increasing total particulate removal by maintaining a
long total residence time by recirculation of the effluent through the oysters more than once.
Dissolved nutrient uptake by macroalgae is also related to flow rate (Wheeler, 1980) and
water exchange (Ugarte & Santelices, 1992), as well as light availability (Hanisak, 1979;
Falkowski, 1983). Consequently nutrient removal by macroalgae may be more efficient in a
faster flowing recirculating system stocked at lower densities to improve water flow and
reduce the boundary layer (Wheeler, 1980).
There will be cost – benefit issues to consider with increased pumping costs involved in
recirculating flow. This is the first study to use recirculating treatment with oysters as
biofilters, and clearly the improvements in water quality were considerably greater compared
with continuous flow. However, the increased costs would have to be weighed up against the
improvements to water quality and the income from enhanced oyster production. Simply
optimising single flow through by optimising flow rates, stocking densities, aeration regimes,
and pre settlement with baffles may prove to be more economically viable.
Sedimentation generally removes larger particles with a high specific gravity, leaving
smaller and motile particles in suspension. For the American oyster (Crassostrea
virginica), a particle size range of 3-4 µm was the most efficiently removed, with
removal of 1-2 µm particles only half as efficient (Haven & Morales-Alamo, 1970).
CHAPTER 4 110
Particles <5 µm were shown to be optimal for ingestion by Saccostrea commercialis
(Thielker, 1981). Greater than 95% of the particles remaining in suspension in shrimp
pond effluent in the present study were in the optimal size range (<5 µm).
High numbers of particulates have been shown to reduce feeding efficiency in oysters
(Loosanoff & Tommers, 1948; Wisely & Reid, 1978). For S. commercialis, less than
2 mg L-1 was optimal, with no production of pseudofaeces. At 18 mg L-1, 50% of the
oysters were producing pseudofaeces, and at 35 mg L-1 all oysters were producing both
faeces and pseudofaeces (Wisely & Reid, 1978).
The lower efficiency of smaller 35 g oysters used in the recirculating experiments is
expected due to the differences in pumping rates by different sized oysters (Wang,
1990). Estimates based on Wang (1990) suggest that the 35 g oysters would be
pumping around 22 L d-1 and the larger 55 g oysters, 64 L d-1. Based on these pumping
rates and the oyster stocking densities, the 55 g oysters filtered the entire volume of
effluent in the raceway in the 2 h period taken for full water exchange. However, the
35 g oysters would have only filtered 20% of the total effluent in the raceway after the
first circuit, and 80% after four circuits. Species differences, changes in water flow rate
and the impacts of high concentration of food and other particulates (Loosanoff &
Tommers, 1948; Ali, 1970) on the filtration rate of oysters may affect these calculations
(Walne, 1972).
4.4.2 Condition of Biofilter Organisms
The high concentrations of suspended solids in shrimp pond effluent are due to faecal
matter, uneaten food, phytoplankton (Hopkins et al., 1994), and scouring the bottom
EFFICIENCY AND CONDITION OF BIOFILT ERS 111
sediments of earthen ponds by the action of aerators (Boyd, 1992). These suspended
particulates can have considerable impacts on the growth and condition of oysters and
macroalgae and their ability to reduce the phytoplankton, bacterial and nutrient
concentrations of the effluent (Hopkins et al., 1993a; Funge -Smith & Briggs, 1998;
Funge -Smith & Briggs, in prep.). To improve the filtration efficiency and the condition
and growth of the biofilters, sedimentation could be carried out. The benefits of
sedimentation to the improvements in water quality that can be obtained by biofilters
has been shown previously (Wang, 1990; Jones et al., 2001b; Chapter 5).
The results from the present study indicate that there may be an optimal amount of
settling with regard to maintaining long term oyster and macroalgal condition and
growth. Oyster volume showed the greatest increase in the 6 h settling treatment, which
may be a result of increased food availability, or even the presence of medium
concentrations of clay particles which have been shown to increase the growth rate of
oysters (Huntington & Miller, 1989). Despite a slight increase in volume in the 6 h
treatment, there was no observed increase in total wet weight (meat plus shell). A
prolonged lack of food can initiate shell growth rather than meat gain (Brown &
Hartwick, 1988). Given previous growth rates recorded for oysters in shrimp effluent
(0.04 g to 55 g in 4 months) (Jakob et al., 1993), the weekly exchange of effluent in the
present study was probably not sufficient to maintain food supply.
The macroalgae also demonstrated a maximum growth response in the 6 h settling
treatment, possibly as a result of increased nutrients from remineralisation of the
remaining particulates (Blackburn et al., 1988). Settling the effluent for 6 h appeared to
be “optimal” for growth of oysters and macroalgae in the present study. This time is
CHAPTER 4 112
likely to vary considerably as a result of differences in effluent composition (seasonal
and overall farm differences), settlement pond design, water flow regime through the
biofilters, and biofilter stocking density, and therefore must be investigated fully and
adapted for the specific application.
The reduced light availability in the 6 h settling treatment may have also helped to
reduce the effects of photoinhibition which were probably present in the 24 h and the
24 h + oyster filtered effluent treatments. The light availability to these treatments
(~200 µE m-2 s-1) was the same irradiance shown to limit growth of Gracilaria
verrucosa (Engledow & Bolton, 1992). A light intensity of 100 µE m-2 s-1 was the
optimal light intensity for maximal growth (Engledow & Bolton, 1992). Light levels as
low as 50 µE m-2 s-1 can saturate growth of Gracilaria and the onset of necrosis can
start as low as 65 µE m-2 s-1 (Bird et al., 1979). Prior to the start of the present study the
macroalgae were showing signs of necrosis due to high light. Incident irradiance was
reduced and the macroalgal condition improved, indicating possible photoinhibition
even at these relatively low irradiances. Gracilaria requires relatively low light, but is
inhibited by the settling of particles onto the mucilage on the thalli (Funge-Smith &
Briggs, in prep.).
Photosynthetic pigments (chlorophyll a and phycoerythrin) of the macroalgae were
elevated in response to high nutrients in the 24 h settled plus oyster filtration treatment,
and in response to low light in the 0 h treatment. Increases in pigments in Gracilaria
have previously been positively correlated with increased nutrient availability (Jones et
al., 1996), and negatively correlated to light availability (Friedlander & Levy, 1995).
EFFICIENCY AND CONDITION OF BIOFILT ERS 113
The photosynthetic response between the seawater control and the settled shrimp effluent
plus oyster filtered treatment may relate to an increase in pigment concentration, or an
increase in the photosynthetic efficiency of the pigments in the macroalgae. The observed
increases in electron transport rate and reduction in photoinhibition is consistent with the
responses to controlled nutrient additions observed in other species of macroalgae (Jones &
Dennison, 1998; Appendix 2). The use of rapid light curves to determine electron trans port
rate under a variety of light intensities provides information about the photosynthetic capacity
and condition of the plant, rather than realistic determination of light saturation intensity.
This technique has been shown to identify changes in plant condition in response to a variety
of environmental stresses including light limitation and inhibition, and nutrient availability
(Hanelt et al., 1994; Herrmann et al., 1995; Cunningham et al., 1996; Hader et al., 1996;
Jones & Dennison, 1998; Appendix 2). The increase in photosynthetic capacity of the
macroalgae incubated in settled and oyster filtered shrimp effluent indicates the potential for
increased growth rates and nutrient uptake.
Tissue N was elevated in the treatment with the highest turbidity (0 h) and, based on the
nutrient excretion observed by Jones et al. (2001b) (Chapter 5), presumably the
highest nutrient availability (24 h plus oyster filtration). Tissue N content of Gracilaria
has been correlated to nutrient availability (Jones et al., 1996) and the concentration of
photosynthetic pigments, especially phycoerythrin (Horrocks et al., 1995; Jones et al.,
1996). The high δ15N in the 0 h treatment may be a response to higher concentrations of
bacteria associated with the high particulate load (Hoch et al., 1994). The considerable
depression in δ15N in the 24 h plus oyster filtration treatment may be a result of higher
discrimination for 14N during nutrient uptake, facilitated by higher nutrient availability
(McClelland & Valiela, 1998).
CHAPTER 4 114
The total concentration of free amino acids was also related to changes in light and nutrient
availability. In addition to being related to increased nutrient availability (Di Martino Rigano
et al., 1992; Jones et al., 1996), high concentrations of free amino acids have been correlated
with reduced light availability (Vergara & Niell, 1995). Specific changes in the amino acid
composition may also be related to increased nutrient availability (Jones et al., 1996), or light
limitation. The percentage of citrulline in Gracilaria has been positively correlated with high
NH4+ availability (Jones et al., 1996).
The interplay between light and nutrient availability had significant impacts on the
physiology of the macroalgae. Some of these changes may be interpreted as indicators of
condition and potential for growth. Both oysters and macroalgae are marketable crops which,
as well as improving the quality of effluent released from aquaculture facilities, may be
additional sources of income which require no further inputs of nutrients or feed to maintain a
viable crop. In addition to direct sale as food on Asian markets, there are several other direct
uses for macroalgae including use in pelleted feeds for shrimp (Briggs & Funge-Smith, in
review), and as fresh feed for abalone, a rapidly expanding aquaculture species (Oakes &
Ponte, 1996). The use of high nutrient content macroalgae such as Gracilaria has been
shown to improve growth in abalone (Fleming, 1995). Extraction of agar and carrageenin
from red macroalgae is probably the most economically viable use at present (Indergaard &
Rueness, 1990). Light and nutrients can have considerable impact on agar and carrageenin
production. Increased dissolved nutrients has been shown to reduce the content of agar and
carrageenin in macroalgae, but increase the gel strength which is important for viability in
commercial processing (Bird et al., 1981).
EFFICIENCY AND CONDITION OF BIOFILT ERS 115
4.4.3 Conclusion
Polyculture (culture of several organisms in the one culture unit) and integrated culture,
which is the culture of several species in discrete culture units (Chien & Tsai, 1985), are
perhaps more ecologically sound methods of aquaculture (Mackay & Lodge, 1983), with the
most efficient use of resources and with the highest resilience against environmental
fluctuation (Chien & Liao, 1995). Management of a polyculture or integrated aquaculture
facility can be more complex with respect to stocking densities, culture techniques and
associated infrastructure, harvesting procedures, and effluent flow management (Chien &
Liao, 1995). Studies like this, however, can provide key information about the interactions
between the effluent and the secondary biofilter crops, and between the biofilters themselves.
This information can be used to better design commercial scale facilities. In particular, the
successful integration of shrimp, oysters and macroalgae in a commercially viable operation
will require proper physical treatment of the effluent prior to biological filtration. It has been
demonstrated in the present study that pre-settlement of particulates can improve the growth
and condition of both oysters and macroalgae incubated in shrimp pond effluent.
Properly designed and managed sedimentation, oyster filtration of particulates (with an
optimised flow regime and appropriate oyster stocking densities), and macroalgal absorption
of dissolved nutrients has the potential to effectively improve effluent water quality and
produce two additional crops from otherwise wasted nutrients.
CHAPTER 5
IMPROVEMENTS IN WATER QUALITY OF AQUACULTURE EFFLUENT
AFTER TREATMENT BY SEDIMENTATION, OYSTER FILTRATION AND
MACROALGAL ABSORPTION
Abstract
Effluent water from shrimp ponds typically contains elevated concentrations of dissolved nutrients and
suspended particulates, compared to influent water. Attempts to improve effluent water quality using filter
feeding bivalves and macroalgae to reduce nutrients have previously been hampered by the high concentration
of clay particles typically found in untreated pond effluent. These particles inhibit feeding in bivalves and
reduce photosynthesis in macroalgae by increasing effluent turbidity. The effectiveness of a three stage effluent
treatment system was investigated. In the first stage, reduction in particle concentration occurred through
natural sedimentation. In the second stage, filtration by the Sydney rock oyster, Saccostrea commercialis
(Iredale and Roughley) further reduced the concentration of suspended particulates, including inorganic
particles, phytoplankton, bacteria, and their associated nutrients. In the final stage, the macroalga Gracilaria
edulis (Gmelin) Silva absorbed dissolved nutrients. Pond effluent was collected from a commercial shrimp
farm, taken to an indoor culture facility and left to settle for 24 h. Subsamples of water were then transferred
into laboratory tanks stocked with oysters and maintained for 24 h, and then transferred to tanks containing
macroalgae for another 24 h. Total suspended solids, chlorophyll a, total nitrogen (N), total phosphorus (P),
NH4+, NO3
-, and PO43-, and bacterial numbers were compared before and after each treatment at: 0 h (initial); 24
h (after sedimentation); 48 h (after oyster filtration); 72 h (after macroalgal absorption). The combined effect of
the sequential treatments resulted in significant reductions in the concentrations of all parameters measured.
High rates of nutrient regeneration were observed in the control tanks, which did not contain oysters or
macroalgae. Conversely, significant reductions in nutrients and suspended particulates after sedimentation and
biological treatment were observed. Overall improvements in water quality (as a percentage of the initial
concentration) were as follows: total suspended solids (12%); total N (28%); total P (14%); NH4+(76%); NO3
-
(30%); PO43- (35%); bacteria (30%) and chlorophyll a (0.7%). These results suggest that sequential treatment of
pond effluent by natural sedimentation, oyster filtration and macroalgal nutrient absorption could significantly
improve shrimp farm effluent water quality.
CHAPTER 5 118
5.1 Introduction
Quantitative comparisons of shrimp farm influent and effluent water have demonstrated that
the effluent can contain elevated concentrations of dissolved nutrients, phytoplankton,
bacteria and other suspended organic and inorganic solids (Ziemann et al., 1992). The
potential adverse environmental impacts from untreated effluent have raised concerns about
the sustainability of shrimp farming (Phillips et al., 1993; Primavera, 1994). This has
prompted the search for cost effective methods of improving effluent water quality prior to
discharge into receiving waters.
Traditional methods of wastewater (sewage) treatment are ineffective and prohibitively
expensive for application in treating shrimp pond effluent (Hopkins et al., 1995b). A
potentially viable alternative is biological treatment of the effluent using oysters and
macroalgae to remove suspended particulates and nutrients (Shpigel et al., 1993b). The
organic component of pond effluent can provide a rich source of food for bivalves (Newell &
Jordan, 1983). Bivalves, such as oysters, can also facilitate the removal of fine inorganic
matter from suspension by coagulating this material into larger, more settleable particles and
egesting them as pseudofaeces (Tenore & Dunstan, 1973).
Previous studies have shown that filtration by oysters can significantly reduce the
concentrations of bacteria, phytoplankton, total nitrogen (N) and total phosphorus (P), and
other suspended particles in shrimp pond effluent (Jones & Preston, 1999; Chapter 3).
However, if sediment loads are too high, oyster filtration can be reduced or cease completely
(Loosanoff & Tommers, 1948). Other studies have observed the problems associated with a
high concentration of suspended solids on the health of oysters (Hopkins et al., 1993a).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 119
These studies indicate the need to reduce the concentration of suspended particles by
sedimentation prior to filtration by oysters.
Regular water exchange in shrimp ponds is required to maintain adequate water quality for
optimal growth conditions. In particular, the toxic effects of high ammonia concentration on
the shrimp (Chien, 1992) can be a critical factor determining water exchange rates (Wajsbrot
et al., 1989; Hopkins et al., 1993a). Although oysters can reduce the concentration of some
particulate and dissolved nutrients (Manahan et al., 1982; Dame, 1996), they can also
increase the concentration of NH4+ through excretion (Hammen et al., 1966). Of the N
absorbed by oysters, approximately 10% goes to growth, 10% to gametes, 2% to byssal
threads, with 50% lost as biodeposition and 27% as excretion (Dame, 1996).
Various species of macroalgae can rapidly assimilate large quantities of dissolved organic
and inorganic nutrients, usually with a preference for NH4+ (D'Elia & DeBoer, 1978; Haines
& Wheeler, 1978; Hanisak & Harlin, 1978; Harlin, 1978; Topinka, 1978; Ryther et al., 1981).
Rhodophyta (red macroalgae) are particularly efficient at taking up nutrients rapidly and have
mechanisms for storing large reserves of nutrients (Vergara et al., 1993). For example, the
red macroalga Gracilaria edulis rapidly assimilates NH4+ (Jones et al., 1996), and another red
macroalgae Kappaphycus alvarezii has been effectively used to assimilate waste nitrogen
from the pearl oysters Pinctada martensi (Qian et al., 1996).
In addition to improving the water quality of shrimp effluent water, macroalgae and oysters
can provide an additional source of income for shrimp farmers. For example, trials into tank
culture of Gracilaria chilensis supplied with salmon seawater effluent have demonstrated
production rates of four times those in wild beds and up to double the agar content
CHAPTER 5 120
(Retamales et al., 1994). In the absence of waste nutrients from salmon cages, shrimp ponds
or other high value aquaculture species, the costs of providing sufficient levels of nutrients
for intensive culture of relatively low value species such as oysters and macroalgae are a
constraint to profitability (Neish, 1979).
This study used controlled laboratory experiments to test the combined efficiency of a three
stage pond effluent treatment system. The water quality after sedimentation, filtration of
particulates by oysters (Saccostrea commercialis) and absorption of nutrients by macroalgae
(Gracilaria edulis) was assessed.
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 121
5.2 Materials and Methods
5.2.1 Experimental Design
Shrimp pond effluent was collected from a commercial Penaeus japonicus shrimp farm in
Moreton Bay, Australia. The farm had a total of 8 ha of ponds and at the time of sampling,
the shrimp biomass in the ponds was approximately 3 t ha-1. A 60 L water sample was
collected from the effluent channel and transported to the laboratory in a plastic drum (35 cm
diameter × 60 cm height). In the laboratory, any settled particles in the sample were
resuspended and 3 replicate 1 L samples were collected for analysis of total suspended solids
(and percent organic of particulates), chlorophyll a, bacterial concentration, total Kjeldahl
nitrogen (TKN) and total phosphorus (TP), and dissolved N (NH4+ & NO3
-/NO2-), and
dissolved P (PO43-). A suite of physico-chemical parameters (temperature, pH, salinity,
dissolved oxygen, and turbidity) were determined with a Horiba U-10 water quality meter
(California, U.S.A.). 30 mL samples of water were collected in test tubes, placed in the dark
for 30 minutes and subsequently placed into a Turner Designs Fluorometer 10-005
(Sunyvale, California, U.S.A.) for determination of in vivo fluorescence (as a measure of
chlorophyll a).
The effluent was left to settle in the dark for 24 hours in the collection drum, with physico-
chemical parameters, in vivo fluorescence, TKN and TP measured at 0.25 h, 0.5 h, 1 h, 2 h,
3 h, 6 h, 12 h, and 24 h. The remaining parameters, dissolved N (NH4+ & NO3
-/NO2-),
dissolved P (PO43-), bacterial concentration, chlorophyll a and total suspended solids were
analysed after 24 h only. A sediment trap (23 cm × 15 cm) was placed on the bottom of the
drum to determine the amount of settled material per litre (after 24 h) and the percent organic
content of the settled particles.
CHAPTER 5 122
Experimental tanks (Fig. 5.1; Plate 5.1) were maintained at 20 - 23°C, and exposed to light
on a 12:12 h light / dark cycle using daylight fluorescent tubes which provided approximately
250 µmol quanta m-2 s-1. After 24 h, 10 L of the settled sample was drained into each of
four, 11 L aerated tanks (26 cm × 17 cm × 25 cm), one as a control (with dead oyster shells),
and three for replicate oyster filtration treatments. The 11 L tanks were each stocked with
16 oysters with a mean wet weight of 40 g. Dead oyster shells were used in the control tanks
to negate any effects on water flow and consequent differences in settling rates. Oysters were
placed on plastic mesh, which was situated on top of the sediment trap. Physico-chemical
parameters, in vivo fluorescence, TKN and TP were measured at 24.25 h, 24.5 h, 25 h, 26 h,
27 h, 30 h, 36 h, and 48 h. The remaining parameters, dissolved N (NH4+ & NO3
-/NO2-),
dissolved P (PO43-), bacterial concentration, chlorophyll a, total suspended solids and
sediment traps were analysed after 48 h only.
After 24 h of oyster filtration, 5 L from each tank was drained into each of four more 11 L
aerated tanks, with one as a control (empty), and three as replicate macroalgal absorption
treatments (100 g wet weight of macroalgae per 5 L) (Fig. 5.1). The water from the oyster
control tank was transferred into the macroalgal control tank, and the water from the replicate
oyster treatment tanks was transferred into the corresponding macroalgal treatment tanks
(Fig. 5.1). Both the oyster and macroalgal control tanks acted as controls for the combined
biological treatment process using both the oysters and macroalgae. Consequently, the
effluent in the macroalgal control tank contained higher concentrations of phytoplankton,
bacteria and other suspended solids, because it had not been treated by the oysters. Physico-
chemical parameters, dissolved N (NH4+ & NO3
-/NO2-), and dissolved P (PO4
3-) were
measured at 48.25 h, 48.5 h, 49 h, 50 h, 51 h, 54 h, 60 h, and 72 h. The remaining
parameters, total suspended solids, TKN, TP, chlorophyll a, bacterial concentration and
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 123
sediment traps were analysed after 72 h only. At the end of each 24 h period, 3 replicate 1 L
samples were collected from each tank and filtered for chlorophyll a extraction, total
suspended solids calculation and the percent organic. Subsamples were taken for TKN and
TP, dissolved N (NH4+ & NO3
-/NO2-), dissolved P (PO4
3-), and bacterial concentration.
At all time periods excepting 0 h, 24 h, 48 h and 72 h, chlorophyll a and TSS were
determined as in vivo fluorescence and turbidity (NTU), respectively. Preliminary analysis
determined correlations between in vivo fluorescence and chlorophyll a, and between
turbidity (NTU) and total suspended solids to be r2 = 0.86 and 0.9, respectively. In vivo
fluorescence measurements were converted to chlorophyll a concentration and turbidity
measurements were converted to total suspended solids for those periods of additional
sampling at 15 mins, 30 mins, 1 h, 2 h, 3 h, 6 h, 12 h.
5.2.2 Analytical Procedures
Chlorophyll a was determined by filtering a known volume of water sample through
Whatman GF/F filters, which were immediately frozen. Acetone extraction and calculation
of chlorophyll a concentration was performed using the methods of Clesceri et al. (1989), and
Parsons et al. (1984).
The filtrate collected from filtering for chlorophyll a analysis was collected in 120 mL
polycarbonate containers and immediately frozen. NH4+ and NO3
-/NO2- and PO4
3- were
determined by the Queensland Health Analytical Services Laboratory (NATA- accredited) in
accordance with the methods of Clesceri et al. (1989) using a Skalar autoanalyser (Norcross,
Georgia, U.S.A.).
CHAPTER 5 124
0 72 48 24 Time (h)
Oysters (10 L)(tanks aerated)
Macroalgae (5 L)(tanks aerated)
RawEffluent Control
(no algae)
100gMacroalgae
100gMacroalgae
100gMacroalgae
16 Oysters
16 Oysters
Control (16 oyster
shells)
16 Oysters
Sedimentation (60 L)(tank not aerated)
Figure 5.1 Design of integrated treatment system stocked with oysters (40 g Saccostrea commercialis), and
macroalgae (Gracilaria edulis).
Plate 5.1 Experimental setup with sedimentation drum (background) and control, oyster, macroalgal filtration
tanks (foreground).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 125
Unfiltered samples for nutrient analysis (total Kjeldahl nitrogen and total phosphorus) were
collected in 120 mL polycarbonate immediately frozen. They were subsequently analysed
within two weeks by the NATA accredited Queensland Health Analytical Services
Laboratory in accordance with the methods of Clesceri et al. (1989) using a Skalar
autoanalyser (Norcross, Georgia, U.S.A.).
Total suspended solids concentrations were determined using the methods of
Clesceri et al. (1989). A known volume of water was filtered onto a pre-weighed and pre-
dried (110 ºC; 24 h) Whatman GF/C glass fibre filter. The filter was then oven dried at 60 ºC
for 24 h and total suspended solids calculated by comparing the initial and final weights.
Volatile suspended solids were determined as loss on ignition by combusting samples in a
muffle furnace for 12 h at 525 ºC (Clesceri et al., 1989).
Bacteria samples were preserved with 2% formalin and kept at 4°C until analysis. A known
volume (0.5 mL - 1 mL) of sample was stained with acridine orange, filtered onto a stained
(Irgalan Black) 2 µm poretics filter and mounted on a slide. Bacteria were counted using
epifluorescence microscopy (Hobbie, 1977).
CHAPTER 5 126
5.3 Results
5.3.1 Suspended Solids
The initial concentration of total suspended solids was 0.60 g L-1. After sedimentation, the
concentration was reduced to 0.17 g L-1. Oyster filtration further decreased the concentration
to 0.02 g L-1. There was no significant change in the oyster control, indicating that material
with a specific gravity greater than water had already settled out of suspension during the
24 h sedimentation period. There was no significant change in the concentration of total
suspended solids in the macroalgal control and treatments (Fig. 5.2A; Table 5.1). During
sedimentation, 1.9 g of material was collected in the sediment traps. In the oyster treatment,
0.42 g was collected, versus 0.02 g in the oyster control. This indicates significant
biodeposition of faeces and pseudofaeces by the oysters (Fig. 5.3).
5.3.2 Organic content
In the first 24 h, sedimentation removed a greater proportion of inorganic particles than
organic particles. The relative organic content of the remaining suspended solids increased
from 23% to 31%. The organic content of the settled material collected in the sediment traps
was 21%, which was lower than the initial water column content of 23%. Sedimentation in
the oyster control tanks did not significantly change the organic content. Oyster filtration
was more effective at removing organic material (phytoplankton, bacteria and detritus),
decreasing the relative organic content of the suspended solids from 31% to 24%. The
organic content of the settled sediment at the end of the oyster treatment was 31%, compared
to 21% at the end of the sedimentation treatment (Fig. 5.4B). These changes mirrored the
changes in the organic content of the suspended particles (Fig. 5.4A). In the macroalgal
treatment and control, the organic content of the suspended solids remained unchanged
(Fig. 5.4A).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 127
Settling Oysters Macroalgae
Control (no oystersor macroalgae)
Oysters Macroalgae
205
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
TS
S (
g L
-1)
0
50
100
150
200
250
300
0 10 20 30 40 50 60 70 80
Time (h)
Ch
loro
ph
yll
a (
ug
L-1
)
Figure 5.2 Changes in total suspended solids (A) and phytoplankton biomass (chlorophyll a) (B) from
sedimentation, oyster filtration and macroalgal absorption. Standard error bars have been plotted, but are too
small to be visible.
A
B
CHAPTER 5 128
Table 5.1 Percentage of original concentrations of various water quality parameters after settling, filtration by
oysters and filtration by macroalgae. * p ≤ 0.05; ** p ≤ 0.01; *** p ≤ 0.001. Percentage of highest concentration
represents the final concentration as a percentage of the highest recorded concentration after sedimentation and
oyster filtration. The percent of initial concentration represents the final concentration as a percentage of the
initial concentration in the untreated effluent. The only differences between the two values are for the dissolved
inorganic nutrients (NH4+, NO3
-, & PO43-).
Percent of highest
concentration
F Value Percent of initial
concentration
F Value
TSS (g L-1 ) 12 1338*** 12 1338***
TSS % Organic (g L-1) 12 507*** 12 507***
TKN (µM) 28 609*** 28 609***
TP (µM) 14 4629*** 14 4629***
NH4+ (µM) 2.3 2987*** 76 0.69
NO3- (µM) 2.2 1176*** 30 30**
PO43- (µM) 4.8 162*** 35 19.6*
Bacteria (no. per L) 30 107*** 30 107***
Chl a (µg L-1) 0.7 202*** 0.7 202***
5.3.3 Chlorophyll a
Sedimentation reduced the chlorophyll a concentration from 180 µg L-1 to 130 µg L-1. In the
oyster control, the concentration of chlorophyll a was further reduced from 130 µg L-1 to
100 µg L-1 due to continued settling. The reduction in the chlorophyll a concentration from
130 µg L-1 to 11 µg L-1 in the oyster treatment was significantly greater than the control. The
concentration of chlorophyll a decreased from 100 µg L-1 to 51 µg L-1 in the macroalgal
control (no macroalgae) and from 11 µg L-1 to 1.5 µg L-1 in the macroalgal treatment. These
decreases indicate additional settling of phytoplankton (Fig. 5.2B; Table 5.1).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 129
0
0.05
0.1
0.15
0 10 20 30 40 50 60 70 80
Time (h)
Set
tled
par
ticl
es (
g L
-1)
Settling Oysters Macroalgae
Control (no oystersor macroalgae)
Oysters Macroalgae
Figure 5.3 Concentration of particles settled per litre from sedimentation and oyster filtration.
5.3.4 Bacteria
During sedimentation, bacterial numbers did not change significantly, whereas the oyster
treatment significantly reduced the concentration of bacteria from 19 × 1010 L-1 to
6 × 1010 L-1. There was no significant change in bacterial numbers in the macroalgal controls
and treatments (Fig. 5.5; Table 5.1).
CHAPTER 5 130
15
20
25
30
35
40
0 10 20 30 40 50 60 70 80
Time (h)
Set
tled
pa
rtic
ula
tes
(% O
rga
nic
)Settling Oysters Macroalgae
1 0 0
2 05
Control (no oystersor macroalgae)
Oysters Macroalgae
15
20
25
30
35
40
TS
S (
% O
rga
nic
)
Figure 5.4 Changes in the organic content of the a) total suspended solids (TSS) and b) settled particles in the
effluent water from sedimentation, oyster filtration and macroalgal absorption. Standard error bars have been
plotted, but are too small to be visible.
A
B
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 131
0
5
10
15
20
25
0 20 40 60 80
Time (h)
Ba
cter
ia (
no
. ×
10
10 L
-1)
Control (no oysters
or macroalgae)Oysters Macroalgae
Settling Oysters Macroalgae
Figure 5.5 Changes in bacterial numbers from sedimentation, oyster filtration and macroalgal absorption.
5.3.5 Dissolved Oxygen
Dissolved oxygen was 6.3 mg L-1 immediately following the initial resuspension of any
settled particles prior to sedimentation. After 24 h of sedimentation without stirring or
aeration, respiration by bacteria and plankton reduced the concentration to 2.6 mg L-1.
Within 2 h of aeration being applied, the concentration of dissolved oxygen was up to
6.3 mg L-1. The concentration peaked at between 7 mg L-1 and 8 mg L-1 in the macroalgal
treatment and control tanks (Fig. 5.6).
CHAPTER 5 132
0
1
2
3
4
5
6
7
8
9
0 20 40 60 80
Time (h)
Dis
solv
ed O
xy
gen
(m
g L
-1)
Control (no oysters
or macroalgae)Oysters Macroalgae
Settling Oysters Macroalgae
Figure 5.6 Changes in water column dissolved oxygen concentrations from sedimentation, oyster filtration and
macroalgal absorption. Standard error bars have been plotted, but are too small to be visible.
5.3.6 Total Nitrogen
The sedimentation treatment reduced the concentration of total Kjeldahl nitrogen (TKN) from
290 µM to 205 µM. The oyster treatment further reduced the concentration to 138 µM, while
in the control the concentration increased to 266 µM. The macroalgal treatment reduced the
concentration from 138 µM to 81 µM, while in the control the concentration increased from
266 µM to 279 µM (Fig. 5.7A; Table 5.1).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 133
0
50
1 0 0
1 5 0
2 0 0
2 5 0
3 0 0
3 5 0
To
tal
Nit
rog
en (
µM
)
0
5
10
15
20
25
0 10 20 30 40 50 60 70 80
Time (h)
To
tal
Ph
osp
ho
rus
(µM
)
Settling Oysters Macroalgae
Control (no oysters
or macroalgae)Oysters Macroalgae
Figure 5.7 Changes in water column total N (A) and P (B) concentrations from sedimentation, oyster filtration
and macroalgal absorption. Standard error bars have been plotted, but are too small to be visible.
5.3.7 Total Phosphorus
Total phosphorus (TP) also decreased during sedimentation from 21 µM to 9.7 µM. The
oyster control increased the concentration to 12.5 µM, whereas the oyster treatment reduced
the concentration to 6.1 µM. The macroalgal control reduced the concentration from
A
B
CHAPTER 5 134
12.5 µM to 11.5 µM and the macroalgal treatment reduced the concentration from 6.1 µM to
2.9 µM (Fig. 5.7B; Table 5.1).
5.3.8 Ammonium
The concentration of water column NH4+ increased significantly during sedimentation from
1.7 µM to 18 µM due to regeneration. Regeneration rates were calculated as µmol h-1 based
on the effluent volumes used during each treatment period (i.e., 60 L for sedimentation, 10 L
for the oyster treatment, and 5 L for the macroalgal treatment). Thus, the observed rate of
regeneration during sedimentation was 40.5 µmol h-1, although this may have been
substantially higher if the loss due to volatilisation and uptake by bacteria and phytoplankton
was taken into account. The rate of NH4+ regeneration in the oyster control was considerably
lower (5.7 µmol h-1) increasing from 18 µM to 29 µM. This was probably a result of the
decrease in the concentration of TSS, thereby reducing the source of remineralisation. The
NH4+ concentration in the oyster treatment increased from 18 µM to 51 µM. During the
macroalgal treatment, the NH4+ concentration was reduced from 51 µM to 1.3 µM (to 1.0 µM
after 2 h). The uptake rate during the first hour (when N concentrations were still saturating
uptake kinetics) was 225 µmol h-1 for the 100 g of macroalgae. Despite the rapid removal of
N from the effluent, this uptake rate is substantially lower than recorded literature values for
other species of Gracilaria (Friedlander & Dawes, 1985; Peckol et al., 1994), suggesting that
removal rates could be improved. The concentration of NH4+ in the macroalgal control
decreased from 29 µM to 1.2 µM, but at a much slower rate (only to 26 µM after 2 h) than in
the macroalgal treatment (Fig. 5.8A; Table 5.1; Table 5.2). The reduction in NH4+ in the
macroalgal control suggests volatilisation, nitrification, or uptake by the high concentration
of bacteria and phytoplankton.
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 135
0
10
20
30
40
50
60A
mm
on
ium
(µ
M)
0
0.5
1
1.5
2
2.5
3
3.5
4
0 10 20 30 40 50 60 70 80
Time (h)
Ph
osp
ha
te (
µM
)
0
2
4
6
8
10
12
14
16
Nit
rate
/ N
itri
te (
µM
)Settling Oysters Macroalgae
Control (no oysters
or macroalgae)Oysters Macroalgae
Figure 5.8 Changes in water column NH4+, NO3
-, PO43- concentrations from sedimentation, oyster filtration and
macroalgal absorption. Standard error bars have been plotted, but are too small to be visible.
A
B
C
CHAPTER 5 136
Table 5.2 Nutrient uptake and release rates for sedimentation, oyster filtration and macroalgal absorption.
Negative symbols represent nutrient uptake, and positive represent nutrient release. The top value for each
treatment is the gross value, the middle value is the control and the bottom value (in bold type) is the net value
after correction for nutrient changes in the control tanks. The last row of results represent the rates of
macroalgal nutrient uptake over the first hour, when nutrient concentrations were still saturating uptake kinetics.
Treatment TKN
µmol h-1
TP
µmol h-1
NH4+
µmol h-1
NO3-
µmol h-1
PO43-
µmol h-1
Sedimentation
Gross
-220
-28
+40.5
+0.05
+0.18
Oysters
Gross
Control
Net
-28
+26
-54
-1.5
+1.15
-2.65
+14
+5.7
+8.3
+5.0
+0.6
+4.4
+1.2
+0.6
+0.6
Macroalgae
Gross
Control
Net
-12
+2.5
-14.5
-0.65
-0.2
-0.45
-10.4
-5.8
-4.6
-2.8
0.0
-2.8
-0.66
-0.26
-0.4
Macroalgae over 1 h
Gross
Control
Net
nd
nd
nd
nd
nd
nd
-225
-7
-218
-10
+1
-11
-3.4
-2.1
-1.3
5.3.9 Nitrate / Nitrite
The concentration of NO3-/ NO2
- was unchanged during the sedimentation period, where the
effluent was not aerated, but increased from 1.0 µM to 1.4 µM in the oyster control and from
1.0 µM to 13 µM in the oyster treatment. There was no significant change in the NO3-/NO2
-
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 137
concentration in the macroalgal control, but the macroalgae in the treatment tanks reduced the
concentration from 13 µM to 0.3 µM (Fig. 5.8B; Table 5.1).
5.3.10 Phosphate
Phosphate was also unchanged during sedimentation, but increased from 0.5 µM to 2 µM in
the oyster control and from 0.5 µM to 3.3 µM in the oyster treatment. The concentration in
the macroalgal control decreased from 2 µM to 0.7 µM, and in the macroalgal treatment the
concentration of P decreased from 3.3 µM to 0.16 µM (Fig. 5.8C; Table 5.1).
These reductions in PO43- equate to a removal rate of 0.66 µmol h-1 in the macroalgal
treatment, while loss in the control was only 0.26 µmol h-1. During the first hour of
macroalgal filtration, the concentration of PO43- in the control decreased at a greater rate than
in the macroalgal treatment, presumably due to phytoplankton and bacterial uptake. This
may have been due to the presence of significantly more phytoplankton and bacteria in the
control tanks, because the control water had not been treated by the oysters.
5.3.11 Nutrient Uptake Rates and Ratios
During the macroalgal treatment, NH4+ was taken up in the first 2 h (hour 48 to hour 50),
NO3- in the next 4 h (from hour 50 to hour 54), and PO4
3- over 10 h (from hour 50 to
hour 60). N in both forms was taken up at faster rates than PO43-.
The TKN: TP ratio in the water column increased during sedimentation from 6.4 to 9.6, and
then during oyster filtration from 9.6 to 10.2, probably due to adsorption of P to settling
particulates. There was no significant change in the oyster control, but the macroalgal control
CHAPTER 5 138
showed some further increase to 10.9, and the macroalgal treatment increased to 12.6
(Fig. 5.9A).
The ratio of DIN: DIP increased from 2.6 to 16.2 during sedimentation, predominantly as a
result of NH4+ remineralisation. During the oyster treatment and control, the ratio dropped to
8.7 and 7.5, respectively. The macroalgal control increased the ratio to 18.8 after 6 h and
then decreased to 2.3, whereas the macroalgal treatment decreased the ratio to 0.8 after 6 h
and then increased it to 4.5. These responses coincided with the rapid drop in the PO43- in the
macroalgal control in the first six hours, and the rapid uptake of NH4+ in the macroalgal
treatment (Fig. 5.9B).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 139
0
5
10
15
20
0 10 20 30 40 50 60 70 80
Time (h)
DIN
:DIP
Ra
tio
Settling Oysters Macroalgae
1 0 0205
Control (no oysters
or macroalgae)Oysters Macroalgae
5
7
9
11
13T
KN
:TP
Ra
tio
Figure 5.9 Changes in water column total N: P ratio (A) and DIN: DIP ratio (B) from sedimentation, oyster
filtration and macroalgal absorption. Standard error bars have been plotted, but are too small to be visible.
A
B
CHAPTER 5 140
5.4 Discussion
Previous studies on the use of sedimentation, bivalves and macroalgae to improve effluent
water quality have generally examined each component separately. For example, Shpigel et
al. (1993b) examined three separate regimes for treating fish pond effluent and concluded
these could be more closely integrated. The present study advanced the concept to the point
of laboratory tests to examine the combined efficiency of sedimentation, biofiltration by
oysters and nutrient uptake by macroalgae.
The ability of a biofilter system to improve the quality of shrimp effluent is likely to vary
depending on the initial water quality. The water quality of shrimp farm effluent depends on
a number of factors, including pond soil type, quality of influent water, stage of growout
season and management practices employed (Ziemann et al., 1992). The present study was
conducted in November, early in the growout season when the concentration of NH4+ was
relatively low (1.7 µM), compared to later in the growout season (65 µM) (Jones et al.,
2001a; Chapter 2). However, there were very high concentrations of total suspended solids
and phytoplankton in the effluent at the time of the study. The concentrations of total
suspended solids (0.6 g L-1), chlorophyll a (180 µg L-1) and bacteria (20 × 1010 L-1), were
respectively 4.5, 5 and 10 times the values measured from the same shrimp farm in the
previous growout season (Jones & Preston, 1999; Chapter 3).
5.4.1 Sedimentation
Sedimentation of shrimp pond effluent was very effective at reducing the total suspended
solid (TSS) load. The concentration was reduced to 12% of the initial concentration with
proportionally more inorganic particles being removed from suspension. Additional benefits
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 141
were reductions in total Kjeldahl nitrogen (TKN) to 70%, total phosphorus (TP) to 47%,
chlorophyll a to 72%, and bacteria to 95% of the initial concentration. The results indicate
that the majority of N in the effluent was associated with phytoplankton and bacteria, rather
than being bound to readily settleable inorganic particles, or detritus. In contrast to this, a
relatively high proportion of the TP appeared to be associated with the readily settleable
particles.
In the present study, the rates of TKN removal by sedimentation and oyster filtration were not
significantly different, however the rate of TSS removal was much slower during oyster
filtration compared with sedimentation. This may be a result of higher N content in the small
(<5 µm) unsettleable particles (Cripps, 1992) removed by the oysters. Alternatively, it may
simply reflect relatively slow settlement rates of phytoplankton compared to TSS. Despite
the reduction in TKN and TP during sedimentation, there was a significant increase in the
concentration of NH4+ in the water column, most likely due to mineralisation of particulate
organic matter (Briggs & Funge-Smith, 1993).
Treatment for the removal of suspended solids must be adapted specifically for the target
effluent. The size and density of the particles, along with the flow rates, surface area, and
retention time affects the efficiency of sedimentation (Chien & Liao, 1995). Effluent with
high algal concentrations may require less sedimentation time, because it can be filtered more
effectively by oysters.
5.4.2 Oyster Filtration
Pre-sedimentation of suspended particles is highly beneficial to oyster filtration. In
particular, the relative increase in the proportion of organic particles (from 23% to 31%) may
CHAPTER 5 142
enhance the oysters’ ability to assimilate particulate nutrients, as well as filter and egest the
sma ller inorganic clay particles as pseudofaeces (Loosanoff, 1949). In the present study, the
concentration of total suspended solids was reduced from 0.6 g L-1 to 0.17 g L-1 after
sedimentation. However, even at this reduced concentration of suspended solids the oyster
pumping rates may have been inhibited. For example, a concentration of 0.1 g L-1 has been
observed to reduce the pumping rate of oysters by up to 87% (Loosanoff & Tommers, 1948).
Ninety five percent of the suspended particulates in the effluent from this shrimp farm were
2 - 4 µm (Jones et al., 2002; Chapter 4). During filtration oysters sort particles by size,
weight (Yonge, 1926) and chemical composition (Loosanoff, 1949; Menzel, 1955). Oysters
preferentially ingest organic material, reject inorganic material, and preferentially ingest N
rich over C rich particles (Newell & Jordan, 1983). Rejected material is prevented from
ingestion by the gills and palps, where it is accumulated and then expelled through the
inhalent opening as pseudofaeces (Barnes, 1994). When food concentration exceeds the
digestive capabilities of the gut, pseudofaeces may contain some digestible food.
Despite rejecting small inorganic particles, oysters may facilitate removal of these particles
from suspension by coagulating them into larger, more settleable particles before egestion as
pseudofaeces. In addition to the fine inorganic particles, oysters filter small organic rich
particles, such as bacteria and very small phytoplankton that would otherwise stay in
suspension (Loosanoff, 1949; Kautsky & Evans, 1987). Oysters can filter particles as small
as 1 µm, although their efficiency in removing the smaller particles (1-3 µm) is two thirds
lower than for larger particles (Haven & Morales-Alamo, 1970).
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 143
The oysters removed large concentrations of phytoplankton, bacteria and other suspended
solids from the water column, and produced faeces and pseudofaeces at a rate of 78 mg (g dry
wt)-1 d-1. This rate is within the range of 66 to 246 mg (g dry wt)-1 d-1 observed for the
Pacific oyster, Crassostrea gigas (Boucher & Boucher-Rodoni, 1988).
Decreases in TP were observed in the oyster treatment due to removal of particulate P
(bacteria, phytoplankton and clay particles with adsorbed P) at a rate greater than the inputs
of PO43-. Shrimp pond effluent often contains a high proportion of small clay particles
(Hopkins et al., 1995b) rich in phosphate (Pomeroy et al., 1965). Bacteria may compete with
sediments for phosphate thereby maintaining the phosphate in the water column (Pomeroy et
al., 1965). The increase in PO43- and TP observed in the oyster control was probably due to
sediment release and the greater increase in PO43- in the oyster treatment due to excretion by
oysters (Asmus et al., 1995). Most of the P filtered by oysters is converted to biodeposits,
with 8% being released as PO43- (Dame et al., 1989), and only 3% absorbed and converted
into biomass (Sornin et al., 1986).
In the present study, oyster filtration reduced the concentration of TKN to 28% and TP to
14% of the initial effluent concentration. These reductions represent an uptake ratio of 20: 1.
Based on the Redfield ratio of 16: 1 in phytoplankton, the oysters are removing more N than
expected. However, this is in contrast to the results of Dame et al. (1989) who observed an
N: P uptake ratio by oysters (Crassostrea gigas) of 2: 1.
During oyster filtration, there was a net decrease in the concentration of particulate nutrients,
and a net increase in the concentration of dissolved inorganic nutrients. In addition to
particulate matter, oysters can also take up dissolved organic and inorganic nutrients and
CHAPTER 5 144
dissolved organic matter (DOM) such as dissolved free amino acids from the water column
(Manahan et al., 1982; Dame, 1996). In particular, oysters can assimilate PO43- directly from
the water column for carbohydrate metabolism, energy transfer and shell deposition
(Pomeroy & Haskin, 1954).
A net NH4+ excretion rate of 0.9 µmol g-1 DW h-1, which is in the range described in the
literature (Boucher & Boucher-Rodoni, 1988), was measured in the oyster treatment tanks.
This takes into account inputs from regeneration, but does not incorporate the losses to
phytoplankton and bacterial uptake of NH4+, volatilisation and nitrification which may
significantly enhance these rates (Boucher & Boucher-Rodoni, 1988).
The low concentrations of dissolved oxygen (DO) during the first two hours of oyster
filtration are likely to have reduced the filtration efficiency. Ingestion rates of bivalves are
reduced under conditions of low DO availability (Sobral & Widdows, 1997). The oysters
were maintained in no flow environments, apart from the water motion due to aeration. Flow
rates have been positively correlated with oyster filtration rates (Walne, 1972), and so oysters
in treatment ponds with flowing water might produce significantly greater improvements in
water quality.
5.4.3 Macroalgal Absorption
Gracilaria edulis in the present study achieved 87% of its total NH4+ uptake within the first
hour. A similar rate (80%) was observed for the red macroalga Kappaphycus alvarezii,
incubated in the nutrients excreted by the pearl oyster Pinctada martensi (Qian et al., 1996).
The initial rate of NH4+ uptake in the present study was 218 µmol h-1 per 100 g wet
-1
(Table 5.2), which is very close to the rate observed by Qian et al. (1996) for 200 g wet-1 for
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 145
K. alvarezii, and is within the range observed for Gracilaria (Rivers & Peckol, 1995) and
other species of macroalgae (Taylor et al., 1998). The NO3- uptake rates in the present study,
however, were considerably lower, with a maximum rate of 15% per hour compared to 66%
per hour for the K. alvarezii. Low light and high concentrations of NH4+ (D'Elia & DeBoer,
1978) can inhibit NO3- uptake. The ability to assimilate NO3
- can also be species specific.
Consistent with the findings for G. edulis, uptake of NO3- by Ulva lactuca was found to be
minimal (Krom et al., 1995). Preferential uptake of NH4+ before NO3
- is typical for most
species of macroalgae and is consistent with the results of Oki & Fushimi (1992) who found
that as the concentration of NH4+ decreased, the uptake rate of NO3
- increased.
In the present study, the fastest rates of net uptake of NH4+, NO3
- and PO43- by macroalgae
were 26, 2.5 and 2.2 times higher than the net oyster release rate, respectively. The net rate
takes into account losses to phytoplankton and bacterial uptake, volatilisation and nitrification
(Table 5.2). The proportion of NH4+ to NO3
- taken up by the macroalgae varied, being 20: 1
during the period of fastest uptake of NH4+ uptake (in the first hour), but only 1.6: 1 over the
entire 24 hours. This overall rate is comparable to the oyster release rates of NH4+ and NO3
-,
which were 1.9: 1. The uptake of NH4+ and NO3
- by K. alvarezii was in similar proportions
to the release rates of the pearl oyster (Qian et al., 1996).
Macroalgae did not significantly change the NH4+: NO3
-/NO2- ratio, but the uptake of N
markedly increased the DIN: DIP on a short time scale, due to the saturating availability of N,
and the ability of Gracilaria edulis to store luxury reserves of N for times of nutrient
limitation (Gerloff & Krombholz, 1966; Jones et al., 1996). This reduction in the DIN: DIP
ratio may have implications for water recirculation and downstream environmental impacts.
CHAPTER 5 146
Analysis of bioindicators downstream of shrimp farms has indicated responses typical of high
NH4+ rich effluent (Jones et al., 2001a; Chapter 2).
The nutrient removal efficiency of macroalgae may be improved with increased light,
particularly for nitrate uptake (Hanisak, 1979; Falkowski, 1983), increased water flow to
reduce the boundary layer (Wheeler, 1980), and higher stocking densities (Ugarte &
Santelices, 1992). The present study was conducted indoors under experimental conditions.
Under full solar radiation, photosynthetic rates could be considerably higher, and nutrient
uptake rates correspondingly higher. In a biological treatment pond, water flow rates would
be determined by the physical volume of water being pumped through the pond, and the
water movement due to the action of paddlewheels or other aerators. In the present study, the
macroalgae was stocked at 2.3 kg m-2, whereas maximum growth rates of Gracilaria
chilensis were achieved by stocking at 4 kg m-2 in winter and 8 kg m-2 in summer (Ugarte &
Santelices, 1992).
Given the low concentrations of NH4+ in the effluent from the shrimp farm at the time of the
study (early in the shrimp growout season), the total reduction in the concentration of NH4+
was lower than expected. The 24% reduction in NH4+ from 1.7 µM to 1.3 µM may not be
indicative of the real potential for this system to remove NH4+. Due to addition of NH4
+ to
the system through the decay of epiphytes and macroalgal tissue, a concentration of
approximately 1 µM may be the equilibrium point between uptake and input from decay.
Based on the fast rates of uptake observed in this study, it can be expected that during the late
phase of the growout season, when the concentrations of NH4+ are considerably higher (up to
60 µM) (Jones et al., 2001a; Chapter 2), that overall removal efficiency will be
considerably higher.
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 147
The percent reductions by macroalgal nutrient absorption have been expressed as overall
reductions from the initial concentration in the shrimp effluent. However, because
remineralisation and oyster excretion significantly increased the concentration of NH4+, NO3
-,
and PO43-, the efficiency of nutrient removal can be expressed as a percentage of the
concentration in the water received by the macroalgae. Based on these calculations, the
concentration of NH4+ was reduced to 2.3%, NO3
- to 2.2% and PO43- to 4.8% (Table 5.1).
5.4.4 Nutrient Regeneration
During sedimentation, the rate of NH4+ regeneration was 1.8 mmol m-2 d-1, although this did
not take into account loss of NH4+ by phytoplankton and bacterial uptake, or by volatilisation.
The NH4+ regeneration rate declined markedly after transfer to the aerated oyster control
tanks. This may have been due to the reduction in particulates, or the result of increased
aeration. These results are consistent with those of Kamiyama et al. (1997), who observed
increases in NH4+ release during low dissolved oxygen conditions, and NO3
- release under
high dissolved oxygen conditions. The NH4+ regeneration rates during sedimentation were
significantly lower than the sediment NH4+ release rates of between 8.4 mmol m-2 d-1 and
18 mmol m-2 d-1 that have been recorded in fish ponds (Riise & Roos, 1997). In the present
study, the NH4+ uptake rate by the macroalgae was 118 mmol m-2 d-1. At this rate, the
macroalgae would be able to remove up to 6.5 times the NH4+ released from the fish ponds
observed by Riise & Roos (1997).
During sedimentation when the concentration of dissolved oxygen was very low, there was
no production of NO3-, despite significant amounts of NH4
+ remineralisation. This is typical
of anaerobic aquaculture ponds (Blackburn et al., 1988). Once under aerobic conditions in
CHAPTER 5 148
the aerated oyster controls, nitrification was observed, although rates were still very low. In
the oyster treatment tanks, there was significant production of NO3-, which is consistent with
the nutrient release by the pearl oyster Pinctada martensi (Qian et al., 1996).
Oysters excrete NH4+, amino acids, urea, uric acid (Hammen et al., 1966) and PO4
3-
(Pomeroy & Haskin, 1954; Dame et al., 1989; Dame, 1996). The nitrogen release rate by the
pearl oyster for NH4+ was 0.52 µmol h-1 and for NO3
-, 0.44 µmol h-1, per oyster (Qian et al.,
1996), compared to 0.52 µmol h-1 and 0.28 µmol h-1, respectively, for S. commercialis in the
present study. It has been suggested that the observed NO3-/NO2
- release from oysters is due
to nitrifying bacteria in the digestive tract of the organism (Saijo & Mitamura, 1971; Boucher
& Boucher-Rodoni, 1988). There could also be free living nitrifying bacteria in the water
column, although there is only negligible production of NO3- in the oyster and macroalgal
control tanks.
The apparent enhancement of nitrification by the oysters resulted in more NO3- being
available for denitrification. By increasing the proportion of NO3--N relative to NH4
+-N,
bivalves such as oysters can enhance nitrification / denitrification coupling, promoting greater
removal of N from the effluent. Rates of denitrification in mussel beds have been observed
to be 21% greater than bare sediment (Kasper et al., 1985), and combined with the mussel
harvest, this represented a 68% greater loss of nitrogen than bare sediment.
Bivalves are known to significantly increase the rates of carbon deposition (Doering &
Oviatt, 1986; Doering et al., 1986) and the rates of benthic flux of DIN (Doering et al.,
1987). Consequently, the sediment NH4+ and organic N pools under bivalves are
significantly higher than those in bare sediments (Kasper et al., 1985). Bivalves enhance
COMBINED TREATMENT USING SEDIMENTATION, OYSTERS AND MACROALGAE 149
movement of organic nitrogen to the sediments, where it decomposes. This decomposition in
the aerobic surficial sediments yields NH4+ (remineralisation) and NO3
-/NO2- (nitrification)
and, in the deeper anaerobic sediments, NO3- is converted into N2 gas (denitrification) (Dame,
1996). Stimulation of nitrification by oysters can result in more NO3- present in the water,
and less NH4+ (Boucher & Boucher-Rodoni, 1988).
By stimulating nitrification, the oysters significantly reduced the NH4+: NO3
-/NO2- ratio from
13: 1 to 4: 1. This has significant management implications, especially in relation to the
possibilities for recycling of treatment pond effluent back into the production ponds.
Aeration and uptake by phytoplankton and bacteria in the macroalgal control tanks reduced
the NH4+: NO3
- ratio to 0.5: 1. Although the ratio of NH4+: NO3
- was lower in the control, the
total concentration of DIN was significantly lower in the macroalgal treatment.
5.4.5 Conclusions
Combining the oysters and macroalgae in the same treatment pond may produce greater
improvements in water quality. Rapid uptake of the nutrients resulting from remineralisation
of oyster biodeposits (Kelly & Nixon, 1984), may reduce potential stimulation of
phytoplankton production. There may be a balance between nutrient enhanced phytoplankton
productivity and the removal of phytoplankton biomass by oysters (Zeitzschel, 1980).
For co-culture of macroalgae and oysters to be successful, the growth requirements of each
species need to be optimised. In particular, if the temperature, water flow rate, or light
availability are not adequate for macroalgal growth there may be more biomass decaying than
being produced. This would result in a decline in water quality and consequently lower
oyster growth rates (Qian et al., 1996).
CHAPTER 5 150
In order to maximise the volume of effluent that this polyculture system can successfully
treat, the maximum final concentration of the various water quality parameters needs to be
determined. Reductions in all parameters were observed to be non linear over 24 h, and
therefore the amount of treatment time would need to be adjusted to maximise the volume of
effluent that can be treated, without sacrificing the required water quality. Using this
combination of polyculture, it was estimated that up to 18 kg N ha-1 d-1 and 15 kg P ha-1 d-1
could be removed.
The results from the present study indicate the ability of sedimentation, oysters and
macroalgae to substantially improve the quality of shrimp pond effluent being released into
the receiving waters (Plate 5.2). It may also make it possible to recirculate the effluent back
into the production ponds, thereby reducing the need for water exchange.
Plate 5.2 Water samples collected: a) before sedimentation; b) after sedimentation; and c) after biofiltration.
A B C
CHAPTER 6
CONCLUSION
The impacts of shrimp pond effluent on the receiving waters were investigated using
bioindicators (Chapter 2), and the potential for biological treatment to improve the water
quality of effluent was investigated (Chapters 3, 4, & 5). It was shown that shrimp pond
effluent can be effectively treated by a combination of sedimentation, oyster filtration of
particulates and macroalgal absorption of dissolved nutrients (Chapter 5).
6. 1 Downstream Impacts
Bioindicators proved to be useful markers for identifying the region of influence of shrimp
effluent discharged into receiving waters. Results indicated that effluent nutrients may be
more widespread than can be detected by traditional water quality sampling techniques
(Chapter 2). Water quality analyses could detect no differences beyond 750 m from the
mouths of the creeks receiving the discharges, but bioindicator responses detected effluent
nutrients up to 4 km away. The amino acid composition, tissue nitrogen content and stable
isotope ratio of nitrogen (δ15N) in seagrasses, mangroves and macroalgae were responsive to
nutrient inputs from shrimp and sewage effluent discharged into two adjacent creeks.
Different responses in these biological indicators revealed that the impacts of shrimp pond
effluent were qualitatively different to impacts of treated sewage effluent, and were spatially
more extensive than identified by water quality analyses (Chapter 2).
Effective in-pond management can significantly reduce effluent loadings of sediment and
nutrients. Remediation techniques such as settlement, wetlands filtering, biological filtering
with bivalves and seaweeds, and bacterial inoculation, all have potential as cost-effective
CHAPTER 6 152
treatment methods for shrimp effluent (Samocha & Lawrence, 1997). Although there are
some lessons to be learnt from research into the treatment of sewage, shrimp farm effluent
has different characteristics, in particular the comparatively high concentrations of suspended
inorganic particulates and high salinity. This thesis presents information on some of the
differences between the shrimp pond effluent and sewage effluent obtained from a
comparative study in Moreton Bay. Shrimp pond effluent was shown to be high in
chlorophyll a and TSS, and high in NH4+, whereas sewage effluent was relatively low in
chlorophyll a, and TSS, and high in NO3- (Chapter 2).
6.2 Efficiency of Biological Filters
Filtration of effluent by oysters significantly reduced the concentrations of chlorophyll a
(phytoplankton), bacteria, total nitrogen, total phosphorus and total suspended solids. In
particular, oyster filtration significantly reduced the concentration of small organic and
inorganic particles (<5 µm) which would otherwise remain in suspension. Despite reducing
the concentrations of particulates and total nutrients, oyster excretion increased the
concentrations of dissolved nutrients (ammonium, nitrate / nitrite, and phosphate).
Macroalgae effectively reduced the concentrations of dissolved nutrients from effluent that
had been pre-filtered by oysters. Macroalgae can take up dissolved organic matter such as
free amino acids, and nutrients such as urea, both of which may be present in shrimp ponds
(Hanisak, 1983). Red macroalgae such as Gracilaria are known to have high uptake rates
and capacity for storage of excess nutrients and they contain high concentrations of
commercially important substances such as agar and carrageenin (Hanisak, 1983).
Macroalgae can also be used as a food supply for animals, such as the abalone Haliotis sp.
(Sorgeloos & Sweetman, 1993), or livestock (Fielder et al., 1994).
CONCLUSION 153
6.3 Condition of Biofilters
The results of this study demonstrated that, when cultured without prior sedimentation of
particulates, the efficiency and condition of oysters and macroalgae was reduced by fouling
from the high concentration of suspended particulates in the effluent. When oysters were
stacked one tray deep, the highest oyster stocking density was the most effective at improving
effluent water quality. However, higher stocking densities with macroalgae and oysters
stacked three layers deep resulted in the death of oysters and senescence of macroalgae. This
may be mitigated by faster water flow, although this would reduce the residence time and
perhaps reduce the removal efficiency of the oysters. Recirculating the effluent through the
oyster treatment significantly enhanced effluent water quality by increasing residence time
without reducing the flow rate.
Integrated treatment incorporating sedimentation prior to biological filtration with oysters and
macroalgae, improved growth and reduced signs of stress in the oysters and macroalgae. The
ratio of NH4+ to NO3
- / NO2- was also reduced, which has positive implications for recycling
of wastewater back into shrimp production ponds, and reducing impacts on receiving waters.
Settled effluent without subsequent biological filtration demonstrated very high rates of
nutrient regeneration.
Despite significant reductions in total suspended solids by natural sedimentation, the
concentration of particulates was probably still inhibiting the filtering efficiency of the
oysters (Loosanoff & Tommers, 1948). Flocculating agents could be used to reduce settling
times, although their use would probably not be cost effective (Norris 1994 in Samocha &
Lawrence, 1997). However, the use of screen filters (Cripps, 1994) or the incorporation of
baffles into pond design can improve the efficiency of sedimentation, and are likely to be the
CHAPTER 6 154
most economically viable and low maintenance option. The main factors affecting settling
rates are particle size and water flow rate. A maximum flow of 4 m min-1, but preferably
1 m min-1 is needed to achieve optimal sedimentation performance (Henderson & Bromage,
1988).
Rates of biodeposition are related to the concentration of suspended particulates. At high
food concentrations oysters produce large amounts of pseudofaeces (i.e., food filtered but not
ingested) and therefore have high deposition rates, demonstrating an inefficient use of filtered
food (Tenore & Dunstan, 1973). Due to the inefficient use of food and high pseudofaeces
production at high food concentrations an integrated aquaculture system may be more
effective if it also includes bivalves capable of existing in a high silt environment such as
clams, deposit feeders such as polychaete worms and carnivores such as flounder
(Tenore et al., 1973).
Oysters can potentially grow at much greater rates in shrimp pond effluent compared with
natural leases. Growth from spat to market size has been achieved in as little as 4 months
(Jakob et al., 1993). In order to maintain adequate improvements in water quality as well as
produce a reliable crop of oysters, the stocking of a full range of oyster sizes is necessary.
This thesis showed that there are significantly different filtration rates between different sized
oysters and stocking densities (Chapters 3 & 4). Filtration performance was found to be
positively correlated with stocking density, however, Holliday et al. (1991) found that oyster
growth rates were inversely correlated to stocking density, although increasing water flow
rates may help to overcome this. To maintain a continuous supply of mature, marketable
oysters, an aquaculture facility would need to stock a full size range. Estimates of how many
CONCLUSION 155
of each size class and the concomitant improvements in water quality have been discussed
(Wang, 1990).
6.4 Scaling up for Commercial Treatment
There are several factors that may affect the performance of the proposed integrated
aquaculture system when scaled up to treat the effluent from an entire shrimp farm:
a) the effects of wind on the settling rates of particles in a sedimentation pond compared to
the controlled conditions in this study;
b) the effects of water flow on the efficiency of the oysters and macroalgae to take up
nutrients and particulates;
c) the problems associated with ensuring the water is mixed sufficiently to permit optimal
biofiltration;
d) the potential long term effects of fouling by suspended solids and epiphytes on the
condition and efficiency of the biofilters.
The use of a sedimentation pond with high walls may reduce the effects of wind on stirring
the water column. Also a smaller, deeper pond may produce a larger percentage of
undisturbed water and provide less fetch for wind disturbance, although the distance that the
particles have to settle becomes greater. Current velocities in the settling ponds should not
exceed 2-4 cm s-1 to prevent resuspension (Warrer-Hansen, 1982).
In the oyster treatment pond, paddlewheels may be ideal to simulate the high recirculating
flow rates provided by bubble aeration in the present study. High flow rates, good mixing
and aeration are important to reduce the problems of fouling and the buildup of metabolites
CHAPTER 6 156
(Thielker, 1981). These characteristics are also required to reduce boundary layers to
facilitate efficient nutrient uptake by the macroalgae (Wheeler, 1980). Oysters and
macroalgae must remain close to the surface to avoid fouling, and sufficient mixing is
required to ensure the biofilters are exposed to all of the effluent in the pond.
Incorporation of sedimentation and biofilters in the one pond may be problematic because of
differences in the optimal depth for biofilter efficiency versus sedimentation efficiency.
Sedimentation could be separated from the biofilters to enable slow, undisturbed (laminar)
flow for sedimentation, and faster mixed flow for the biofilters. The culture of oysters and
macroalgae in separate ponds (oysters first, and then macroalgae) may help to improve the
condition and efficiency of the macroalgae because of the reduction in suspended solids
effected by the oysters. However, co-culture in the one pond may be more efficient, as
Qian et al. (1996) demonstrated that co-culture of bivalves and macroalgae results in higher
growth rates than monospecific cultures. Improvements in bivalve growth could be due to
removal of metabolites by the macroalgae, while the higher growth rates of macroalgae may
be a result of increased nutrient availability from excretion by the bivalves.
6.5 Other Potential Biofilters
Bivalves other than oysters have a potential application as biofilters. Mussels may be more
effective than oysters due to their higher filtration efficiency (Jørgensen, 1966; Tenore &
Dunstan, 1973), although oysters have a higher assimilative capacity (Tenore et al., 1973).
Clams could also be better because of their ability to survive burial in high concentrations of
silt, although their filtration rates are the lowest of the three bivalves.
CONCLUSION 157
There may also be potential benefits in using macroalgae other than Gracilaria. Despite the
success of using high value species of red macroalga such as Gracilaria sp., there may be
certain species of macroalgae which are better able to tolerate the high suspended solid load
in shrimp farm effluent. Enteromorpha sp. (a filamentous green macroalga) bloomed
prolifically in the experimental raceways during non sampling times, and was observed to
have very fast growth rates, with no signs of fouling.
6.6 Management Implications and Potential Problems with Biofiltration / Polyculture
Although the treatment of shrimp pond effluent through the integration of natural
sedimentation, oyster filtration and macroalgal absorption has been shown to be successful
(Chapter 5), there are still potential problems associated with very high suspended solid loads
(Chapter 4). The concentration of suspended solids in the pond effluent can be affected by
the concentration in the influent water, the pond soil type, pond shape, feed management and
aerator design and implementation. Reducing the concentration of suspended solids may be
best accomplished by changes to in-pond management practices.
Due to the origins of shrimp feed pellets (ground shellfish, crustaceans, and fish), they often
contain relatively high concentrations of heavy metals due to bioaccumulation, resulting in
elevated levels in the shrimp farm effluent. There has been some concern that shrimp pond
effluent may result in elevated concentrations of heavy metals in oysters and macroalgae used
as biofilters, thereby making them unsafe for human consumption. However, oysters
cultured in shrimp pond effluent were found to be free of human pathogens and fit for human
consumption (Shpigel et al., 1993b). Further studies are needed to examine the effects of
shrimp pond effluent on the quality of oysters in relation to national and international food
standards.
CHAPTER 6 158
6.7 Benefits of Polyculture or Integrated Aquaculture
Intensive shrimp farming without water exchange may be possible if feed management
practices are modified so that small portions of feed are delivered at frequent intervals
(Hopkins et al., 1995a), and biofilters such as oysters and macroalgae are incorporated to
reduce the loading of particulates and dissolved nutrients. These techniques not only reduce
the impacts of the effluent on the environment, but they can also have considerable
advantages for maintaining in-pond water quality and preventing the transfer of disease from
contaminated water pumped in after water exchange.
In addition to improving the effluent water quality from shrimp ponds, the culturing of
oysters and macroalgae in this nutrient rich wastewater can lead to rapid growth rates in these
commercially valuable species. The culture of oysters and macroalgae are traditionally
conducted using extensive management practices, in comparison to shrimp, fish and other
intensively farmed aquaculture species. Maintaining food supply and nutrients for intensive
culture of both oysters and macroalgae can be costly and, it has been suggested that for
commercial success, would require development of polyculture with other crops (Neish,
1979).
Polyculture is a more ecologically sound method of aquaculture (Mackay & Lodge, 1983)
with an efficient use of resources, and a high resilience to environmental fluctuation (Chien
& Liao, 1995). In addition to meeting environmental standards to reduce effluent loads of
nutrients and total suspended solids, shrimp farmers may potentially benefit economically
from the production of biofilter crops.
CONCLUSION 159
The concept of polyculture is regarded by many as having potential benefits, environmentally
and economically, but the ideas have not been widely accepted (Wang & Jakob, 1991).
Sedimentation ponds are ideal as primary treatment, with oysters and macroalgae functioning
as secondary and tertiary treatment. Possibly the main factor inhibiting the uptake of this
type of polyculture is the problem associated with culture and maintenance of these
organisms in the pond environment (Chien & Liao, 1995). However, in addition to
significant water quality improvements, high yields of oysters and macroalgae can be
achieved (Shpigel et al., 1993b).
6.8 Future Research
The results of this study demonstrated that natural abundance of stable isotope ratios of
nitrogen were useful indicators for detecting the range of influence of pond effluent nutrients
in the discharge environment. Analysis of the natural abundance of δ13C of seagrass,
mangroves and macroalgae could also be used to determine whether the C was from
decomposing mangrove leaves or excreted from phytoplankton (Primavera, 1996).
Differences in the δ13C have also been correlated to differences in phytoplankton species
composition (Fogel et al., 1992), which may make it possible to trace the δ13C signature back
to a nutrient source with a specific phytoplankton composition. The combined analysis of
two elements also improves the ability to resolve sources with similar isotopic signatures
(Ziemann et al.1984).
To better identify the region of potential impacts from shrimp farm effluent, enriched stable
isotope tracer techniques (e.g., 15N labelled NH4+ released in the farm’s discharge canal)
could be used to quantify effluent nutrient uptake by biota in the receiving waters. Controlled
CHAPTER 6 160
laboratory experiments are also needed in order to discern the relationships between shrimp
and sewage effluent and the effects on amino composition, δ15N and δ13C.
Investigations into the use of other species of bivalves and macroalgae, as well as
incorporation of sediment fauna such as polychaetes could be conducted. The optimal
combination will probably relate to effluent composition, commercial viability and ease and
cost of maintenance. Certain species of bivalves and macroalgae could potentially improve
the quality of effluent from freshwater aquaculture. The use of duckweed (Lemnaceae) has
already been trialed and found to be successful, with very high growth on effluent from carp
and tilapia. It was found to significantly reduce dissolved nutrient concentrations and was
also used as feed for the cultured fish (Skillicorn et al., 1998).
Application of the integrated treatment system (Chapter 5) needs to be tested on a pond scale
to determine the optimal combination of biofilter density, pond depth, aeration and water
flow rate to maximise water quality improvements while maintaining good condition and
growth of the biofilters. These are likely to be a compromise and the optimal setup will be
dependant on the effluent composition, and the species of bivalves and macroalgae used.
6.9 Summary
This thesis has developed techniques for identifying the influence of shrimp pond effluent on
receiving waters using seagrass, macroalgae and mangroves as biological indicators.
Biological indicator responses indicated that the impacts of aquaculture effluent were
qualitatively different to sewage effluent, and were spatially more extensive than identified
by water quality analyses. To reduce these impacts, effluent treatment techniques
incorporating biological filters were investigated.
CONCLUSION 161
Filtration by oysters significantly reduced the concentrations of particulates, but increased the
concentrations of NH4+, NO3
-, and PO43-. Nutrient uptake by macroalgal significantly
reduced the dissolved nutrient concentrations, in particular the concentration of NH4+.
Differences in oyster size, stocking density, and water flow regime had significant impacts on
the condition of the oysters and macroalgae as well as their ability to improve the quality of
the shrimp pond effluent. The efficiency and condition of the oysters and macroalgae was
reduced by high suspended particulate loads in the effluent, however, an integrated system
incorporating natural sedimentation prior to biological filtration proved effective at
optimising oyster and macroalgal performance.
The integrated system effected significant improvements in the water quality of effluent
being released from shrimp ponds. These improvements may be sufficient to enable
recirculation of effluent back into the shrimp production ponds creating a more controlled
system with minimal environmental impacts (Fig. 6.1). Filtration and absorption by various
marine organisms can be effective for monitoring and reducing the environmental impacts of
aquaculture effluent.
CHAPTER 6 162
Figure 6.1 Diagrammatic design of water flow for typical untreated shrimp farms (left) and a design to
incorporate physical (sedimentation) and biological (oyster filtration and macroalgal absorption) treatment
(right).
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APPENDIX 1
FACTORS LIMITING PHYTOPLANKTON BIOMASS IN THE BRISBANE RIVER
AND MORETON BAY
Jones, A.B., Dudley, B.J., & Dennison, W.C. 1998. Factors limiting phytoplankton biomass
in the Brisbane River and Moreton Bay. In: Tibbets, I.R., Hall, N.J., & Dennison, W.C.
(eds.). Moreton Bay and Catchment. School of Marine Science, The University of
Queensland, Brisbane. pp. 179-186.
In: Tibbetts, I.R., Hall, N.J. & Dennison, W.C. eds (1998) Moreton Bay and Catchment.School of Marine Science, The University of Queensland, Brisbane. pp. 301-308.
Factors Limiting PhytoplanktonBiomass in the Brisbane River andMoreton Bay
Adrian B. Jones, Bernard J. Dudley and William C. Dennison
Department of Botany, The University of Queensland, Brisbane Qld 4072
AbstractIncreasing eutrophication of coastal marine environments has led to the development of nutrient samplingprograms to monitor water quality. Various shortcomings of chemical analyses common in the majorityof these sampling programs have identified the need to develop biological indicators (bioindicators) thatcan be used to detect the source, fate and impact of nutrients within eutrophic systems. Using phytoplanktonbioassays, we tested the accepted model that phosphorus (P) limits phytoplankton biomass in freshwaterand nitrogen (N) is limiting in coastal marine waters. The study was conducted along a salinity gradientfrom Lake Wivenhoe (0 ppt), through the nutrient rich, highly turbid Brisbane River to Moreton Bay(35 ppt). Water samples from seven sites along the transect were spiked to make treatments with thefollowing nutrient concentrations: 30 µM NH4
+; 200 µM NO3-; 20 µM PO4
3-; 66 µM SiO32+; all nutrients
combined and an unspiked control. Chlorophyll a fluorescence was measured daily for 7 d in eachtreatment. Nutrient limitation was inferred if there was an increase in chlorophyll a fluorescence innutrient spiked samples relative to the control. Light limitation was inferred from an increase inchlorophyll a fluorescence in controls. Phytoplankton bioassay results indicate that phytoplankton biomasswas limited by: N and P in the freshwater sites; N in the upper reaches of the Brisbane River; light in themiddle reaches; and N and silica (Si) in the lower reaches. Within Moreton Bay, phytoplankton populationsexhibited no response to nutrient addition. These results suggest that the generalised models of nutrientlimitation may not be applicable to all regions. In particular, P may not limit phytoplankton at anysalinity regime within the Moreton region.
IntroductionThe Brisbane River is the largest tidal estuary flowing into Moreton Bay, and is characterisedby high turbidity. It receives large inputs of nutrients from both point and non-point sourcessuch as terrestrial runoff, sewage treatment plants and release from resuspended sediments(Moss, 1990). To assess the role of nutrients and suspended solids in eutrophication, traditionalwater quality analyses involved periodic sampling of parameters such as dissolved inorganicnitrogen (DIN) and phosphorus (DIP), chlorophyll a, and total suspended solids (TSS). Thesetechniques are limited as they only provide an instantaneous measurement at the time of watercollection whereas large fluctuations in the concentrations of dissolved nutrients can occur onshort time scales in estuaries (Wheeler & Björnsäter, 1992; Valiela, 1995). Additionally, theseanalyses do not directly assess the impact of eutrophication on marine life in the system (Lyngby,1990). Bioassays can be used to investigate the nutrient responses of the phytoplanktoncommunity, thereby providing information on the history of nutrient availability at a site.
Some bioassay studies supply all but one nutrient to each treatment containing an axenic cultureof phytoplankton (Smayda, 1974; Hitchcock & Smayda, 1977). If the particular treatmentshows lower growth rates than the treatment with all nutrients added, then that nutrient isconsidered to be limiting. The technique used in the present study has been employed widely(Valiela, 1995) and uses ambient phytoplankton populations spiked with single nutrients. Rapidincreases (before the other ambient nutrients have been assimilated) in phytoplankton biomassin such treatments indicate that the nutrient was limiting for the ambient phytoplanktoncommunity.
Jones, Dudley & Dennison
Moreton Bay and Catchment302
In Moreton Bay, phytoplankton bioassay techniques have been used to assess both short term(~15 hr) physiological responses to nutrient enrichment and long term (up to 7 d) responses inbiomass to nutrient enrichment. Short term bioassays measure CO
2 uptake via 14C after 15 h
incubations in artificially increased nutrient concentrations (O’Donohue & Dennison, 1997).Long term bioassays examine changes in algal biomass with nutrient additions (O’Donohue etal., this volume). Changes are measured as in vivo fluorescence, which can be directly correlatedto chlorophyll a concentration. Interpretations of the nutrient(s) limiting to phytoplankton canbe made based on how rapidly the population increases and to which nutrient(s) they respond.In some cases the population may be adapted to oligotrophic conditions and may not be capableof rapid assimilation of the nutrients. In a system where nutrients are available, but increases inbiomass are limited by the absence of one particular nutrient (most commonly nitrogen), thensupplying this nutrient will result in increases in biomass. When the phytoplankton communityis not limited by either N or P alone; the addition of a combination of N and P may produce amarked response. Light limitation may be inferred from increases in biomass in the controls(no nutrient), after suspended solids settle out of suspension.
This investigation was conducted to identify factors limiting phytoplankton biomass in theBrisbane River estuary and Moreton Bay, and to define more accurately where efforts shouldbe directed to best monitor water quality in the region. Ultimately, this information will benefitdecision making on a number of management issues, particularly nutrient removal strategies.
Materials and Methods
Seven sites were selected along a transect from Lake Wivenhoe, along the Brisbane River andinto Moreton Bay (Figure 1). The transect spans the full salinity range from freshwater (0),through the tidal reaches of the river, to full salinity (35) seawater. Sites will be referred tothroughout the text as distance in kilometres upstream (negative) or distance downstream(positive) from the mouth of the Brisbane River.
Water quality
Chlorophyll a was determined by filtering a known volume of water sample through WhatmanGF/F filters, which were immediately frozen. Acetone extraction and calculation of chlorophylla concentration was performed using the methods of Clesceri et al. (1989) and Parsons et al.(1989).
The water collected from filtering for chlorophyll a analysis was transferred into 120 mLpolycarbonate containers and immediately frozen. NH
4+ and NO
3-/NO
22- were determined within
two weeks of sampling using the methods of Parsons et al. (1989).
The concentration of total suspended solids (TSS) was determined by filtering a known volumeof water onto a pre-weighed and pre-combusted (110ºC; 24 h) Whatman GF/C glass fibrefilter. The filter was then oven dried at 60ºC for 24 h and TSS calculated by comparing theinitial and final weights.
Secchi depth was determined by lowering a 30 cm diameter secchi disk (black and whitealternating quarters) through the water column until it was no longer possible to distinguishbetween the black and white sections.
Limitation of phytoplankton biomass
303Moreton Bay and Catchment
Bioassays
Phytoplankton bioassays were conducted with ambient phytoplankton assemblages collectedfrom seven sites in the Brisbane River and Moreton Bay (Figure 1). One 30 L drum of waterwas collected from each site, kept cool and shaded, and returned to an outdoor incubationfacility. Four litres of water from each site was filtered through a 200 µm mesh (to screen outthe larger zooplankton grazers) into sealed transparent 6 L plastic containers and placed inincubation tanks filled with water (2 m diameter, 0.5 m deep). Temperature was maintained at±2°C of the ambient found at each site by flowing water through the tanks and light levels weremaintained at 50% of incident irradiance with neutral density screening. For each site therewere six bioassay containers, each with a different nutrient treatment. Samples were spiked tomake the following concentrations: NO
3- (200 µM); NH
4+ (30 µM); PO
43-(20 µM); SiO
32+
(66 µM); all nutrients at those concentrations (+All); and a control (no nutrient addition). Theconcentrations were chosen to ensure saturation by each particular nutrient based on the highestambient concentrations at the study sites. In vivo fluorescence was measured for all treatmentsdaily for 7 d, using a Turner Designs Fluorometer.
The potential of light and nutrients to stimulate significant increases in phytoplankton biomass(blooms) in the bioassays was investigated. The nutrient control treatments functioned as lightresponse treatments because sedimentation of suspended solids in the samples increased lightavailability above ambient levels. Light stimulated bloom potential was calculated as thedifference between initial and maximum in vivo fluorescence values in the control water sampleover 7 d. Nutrient stimulated bloom potential was calculated as the difference between the+All nutrients treatment and the control.
Figure 1. Map of study sites in the Brisbane River and Moreton Bay, Queensland, Australia. Distancesare relative to the mouth of the river (negative upriver from the mouth and positive intoMoreton Bay). Site 1 – Lake Wivenhoe (-145 km); Site 2 – Karana Downs (-82 km); Site 3– Bremer River Junction (-72 km); Site 4 – Long Pocket (-36 km); Site 5 – Gateway Bridge(-12 km); Site 6 – South of Fisherman’s Island (+8 km); Site 7 – Myora (+33 km).
N
BrisbaneCity
⊗-145 km
-82 km
-72 km
-36 km
-12 km
+8 km
+33 km
Jones, Dudley & Dennison
Moreton Bay and Catchment304
Results
Water quality
The seven study sites occur along a salinity gradient from freshwater at -145 km to full strengthseawater in the Bay sites. Water column NH
4+ concentration ranged from <1.5 µM at +33 km
to a peak of 11 µM at -12 km. NO3- ranged from <4 µM at +33 km to 112 µM at -36 km, near
the middle of the river. TSS concentration ranged from 4 mg/L at +33 km to 23 mg/L at -36 kmand the chlorophyll a concentration from 0.5 µg/L at +33 km to 12 µg/L at -72 km (Figure 2).
Bioassays
Freshwater sites (-145 km and -82 km) demonstrated responses in phytoplankton biomass inthe +All treatments, with little or no difference in the controls and other nutrient additions. Atthe estuarine sites (-36 km, -12 km, +8 km), phytoplankton biomass increased in all treatments,and the control. At -72 km, the response in the control was not as great as that in the treatments.The response of phytoplankton at +8 km was primarily to nitrogen (NH
4+ and NO
3-) and then
to silica (SiO3). Populations at the oceanic site (+33 km) showed no almost no response
(Figure 3).
DiscussionNutrient (source and concentration) and light availability (from secchi depth measurements)vary considerably along the gradient of the Brisbane River and Moreton Bay (Figure 2). Thepeak in NH
4+ at the -12 km site may be due to a fertiliser plant, located at -7 km (Moss, 1990),
and the Luggage Point sewage treatment plant near the river mouth (0 km) (Moss et al., 1992).The relatively high concentrations of NO
3- (compared to NH
4+) at most of the mid and upper
reaches of the river may result from non-point sources, enhanced nitrification and preferentialuptake of NH
4+ by phytoplankton (Lipschultz et al., 1986).
The response of the bioassays (Figure 3) to the +All nutrients treatment at the -145 km and-82 km sites indicates limitation by more than one nutrient. At the -72 km and -36 km sites,there was no increase in biomass above the control, indicating a response to increased lightdue to sedimentation of particles. At the -12 km site there was also a light response, but N andSi stimulated biomass above the control. The +8 km site had almost no response in the control,but strong responses to N and Si treatments. This indicates that light is no longer a factorcontrolling biomass in this region of the estuary and is consistent with the higher light availability(secchi disk readings) at this site. The +33 km site showed almost no response except for the+All treatment on the seventh day. The responses in the +All treatments at day 7 may be due tobottle effects, as all the other sites responded within the first 4 days. This response suggeststhat the phytoplankton community within the Brisbane River is better adapted to respondrapidly to high nutrient availability compared with those in the low nutrient waters of easternMoreton Bay.
There was an inverse relationship (r2 = 0.93) between nutrient concentration and light availability(as total suspended solids) along the study transect. At some sites phytoplankton biomass waslight limited (tidal estuary), and at others nutrient limited (freshwater and marine). Lightlimitation was observed when phytoplankton biomass from the highly turbid mid-river sitesincreased in controls after the suspended solids in the sample had settled out. This responseindicates that once light limitation was removed, controls had sufficient ambient nutrientconcentrations to grow as well as the treatments, to which nutrients had been added.
Limitation of phytoplankton biomass
305Moreton Bay and Catchment
Bloom potentials, calculated as the difference in maximum biomass (as in vivo fluorescence)between the treatment and control, represent the relative increase in biomass given saturatinglight or nutrient conditions. The upriver sites show the greatest nutrient-induced bloom potentialdue to relatively high light availability coupled with a environment containing relatively highconcentrations of nutrients. The highly turbid midriver sites had the highest light-inducedbloom potential (Figure 4).
Figure 2. Water quality parameters at the seven sites along the Brisbane River/Moreton Bay studytransect from Lake Wivenhoe (-145 km) to Myora in eastern Moreton Bay (+33 km).
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Our new model (Figure 5) describes N and P limitation in freshwater sites, N limitation in theupper reaches of the river, light limitation in the middle reaches, and N and Si limitation in thelower reaches and the Bay. In contradiction to the widely accepted nutrient limitation model,we found no P limitation of phytoplankton biomass in samples from the freshwater sites. Thistrend has been observed elsewhere in southeast Queensland, in the Maroochy and TweedRivers. This departure from the accepted worldwide trend may be explained by a number of
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Limitation of phytoplankton biomass
307Moreton Bay and Catchment
Figure 4. Light stimulated and nutrient stimulated bloom potential along the Brisbane River/MoretonBay study transect.
Figure 5. New and old models of factors limiting phytoplankton biomass in the Brisbane River andMoreton Bay system.
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factors including large inputs from fertilisation, nutrient content of Australian soils and fromthe high loading of suspended solids to which P binds.
These results demonstrate the need to develop a better understanding of the interactions oflight and nutrients as factors limiting phytoplankton biomass in the Brisbane River estuary andMoreton Bay. Phytoplankton bioassays indicate that the established model for limiting factorsmay not apply to this region, and that such assays, in conjunction with traditional water qualitymeasurements, provide key information not available from traditional water quality monitoringprograms.
AcknowledgementsWe would like to thank the Marine Botany practical class (BT 320) for sample collection andanalysis. Andrew Moss (Queensland Department of Environment) and James McEwan(Brisbane River Moreton Bay Wastewater Management Group) provided assistance withinterpretation of results.
ReferencesClesceri, L.S., Greenberg, A.E. & Trussel, R.R. (1989) Standard methods for the examination of water
and wastewater. pp. 253-256. American Public Health Association, New York.
Hitchcock, G.L. & Smayda, T.J. (1977) Bioassay of lower Narragansett Bay waters during the 1972-1973 winter-spring bloom using the diatom Skeletonema costatum. Limnology and Oceanography22: 132-139.
Lipschultz, F., Wofsy, S.C. & Fox, L.E. (1986) Nitrogen metabolism of the eutrophic Delaware Riverecosystem (USA). Limnology and Oceanography 31: 701-716.
Lyngby, J.E. (1990) Monitoring of nutrient availability and limitation using the marine macroalgae,Ceramium rubrum (Huds.) C. Ag. Aquat. Bot. 38: 153-161.
Moss, A.J. (1990) Turbidity and nutrient behaviour in the estuary. In: The Brisbane River. A source-book for the future (eds Davie, P., Stock, E. & Low Choy, D.) pp. 307-311. Australian LittoralSociety Inc. in assn with Queensland Museum, Brisbane. 427 pp.
Moss, A.J., Connell, D.W. & Bycroft, B. (1992) Water quality in Moreton Bay. In: Moreton Bay in theBalance (ed. Crimp, O.N.) pp. 103-114. Australian Littoral Society Inc. and the Australian MarineScience Consortium, Moorooka, Queensland.
O’Donohue, M.J. & Dennison, W.C. (1997) Phytoplankton productivity response to nutrientconcentrations, light availability and temperature along an Australian estuarine gradient. Estuaries20(3): 521-533.
Parsons, T.R., Maita, Y. & Lalli, C.M. (1989) A Manual of Chemical and Biological Methods for SeawaterAnalysis. Oxford, Pergammon Press. 173 pp.
Smayda, T.J. (1974) Bioassay of the growth potential of the surface water of lower Narragansett Bayover an annual cycle using the diatom Thalassiosira pseudonana (oceanic clone, 13-1). Limnol.Oceanog. 19: 889-901.
Valiela, I. (1995) Marine Ecological Processes 2nd edn. New York, Springer Verlag. 686 pp.
Wheeler, P.A. & Björnsäter, B.R. (1992) Seasonal fluctuations in tissue nitrogen, phosphorus and N:Pfor five macroalgal species common to the Pacific northwest coast. J. Phycol. 28: 1-6.
APPENDIX 2
PHOTOSYNTHETIC CAPACITY IN CORAL REEF SYSTEMS: INVESTIGATIONS
INTO ECOLOGICAL APPLICATIONS FOR THE UNDERWATER PAM
FLUOROMETER
Jones, A.B. & Dennison, W.C. 1998. Photosynthetic capacity in coral reef systems:
investigations into ecological applications for the underwater PAM fluorometer. In:
Greenwood, J.G., & Hall, N.J. (eds.). Proceedings of the Australian Coral Reef Society 75th
Anniversary Conference, Heron Island October 1997. School of Marine Science, The
University of Queensland, Brisbane. pp. 105-118.
In: Greenwood, J.G. & Hall, N.J., eds (1998)Proceedings of the Australian Coral Reef Society 75th Anniversary Conference, Heron Island October 1997.School of Marine Science, The University of Queensland, Brisbane. pp. 105-118.
Photosynthetic Capacity in CoralReef Systems: Investigations intoEcological Applications for theUnderwater PAM FluorometerAdrian B. Jones and William C. Dennison
Department of Botany, The University of Queensland, Brisbane Qld 4072
ABSTRACTA submersible pulse amplitude modulated (PAM) fluorometer was used to determine the effects ofdesiccation, ultraviolet radiation, changes in solar radiation and nutrient availability on the photosyntheticapparatus of a variety of marine plants (zooxanthellae, benthic microalgae and macroalgae) at HeronIsland, Great Barrier Reef, Australia. The PAM measures photosynthesis as irradiance-dependentphotosystem II electron transport. There were a number of interspecific and intraspecific variations inelectron transport rate (ETR) based on physiological and morphological differences, and the plant’s responseto changes in environmental conditions. The highest ETR was found in the zooxanthellae of the clamTridacna maxima, and the lowest in the calcified green macroalga Halimeda opuntia. Factors such aswater velocity, ultraviolet radiation, solar radiation (total irradiance and spectral changes), desiccation, nutrientavailability and algal pigment content were hypothesised as influencing intraspecific changes in ETR. Aseries of experimental manipulations were conducted to test these hypotheses. Reef flat algae was shadedto 50% of incident solar radiation and 0% of ultraviolet radiation. Samples of macroalgae were collectedfrom the reef flat and 15 m depth and allowed to desiccate to determine if different populations of the samespecies could adapt physiologically to different environmental conditions. Reef flat samples were collectedand incubated in seawater enriched in nitrogen and phosphorus to test for nutrient limitation. Significantdifferences in the ETR of the plants tested highlighted the impacts of various environmental parameters onphotosynthetic capacity. Samples from regions with higher water velocities on the reef flat had significantlyhigher ETRs. Screening of ultraviolet radiation increased the maximum ETR of certain species, whileprolonged periods of shading reduced the maximum ETR of some species more quickly than others.Desiccation responses were the same between deep collected and reef flat populations, although increasedlight and temperature did reduce the maximum ETR of the deep collected samples. Fertilisation responsesvaried between species. The results indicate that PAM fluorometry can be used as a tool for in situ nondestructive assessment of the effects of various ecological parameters on photosynthetic activity in marineplants.
INTRODUCTIONVarious environmental factors can influence the temperature, nutrient availability, light regimeand cellular water content of marine macroalgae, and symbiotic microalgae. These factorscan be a significant influence on photosynthetic capacity, and in turn on the productivity ofthe entire coral reef system (Dring, 1982; Matta & Chapman, 1995). Photosynthesis in marineplants is traditionally measured as oxygen production. This requires containment of the plantin chambers in the laboratory, or in situ. This method is time-consuming, can be destructiveto the plant and creates an artificial environment which may not sufficiently simulate naturalconditions (Hanelt et al., 1994; Hader et al., 1996b). In contrast, pulse amplitude modulated(PAM) fluorescence enables rapid measurement of photosynthetic responsesnon-destructively in situ with minimal interference to the plant’s immediate environment.
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The development of portable pulse amplitude modulated (PAM) fluorometers, in particularthe submersible DIVINGPAM (Walz GmbH. Effeltrich, Germany), has facilitated many areasof novel research on photosystem physiology, regulation and ecological adaptation in marinemacroalgae and other aquatic plants (Cunningham et al., 1996). PAM fluorescence has beenused successfully to study photosynthesis in corals (Warner et al., 1996), macroalgae (Haneltet al., 1994; Herrmann et al., 1995; Hader et al., 1996a; Hader et al., 1996b), and benthicmicroalgae (Hartig et al., 1998; Kromkamp et al., 1998). These studies have also successfullyused this technique to determine the effects of environmental parameters such as temperature,desiccation, photosynthetically active radiation (PAR) and ultraviolet radiation (UVR) onthe photosynthetic apparatus of the organism.
The PAM fluorometer measures photosynthesis as irradiance-dependent photosystem IIelectron transport. A saturating pulse of light allows measurement of the electron transportrate (ETR) by saturating the plastiquinone pool between PSII and PSI. The PAM measureschlorophyll a fluorescence before and after the saturating pulse allowing subsequentcalculation of the photosynthetic yield. The yield is corrected for the specific leaf absorbanceand then halved to account for the two quanta of light required per electron. This final figure,the electron transport rate, has been successfully correlated to traditional measures ofphotosynthesis, such as changes in oxygen evolution (Xiong et al., 1996) and 14C uptake(Hartig et al., 1998). For a detailed explanation of the mechanisms involved in using PAMfluorescence to measure photosynthesis as electron transport rate, refer to Schreiber et al.(1994).
Short term changes in ambient solar radiation may confound measurements of true ETRwhen using the PAM to make single instantaneous yield calculations (Critchley & Gademann,in prep). Generation of rapid light curves (RLC’s) by the PAM avoids some of these problemsby providing the specimen with periods of increasing actinic light (Critchley & Gademann,in prep), during which the response to various irradiances can be measured. The result is anETR versus PAR curve, which can be used to determine the maximum potential rate ofphotosynthesis and the irradiance at which photoinhibition may occur.
The aim of this research was to investigate potential uses of PAM fluorescence to determinethe impacts of a range of environmental variables, including light availability (PAR and UVR),desiccation stress, water motion, and nutrient availability on the photosynthetic apparatus ina variety of marine plants.
MATERIALS AND METHODSA DIVINGPAM (submersible pulse amplitude modulated fluorometer) was used to determinethe electron transport rate (ETR) in a variety of photosynthetic organisms and their responseto changes in certain environmental parameters at Heron Island, Great Barrier Reef, Australia(Figure 1). The island is characterised by a large reef flat (approximately 2 m deep at hightide) with a wide variety of macroalgae (Rogers, 1997). The presence of beachrock on thewindward side of the island has led to the formation of a gutter just offshore from the beach(Coote, 1984; Kan, 1995). This region is characterised by higher water flow rates, an increaseddensity of benthic microalgae, and a reduced density of corals and attached macroalgae(Rogers, 1997; Heil et al., in review).
Experiments were conducted across the reef flat and down the reef slope out from the southernside of the island. The ETR was determined at a range of light intensities by generation of
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rapid light curves (RLC’s) using the PAM. The RLC was generated over a 90 sec period with10 sec periods of actinic irradiance and 1 sec saturating pulses for measurement of quantumyield (Critchley & Gademann, in prep.). Data was stored in the diving PAM for subsequentdownload to computer. Measurements were standardised by placing the fibre optic cable ina leaf clip, attached to the plant near the growing tip of the thallus. When used with filamentousalgae such as Chlorodesmis fastigiata the filaments were arranged to simulate a flat thallus.When used on corals or clams a specially designed coral clip was used to ensure no movementof the fibre optic cable during the period of the RLC. Three replicate RLC’s were conductedfor each species on neighbouring individuals. In all cases visibly bleached or otherwise stressedplants were avoided, and only unshaded specimens were sampled. For each experiment,measurements were taken at approximately the same time of the day and same phase of thetide to limit the effects of changes in temperature and light history.
An initial survey of ten species from the reef flat was conducted to compare maximum ETR,and to determine an appropriate species for testing the effects of changing environmentalconditions. C. fastigiata was chosen because it is ubiquitous across all of Heron reef andtherefore is capable of adapting to a variety of habitats.
The ETR of C. fastigiata was measured along a transect conducted at 20 m intervals fromthe beach at Heron Island, through the gutter, across the reef flat to the reef crest (Figure 1).Another transect was conducted down the reef slope, with ETR of C. fastigiata measured at2, 5, 10 and 15 m depths (Figure 1).
Further responses in ETR to changes in light availability were determined by shading a numberof macroalgal species in situ on the reef flat to 50% of incident solar radiation and 0% ofUVR using 1 m2 quadrats of shade cloth and polycarbonate sheeting respectively. Threereplicate shade and three replicate UV screens were constructed and placed approximately10 cm above the macroalgae. ETR was measured on afternoon low tides everyday for 4 d.
Sampling Transect
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Figure 1. Location of study site at the southern reef flat of Heron Island, Great Barrier Reef, Australia.
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Adaptation between populations of the same species to desiccation, high light and temperaturestress was determined by collecting C. fastigiata from 15 m and from the reef flat. Sampleswere maintained in flow-through aquaria (water is pumped from the reef flat) for 1h prior todesiccation to allow the plants to adapt to the same light, temperature and water flow regime.The samples were then desiccated in the shade for one hour, with RLCs performed at 0 mins,15 mins, 30 mins and 60 mins.
The effect of nutrient enrichment on the maximum ETR of macroalgae (C. fastigiata,Padina tenuis and Colpomenia sinuosa) and coral (Acropora sp.) was determined. Sampleswere collected from the reef flat and placed in six flow-through aquaria. Three replicateaquaria were fertilised (88g m-2 N, 22g m-2 P) with slow release Osmocote fertiliser andthree unfertilised control treatments were maintained. Three replicate samples of each specieswere maintained in each aquarium. Samples were allowed to acclimate and respond to nutrientadditions for 10 days prior to sampling ETR.
RESULTSOf all the photosynthetic plants surveyed with the PAM, the zooxanthellae in the clam Tridacnamaxima had the highest maximum ETR (230 µmol e - m-2 s-1), considerably higher than thosein the coral Acropora sp. (70 µmol e - m-2 s-1). Chnoospora implexa (150 µmol e - m-2 s-1) andColpomenia sinuosa (125 µmol e - m-2 s-1) from the Phaeophyta had the highest rates of themacroalgae, followed by Plocamium microcladioides (95 µmol e - m-2 s-1) (Rhodophyta) andChlorodesmis fastigiata (60 µmol e- m-2 s-1), Caulerpa racemosa (40 µmol e- m-2 s-1), andHalimeda opuntia (15 µmol e- m-2 s -1) (Chlorophyta). The maximum ETR of benthicmicroalgae was 45 µmol e- m-2 s-1.
The ETR of C. fastigiata exhibited peaks in the gutter (43 µmol e- m-2 s-1) and at the reefcrest (45 µmol e - m-2 s-1), corresponding to the regions of fastest water flow on the reef flat(Heil et al., in review). The ETR of samples across the rest of the reef flat were fairlyconsistent, ranging from 31 µmol e- m-2 s-1 to 33 µmol e- m-2 s-1 (Figure 2). The ETRs of thesamples taken from depth were higher than those from the reef flat, increasing to a peak of70 µmol e- m-2 s-1 at 10m depth, and dropping to 56 µmol e- m-2 s-1 at 15m (Figure 3).
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The samples of C. fastigiata collected from 15 m for the desiccation experiment weremaintained in aquaria prior to desiccation. After one hour in flow-through aquaria the maximumETR was reduced from 55 µmol e- m-2 s-1 to 16 µmol e- m-2 s-1. This most likely relates to thehigher light environment or increased water temperature in the aquaria. During desiccation,the rate of decline in maximum ETR between the two populations was not significantlydifferent, indicating no adaptation to desiccation by the reef flat algae (Figure 4).
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Figure 4. Maximum ETR of Chlorodesmis fastigiata collected from the reef flat and from 15m depth.During the first hour samples were maintained in flow-through aquaria. For the second hourthey were desiccated in the shade.
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Shading to 50% of incident light reduced the maximum ETR in C. fastigiata from60 µmol e- m-2 s-1 to 25 µmol e- m-2 s-1 after 1 d. No further reduction in ETR occurred overthe next 3 d. Photosynthesis in Chnoospora implexa declined from 160 µmol e- m-2 s-1 to100 µmol e- m-2 s-1 after 1 d. Further reductions in maximum ETR over the next 3 d reducedthe maximum ETR to 50 µmol e- m-2 s-1 (Figure 5).
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Figure 5. The response of ETR versus PAR curves in Chlorodesmis fastigiata and Chnoospora implexato shading (50% of incident light) over four days.
Total shading of UVR resulted in an increase in the maximum ETR in Padina tenuis from40 µmol e- m-2 s-1 to 80 µmol e- m-2 s-1 over a 4 day period. There was no change however inthe maximum ETR of C. fastigiata (Figure 6).
Fertilisation significantly increased the maximum ETR in P. tenuis from30 µmol e- m-2 s-1 to 45 µmol e- m-2 s-1. There was no change in the maximum ETR ofC. fastigiata, but the photoinhibitory response was reduced. By contrast, Colpomenia sinuosashowed no significant response to fertilisation and Acropora sp. showed a slight decline(Figure 7).
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DISCUSSIONComparisons of the maximum ETR indicate that the photosynthetic activity between speciesis variable, ranging from 15 µmol e- m-2 s-1 for Halimeda opuntia to 230 µmol e - m-2 s-1 forTridacna maxima (Table1). Titlyanov et al. (1994) reported that of twenty species ofmacroalgae tested, Halimeda sp. had the lowest photosynthetic capacity. It is unclear whythe maximum ETR of the zooxanthellae of T. maxima is so much higher than those in Acroporasp. It may be due to the presence of a different species of Symbodinium (Rowan & Powers,1991) or due to the presence of more autofluorescent chromatophores in T. maxima(Schlichter et al., 1994). Autofluorescent chromatophores transform short wavelengthirradiance, which is less suitable for photosynthesis, into longer wavelengths which arephotosynthetically more effective.
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Figure 6. The response of ETR versus PAR of Chlorodesmis fastigiata and Padina tenuis to screeningof 100% of incident UVR.
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Table 1. The maximum recorded ETR for selected photosynthetic marine organisms on Heron Islandreef flat
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Figure 7. The response of ETR versus PAR to fertilisation (88g m-2 N, 22g m-2 P).
Group Species Maximum ETR (µmol e- m-2 s-1)
Phaeophyta Chnoospora implexa 150
Colpomenia sinuosa 125
Padina tenuis 40
Rhodophyta Plocamium microcladioides 95
Chlorophyta Chlorodesmis fastigiata 60
Caulerpa racemosa 40
Halimeda opuntia 15
Microalgae Tridacna maxima (zooxanthellae) 230
Acropora sp. (zooxanthellae) 70Benthic microalgae 45
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Although the ETR of the benthic microalgae was not relatively high(45 µmol e- m-2 s-1), its high biomass across the reef flat (Heil et al., in review) makes itpotentially a very large contributor to total primary production. Interestingly, the specieswith the highest ETRs are from the Phaeophyta (Chnoospora implexa and Colpomeniasinuosa) and Rhodophyta (Plocamium microcladioides). The higher ETRs of the brown andred algae relative to the green algae (Chlorophyta) may be due to the enhanced light capturingability of their accessory pigments (Rowan, 1989).
For calculation of the absolute ETR based on PAM fluorometry, it is necessary to carefullymeasure the mean specific absorption coefficient of the algae (Hartig et al., 1998). Due tothe difficulty in obtaining an accurate value, most workers are currently using the defaultabsorbance of 0.84, which was calculated as an average based on analysis of terrestrial leaves.The use of this value is acceptable for comparing the ETR of individuals within one species.It must be realised, however, that there will be some error associated when comparing specieswith very different morphologies, pigment complements and chloroplast configurations unlessthe species specific absorption coefficients are calculated (Cunningham et al., 1996).
The transect conducted across the reef to measure the maximum ETR ofC. fastigiata shows definite peaks at the gutter just offshore from the beach, and at the reefcrest. The gutter is formed due to scouring from the action of waves returning from thebeachrock with more energy (Coote, 1984; Kan, 1995). The region is characterised by highercurrent velocities at the sediment surface than the rest of the reef flat (Heil et al., in review).The increased water motion in these areas reduces the boundary layer allowing faster diffusionof CO
2 across the cell membrane (Shashar et al., 1996). This may explain the higher ETR
measured in C. fastigiata at these sites. Benthic microalgae within this same gutter region atHeron Island have a much higher ETR than other areas on the reef flat (Heil et al., in review).Another possible explanation for the higher rates in the gutter is the increased depth in thisregion, which may provide slightly more protection from PAR and UVR.
The highest ETR in C. fastigiata along our depth transect was recorded at10 m. This could be due to the increased concentration of chlorophyll a within plants atdepth (Titlyanov et al., 1992; Titlyanov et al., 1994) or may be a reflection of photoinhibitionin plants at the surface due to exposure to high solar radiation or high UVR. The subsequentdecline in ETR observed at 15 m may be due to reduced light availability from attenuationthrough the water column, or because of changes in the spectrum of available PAR. As achlorophyte, C. fastigiata absorbs predominantly red and blue light for photosynthesis(Rowan, 1989) and the red wavelength is the first to be absorbed by water. Other workershave also found optimal photosynthetic rates at intermediate depths. In particular, Hader etal. (1997) demonstrated optimal photosynthesis using oxygen evolution measurements ofanother chlorophyte (Caulerpa prolifera) at a depth of 5 m. Photoinhibition at mid rangedepths may be further reduced because UVR is removed at a much greater rate than PARthrough the water column (Franklin & Forster, 1997).
The impact of desiccation on the rate of decline in the maximum ETR ofC. fastigiata was not different between the reef flat and 15 m samples. However, the initialdecline in the ETR of the 15 m deep samples from 55 µmol e - m-2 s-1 to 16 µmol e- m-2 s-1 isprobably because of the increased solar radiation or increased temperature in the aquaria.This indicates that adaptations to localised reef environments have occurred betweenindividual populations of C. fastigiata. This is consistent with the work of Porst et al. (1997)and Franklin et al. (1996), who observed a drastic decline in the effective quantum yield
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when plants collected from depth were exposed to high irradiance at the surface. Previousstudies have demonstrated increases in photosynthesis (measured by oxygen evolution) duringinitial emersion when the thallus has lost 10-20% of its initial water content (Johnson et al.,1974; Quadir et al., 1979; Dring & Brown, 1982; Beer & Eshel, 1983), although aftersubstantial desiccation photosynthesis declines (Dring & Brown, 1982). The initial increasemay be due to a reduction in the boundary layer when the water covering the surface of thethallus has evaporated, thereby allowing for faster rates of diffusion of CO
2. In this study,
there was no initial increase in ETR during desiccation. This may be related to differences inmorphology between C. fastigiata and the species tested by other workers. The samples ofC. fastigiata for this study were collected from near the gutter region of the reef flat and soare never exposed at low tide. Perhaps if they had been collected from the reef crest, whichis exposed at low tide, there may have been adaptations to resist desiccation.
Both PAR and UVR can be important inhibitors of photosynthesis in marine plants (Hader etal., 1996a). Shading to 50% of incident light reduced the maximum ETR of C. fastigiata andC. implexa. There was a gradual reduction in maximum ETR over 4 d by C. implexa comparedto the relatively quick decline over 1 d in C. fastigiata. This suggests that C. implexa hasmore efficient photoadaptive mechanisms to better tolerate changes in light availability. It isalso interesting to note the lack of photoinhibition in C. implexa at higher PAR during therapid light curve, another indication of resistance to changes in light availability. The alphaof the ETR versus PAR curves was also significantly different between the two species, withC. fastigiata being more efficient at lower irradiances, even though it was C. implexa whichwas better able to cope with the reduced light regime.
Padina tenuis had a significantly higher maximum ETR after being shaded from 100% ofthe UVR, which is consistent with the findings of Hader et al. (1996c) and Porst et al.(1997) who observed that UVR had an overproportional inhibitory effect on thephotosynthesis of two species of chlorophyte. However, in this study C. fastigiata showedno significant change even after 4 d. The filaments that make up C. fastigiata are each onecell, which allows the plant to move its chloroplasts to the tip for photosynthesis, or downinto the base of the filaments for repair. This cytoplasmic streaming has been postulated asa mechanism by which this species can mitigate the photoinhibitory effects of UVR and highPAR (Franklin & Larkum, 1997).
The fertilisation responses in P. tenuis and C. fastigiata are consistent with an increase inpigment concentrations facilitating higher photosynthesis and reduced photoinhibition athigher light (Titlyanov et al., 1994). Fertilisation of Acropora sp. resulted in a slight depressionin maximum ETR, although the initial fluorescence was significantly higher. Higher initialfluorescence may indicate that there is greater biomass of zooxanthellae within the coral.Several studies (Hoegh-Guldberg & Smith, 1989; Dubinsky et al., 1990; Stimson & Kinzie,1991) have reported that supplying nitrogen to coral resulted in increases in the populationof zooxanthellae, but reductions in the per cell rates of photosynthesis. They hypothesisedthat this was due to shading and competition for CO
2 because of the greater algal density. The
host corals may also play a role in preventing access of the zooxanthellae to the intracellularnutrients within the host, which are speculated as being quite low (Muscatine, 1980; D’Elia& Cook, 1988; Snidvongs & Kinzie III, 1994). Another possible theory for the reduction inphotosynthesis in the N and P fertilised treatment relates to the effects of dissolved inorganicphosphorus on inhibition of calcification by the coral host (Simkiss, 1964). This reducesthe availability of CO
2 to the symbiotic algae (Snidvongs & Kinzie III, 1994).
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In conclusion, our investigations reveal that PAM fluorescence measurements can be usedto assess the photosynthetic response of macroalgae, zooxanthellae and benthic microalgaeto a variety of environmental parameters such as desiccation, light, and nutrient availability.
ACKNOWLEDGEMENTSThe authors would like to thank the students of the field courses, Coral Reef Biology andGeology (1996) and Terrestrial and Marine Environmental Physiology (1997) for help withsample collection. We would like to thank two anonymous reviewers whose comments ledto substantial revision and improvement to the manuscript.
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