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The effects of fish waste and oxytetracycline on the microbenthos Wu, Henry C. 1992

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THE EFFECTS OF FISH WASTE AND OXYTETRACYCLINE ON THE MICROBENTHOS  By HENRY C. WU B.Sc.(H), The University of British Columbia, 1988 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTERS OF SCIENCE in THE FACULTY OF GRADUATE STUDIES (Department of Zoology)  We accept this thesis as conforming to the required standard  THE UNIVERSITY OF BRITISH COLUMBIA April 1992 © Henry C. Wu, 1992  In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission.  (Signature)  Department of The University of British Columbia Vancouver, Canada  Date  DE-6 (2/88)  ftr-^3o /97-  11  ABSTRACT  Fish feed, feces and oxytetracycline (OTC) were added to sediment microcosm tanks in 1990, and the physical and biological changes in the benthic environment were followed over a period of 28 days. The main objective was to study the response of the microfauna (benthic microbial and protozoan communities) to fish waste. Results indicated that anoxic and highly reducing conditions were quickly reached in the sediments within a day, with a shift away from microbially mediated sulphate reduction to methanogenesis in the degradation of fish waste. Most of the OTC was quickly washed out of the fish waste when added to the sediments, with small quantities of the antibiotic persisting throughout the duration of the study. Bacterial abundance rapidly increased in the sediments when fish waste and oxytetracycline were added. While there was a potential for OTC to act on the microfaunal communities, this effect could not be conclusively demonstrated.  A second study was conducted in which fish feed, feces and oxytetracycline were added gradually to sediment microcosm tanks in 1991 for 20 days. The main objective of this study was to examine the changes in the physical-chemical regime and the benthic microfaunal community over time. The gradual addition of fish waste to the sediments resulted in microbially mediated sulphate reduction being the main route for the degradation of fish waste. As most of the oxytetracycline was quickly washed out of the fish waste, the amount of the antibiotic was insufficient to affect the bacterial and protozoan communities. There were two rapid increases in ciliate abundance, with the initial bloom occurring before the increase in bacterial abundance, and the decrease in the ciliate population was correlated to increased reducing conditions in the sediments. A large number of heterotrophic microflagellates developed after the ciliate population declined, and these flagellates were observed to be actively grazing on the bacterial population. The short time interval of the study was considered to be insufficient for noticeable changes to have occurred in the meio- and macrofauna. The rapid response of the microfauna to fish waste suggests that changes in the protozoan community could be used as a biological tool in monitoring the impact of fish waste on the environment. More efficient management at aquaculture facilities could minimize environmental problems.  iii TABLE OF CONTENTS  ABSTRACT ^ TABLE OF CONTENTS ^ LIST OF FIGURES ^ LIST OF TABLES ^ ACKNOWLEDGEMENTS ^ GENERAL INTRODUCTION ^  vi viii ix 1  CHAPTER 1. THE EFFECTS OF FISH WASTE AND OXYTETRACYCLINE ON THE MICROBENTHOS. A PRELIMINARY INVESTIGATION. 1.1 INTRODUCTION ^  5  1.2 OBJECTIVES ^  9  1.3 MATERIALS AND METHODS 1.3.1 EXPERIMENTAL DESIGN ^  9  1.3.2 APPLICATION OF FISH WASTE AND OXYTETRACYCLINE ^  10  1.3.3 SAMPLING DESIGN ^  12  1.3.4 DATA ANALYSIS^  17  1.4 RESULTS 1.4.1 SEDIMENT COMPOSITION ^  17  1.4.2 VISUAL OBSERVATIONS ^  17  1.4.3 REDOX PROFILES (Eh) ^  20  1.4.4 DISSOLVED OXYGEN ^  23  1.4.5 CARBON: NITROGEN RATIOS ^  23  1.4.6 OXYTETRACYCLINE ^  27  1.4.7 MICROBIAL ABUNDANCE^  27  1.4.8 PROTOZOAN ABUNDANCE ^  30  iv 1.5 DISCUSSION 1.5.1 REDOX MEASUREMENTS ^  30  1.5.2 MICROBIAL MAT ^  32  1.5.3 DISSOLVED OXYGEN ^  33  1.5.4 CARBON:NITROGEN RATIOS ^  33  1.5.5 OXYTETRACYCLINE^  35  1.5.6 MICROBIAL ABUNDANCE^  37  1.5.7 POTENTIAL EFFECTS OF OTC ON MICROFAUNA ^  38  1.5.8 ENVIRONMENTAL CONCERNS ^  44  CHAPTER 2. THE EFFECTS OF FISH WASTE AND OXYTETRACYCLINE ON THE MICRO- AND MEIOFAUNA. 2.1 INTRODUCTION ^  49  2.2 OBJECTIVES ^  51  2.3 MATERIALS AND METHODS 2.3.1 EXPERIMENTAL DESIGN ^  51  2.3.2 THEORMCAL CALCULATION OF ORGANIC LOADING AND OTC ^  54  2.3.3 APPLICATION OF FISH WASTE AND OTC ^  56  2.3.4 SAMPLING DESIGN^  56  2.4 RESULTS 2.4.1 VISUAL OBSERVATIONS ^  58  2.4.2 REDOX PROFILES (Eh) ^  59  2.4.3 DISSOLVED OXYGEN ^  61  2.4.4 OXYTETRACYCLINE ^  65  2.4.5 BACTERIAL ABUNDANCE ^  65  2.4.6 MICROBIAL MAT^  67  2.4.7 PROTOZOAN ABUNDANCE ^  68  2.4.8 DIATOM, NEMATODE AND COPEPOD ABUNDANCE ^  70  V 2.5 DISCUSSION 2.5.1 REDOX MEASUREMENTS  ^ 72  2.5.2 MICROBIAL MAT ^  76  2.5.3 DISSOLVED OXYGEN ^  76  2.5.4 OXYTETRACYCLINE ^  77  2.5.5 MICROFAUNAL ABUNDANCE ^  78  2.5.6 MICROFAUNAL INTERACTIONS ^  81  2.5.7 DIATOMS ^  84  2.5.8 MEIOFAUNA^  85  2.5.9 INTERACTIONS BETWEEN MICRO- AND MEIOFAUNA ^  87  2.5.1 0 MACROFAUNA ^  88  2.5.11 PROTOZOANS IN BIOLOGICAL MONITORING ^  89  2.6 SUMMARY ^  91  2.7 RECOMMENDATIONS FOR FUTURE STUDIES ^  92  APPENDIX A. EXTRACTION OF ATTACHED BACTERIA IN SEDIMENTS BY ULTRASONICATION^  94  APPENDIX B. EXTRACTION OF OXYTETRACYCLINE FROM SEDIMENTS BY HIGH-PERFORMANCE LIQUID CHROMATOGRAPHY. A CALIBRATION AND RECOVERY STUDY ^  REFERENCES ^  103  111  vi  LIST OF FIGURES Figure 1.1. Generalized SAB diagram of changes along a gradient of organic enrichment. Where S=Species Diversity, A=Total Abundance, B=Total Biomass, PO=Peak of Opportunists, E=Ecotone Point, TR=Transition Zone. (From Pearson & Rosenberg 1978) ^  6  Figure 1.2. The arrangement of the ten, 20 litre cylindrical benthic microcosm tanks with black plastic covers, side profile. A header tank supplies a common source of seawater to each microcosm ^  11  Figure 1.3. A schematic of the 7 transect lines used for randomly sampling the benthos, with the central drainage pipe. The sediment cores would be taken along the transect line ^  13  Figure 1.4. A 3 cc syringe-corer, modified to obtain Eh depth profiles at every 0.5 cm ^ 14 Figure 1.5. Grain size composition by relative percentages of sediment obtained from 30 to 40 ft depth near Spanish Banks/U.B.C. The major constituents were the fine to medium sand at 31% (from 180-355 gm), followed by the very fine sand at 29% (from 75-180 gm), and the coarse silt/mud at 26% (< 53 gm) ^ 18 Figure 1.6. Reduction-oxidation time-series depth profiles of the replicate tanks for the Control Treatment in 1990. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment surface 21 Figure 1.7. Reduction-oxidation time-series depth profiles of the replicate tanks for Treatment A in 1990, a low dosage of oxytetracycline. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment ^  22  Figure 1.8. Reduction-oxidation time-series depth profiles of the replicate tanks for Treatment B in 1990, a high dosage of oxytetracycline. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment 24 Figure 1.9. Comparison of the dissolved oxygen levels in seawater between the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3 ^  25  Figure 1.10. Carbon:nitrogen ratios of the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3 ^ 26 Figure 1.11. The concentrations of oxytetracycline (ppm) for Treatment A (2.53 g of OTC) and Treatment B (10.13 g of OTC). Plotted values are the Mean ± 1 Standard Deviation, n=3 ^ 28 Figure 1.12. Bacterial abundance (cells x 10 9 • 1 sediment dry weight) between the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3 ^ 29  vii Figure 1.13. Electron free-energy diagram for the biologically mediated redox sequences of the degradation of organic carbon, CH 2 O (acting as the electron donor). Note the relationship between the change in energy (G°), redox processes and the Eh (mV) (from Zehnder & Stumm 1988) 31 Figure 2.1. The arrangement of the 200 L tanks in which fish waste and oxytetracycline were applied. The order of the tanks for the treatments was randomly chosen' ^ 52 Figure 2.2. Grain size composition by relative percentages, of the sediment obtained from the intertidal zone during low tide at Spanish Banks/U.B.C. The majority of the sediment is comprised of fine to medium sand (65%), followed by a medium to 53 coarse sand component (26%) ^ Figure 2.3. A schematic of the division of grids used to randomly sample the sediments of the 200 L tanks. The sediment is divided into a system of 3 columns (A-C) and 5 rows (1-5) with a central drainage pipe and 1.5" border which was not sampled to avoid edge effects ^ 55 Figure 2.4. Reduction-oxidation time-series depth profiles of the replicate tanks for the Blank Treatment, followed over 20 days in which no fish waste or oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. 60 Measurements began 0.5 cm above the sediment surface ^ Figure 2.5. Reduction-oxidation time-series depth profiles of the replicate tanks for the Control Treatment, followed over 20 days in which fish waste but no oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began 0.5 cm above the sediment surface ^ 62 Figure 2.6. Reduction-oxidation time-series depth profiles of the replicate tanks for the OTC Treatment, followed over 20 days in which fish waste and oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. 63 Measurements began 0.5 cm above the sediment surface ^ Figure 2.7. Comparison of the dissolved oxygen levels in seawater between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3 ^  64  Figure 2.8. Comparison of bacterial abundance (cell x 10 9 -g- ' sediment dry weight) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation 66 Figure 2.9. Comparison of protozoan abundance over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added); where A=ciliate abundance (numbers x 10 3 -cm-2 ); B=heterotrophic microflagellate abundance (numbers x 10 6 -cm-2 ). Plotted values are the Mean ± 1 Standard Deviation ^ 69 Figure 2.10. Comparison of diatom abundance (numbers x 10 2 -cm -2 ) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted 71 values are the Mean ± 1 Standard Deviation ^  viii Figure 2.11. Comparison of nematode abundance (numbers•cm -2 ) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation ^ 73 Figure 2.12. Comparison of harpacticoid copepod abundance (numbers•cm -2 ) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation 74 APPENDIX A  Figure 1. Accumulation of bacteria attached to sediments following ultrasonication times of 30, 60, 90, 120, 180, 240 and 300 seconds, of a formalin preserved sand sample. Values presented are the Mean ± 1 Standard Deviation, n=5 ^ 97 Figure 2. The destructive effects on a bacterial suspension extracted from the sediment, following ultrasonication times of 30, 60, 90, 120, 180, 240 and 300 seconds. Values presented are the Mean ± 1 Standard Deviation, n=4 ^ 98 Figure 3. The percentage of bacteria remaining on sediments, following ultrasonication periods of of 90, 150, 210 and 270 seconds. The relationship is linearized by log transformation and fitted with a first order regression and 95% confidence limits, n=16 100  LIST OF TABLES  TABLE 1. Total Carbon content (percent) of the sediments in the Control, Treatment A, and Treatment B. Values presented are the Mean % ± 1 Standard Deviation (SD)....34  ix ACKNOWLEDGEMENTS  I would like to thank my supervisor Dr. T.R. Parsons, and the other members of my Research Committee (Dr. C.D. Levings and Dr. R. Petrell) for their advice and continuing support throughout this study. My thanks to the following people for their invaluable assistance; S. Mattice in the setup and maintenance of the microcosm tanks, Dr. H. Rogers who kindly provided access to his lab and HPLC (DFO, West Vancouver) and M. Sadar in the analysis of oxytetracycline in 1990, Dr. K. McErlane who provided an HPLC in 1991 and to R. Aoyama for his suggestions during the analysis of oxytetracycline. I am especially grateful for the useful criticisms of my Research Committee, and from Dr. M. St. John, Dr. C. Lalli and J. Berges on various drafts of this thesis. The suggestions provided by SCARL (Statistical Consulting and Research Laboratory, Dept. Statistics), J. Berges, D. Montagnes, T. Sutherland and M. Adl during the course of this research were particularly helpful. Funding for this project was made available from NSERC operating grants to Dr. T.R. Parsons with additional support by the Department of Fisheries and Oceans through Dr. C.D. Levings, and from operating grants provided by the Donner Foundation. My deepest gratitude to my parents and Jo for their encouragment, patience and understanding, and to whom I would like to dedicate this thesis.  1  GENERAL INTRODUCTION  World aquaculture (the production of finfish, crustaceans, molluscs and seaweeds) has increased greatly in the last three decades (Iwama 1990), from 5.6 million t in 1975 to a projected value of over 22.2 million t in the year 2000. The socioeconomic aspects of aquaculture can be considerable, being an important industry in some countries (e.g. Norway) and employing many people (Ridler 1984, Ford 1984, Folke & Kautsky 1989). In British Columbia, the aquaculture industry dates back to 1912 with Pacific oyster cultivation, while trout (freshwater) culture began in the 1950s and salmon (marine) culture in the 1970s. The farming of other aquatic species in B.C. such as clams, scallops, mussels, Arctic char, black cod (sablefish), sturgeon and marine plants (e.g. kelp) is just beginning. By far the largest and most successful activity (total value) has been in marine fish culture. While salmon farming is a relatively young industry, there have been large increases in the number of farms and production, from eight operating farms producing 107 t of chinook and coho salmon in 1985 (Korman 1989), to 125 farms in 1991 producing 16,500 t of chinook, coho, and Atlantic salmon worth $105 million (B.C. Salmon Farmers Association 1992). An additional 68 farm sites could be in operation within the next few years once licensing and development is completed. Salmon farming along the coast of British Columbia is intensive in local areas. The rearing of fish occurs in floating net pens and the diet is composed almost exclusively of artificial feed. Unfortunately, the output from this intensive culturing practice is unavoidably released locally into the marine environment, affecting water and sediment quality. This output is composed of the uneaten feed and metabolic by-products of the fish, of which the major constituents are organic carbon and nitrogen compounds (carbohydrate, lipid and protein), followed by phosphate, ammonium, urea, bicarbonate, vitamins, therapeutants and pigments (e.g. Gowen & Bradbury 1987, Enell 1987, Aquametrix Research Ltd. 1988). The particulate excess feed and fecal material sink to the sediments in  2  the immediate area around the fish farms, while the soluble components (e.g. ammonium, urea) are diluted and dispersed to the water column (e.g. Hall & Holby 1986, Enell 1987). The magnitude and nature of this fish waste is affected by the biomass of the farm and feeding methods used, which determine the amount of food fed and wastage levels from uneaten feed and fecal material (Iwama 1990). The size of the fish determines the size of the feed and types of feed used (e.g. starter feed vs production feed, dry vs moist feeds) all of which are involved in the rate of breakdown in the excess feed and the type of fecal matter produced. These are important characteristics to consider as the size of the particles are inversely related to the areal dispersal pattern (smaller, lighter particles sink slower and are dispersed farther) which further depends on the physical oceanography of the region (currents, tides, and bottom topography) and on the distance of the net pen above the ocean floor (Weston 1986). Salmon farming will result in changes to the nearby marine environment, altering the physical, chemical and biological regimes. There are several potential environmental effects (as reviewed by Weston 1986, with some modifications as increased research has broadened the scope of problems):  (1)  Changes in water circulation.  (2)  Sedimentation and accumulation of the fish waste beneath the culture operation.  (3)  Alterations in water chemistry.  (4)  Changes in plankton and benthic algal biomass and productivity.  (5)  Changes in structure of the benthic faunal communities.  (6)  Introduction of exotic species and subsequent changes in the genetic fitness of wild stocks.  (7)  The effects of antibiotics on the surrounding biota and development of bacteria pathogenic to humans.  (8)  Disease transmission from cultured to wild stocks.  3  Since the review of the environmental impacts of mariculture by Weston (1986), there has been a great deal more interest worldwide of the potential environmental problems associated with this rapidly expanding industry (Gowen & Bradbury 1987, Eng et al. 1989, Munro 1990, Iwama 1990). Further research to quantify the nutrient flux from mariculture activities has resulted in the development of models (Hall et al. 1990, Holby & Hall 1991, Ackefors & Enell 1990). The possibilities for disease and parasite transmission, and for gene transfer between wild and escaped farmed fish resulting in reduced genetic variability have also been examined (Windsor & Hutchinson 1990, Carss 1990, Egidius et al. 1991). There have been recent investigations on the effects of antibiotics (used to treat disease outbreaks at fish farms) on the microbial community in marine sediments in Norway (oxytetracycline, oxolinic acid, furazolidone; e.g. Samuelsen et al. 1991 & 1992, Nygaard et al. 1992). Additional work continues on the behaviour of these antibiotics in the marine environment and fauna (Samuelsen et al. 1988, Lunestad & Goksoyr 1990, Bjorklund et al. 1990 & 1991) and on the potential for development of antibiotic resistance (Torsvik et al. 1988, Husevag et al. 1991). The physical and chemical changes of the sediments (due to the high sedimentation rates of fish waste from net and cage pens) have been studied (Hall & Holby 1986, Samuelsen et al. 1988, Kaspar et al. 1988). The possibility that nutrientenriched environments (a condition known as hypernutrification or eutrophication) around aquaculture facilities may support increases in the planktonic and benthic algal communities has also been considered (Weglenska et al. 1987, Parsons et al. 1990, Korman 1989, Stirling & Dey 1990, Carr & Goulder 1990). The response of the macrofauna to enrichment of the benthic environment by excess fish feed and fecal matter has been well established (Brown et al. 1987, O'Connor et al. 1989, Lumb 1989, Lumb & Fowler 1989, Frid & Mercer 1989, Ye et al. 1991).  4  However, the effects of fish waste should not be restricted solely to the macrofauna, as an integral benthic community consists of micro- and meiofaunal components as well (Fenchel 1978). But changes in the micro- and meiofauna due to organic enrichment resulting from aquaculture activities have never been considered before. Thus, the overall objective of this study was to examine the responses of the microfaunal community (and the meiofaunal community to a limited extent) to fish waste and the antibiotic oxytetracycline. Oxytetracycline was chosen as it has been the most commonly used drug in the treatment of diseases at fish farms for approximately the last two decades (Grave et al. 1990). For a comprehensive review of the environmental impacts associated with aquaculture, the reader should refer to Weston (1986), Gowen & Bradbury (1987), Iwama (1991) and Levings (1992).  5  1. THE EI4ECTS OF FISH WASTE AND OXYTETRACYCLINE ON THE MICROBENTHOS. A PRELIMINARY INVESTIGATION.  1.1 INTRODUCTION  There is a lack of comprehensive research on changes in the microbenthic communities (bacterial and protozoan populations) beneath fish farms caused by fish waste and oxytetracycline, a commonly used antibiotic. Previous studies have concentrated mainly on changes in meio- and macrofaunal communities in the sediments due to fish wastes (e.g. polychaetes, molluscs, amphipods, copepods, snails; Weston 1986 & 1990, Moriarty 1986, Brown et al. 1987, Gowen & Bradbury 1987, Lumb 1989, O'Connor et al. 1989, Ritz et al. 1989). The terms micro-, meio- and macrofauna refer to the faunal communities of the benthic environment, and are used interchangably with micro-, meioand macrobenthos. Only a few recent reports have examined the benthic diatom communities around aquaculture sites (Stirling & Dey 1990, Carr & Goulder 1990), and the remaining studies have focused on bacterial abundance and resistance to antibiotics (e.g. Torsvik et al. 1988, Husevag et al. 1991, Samuelsen et al. 1988, 1991 & 1992). Because benthic biota are tolerant of a certain amount of environmental variation, the responses of each benthic species will be determined by the magnitude of organic enrichment (Blackstock 1984). Characteristic changes in the species, abundance and biomass of the meio- and macrobenthic communities in response to pollution are both temporal and spatial (e.g. Pearson 1975, Pearson & Rosenberg 1978, Bagheri & McLusky 1982, Oviatt et al. 1987, Warwick et al. 1987, Whitehurst & Lindsey 1990, Moore & Rodger 1991). As organic enrichment decreases along a gradient from a central source (Fig. 1.1), there is an initial small increase in biomass corresponding to large abundances of small opportunistic species, followed by a decline to the ecotone point. This ecotone point is a transition zone in which a community is poor in species, abundance and biomass  6  PO  •  E •  4r....■■■■••  TR  elr.■.............►  INCREASING ORGANIC INPUT  Figure 1.1. Generalized SAB diagram of changes along a gradient of organic enrichment. Where S =Species Diversity, A =Total Abundance, B=Total Biomass, PO=Peak of Opportunists, E=Ecotone Point, TR=Transition Zone. (From Pearson & Rosenberg 1978).  7  (Pearson & Rosenberg 1978). After the ecotone point and continuing along the gradient of decreasing organic enrichment, biomass increases to a second maximal level with a greater variety of species and lowered abundance. Species diversity, abundance and biomass levels eventually stabilize to resemble non-polluted communities at the outer boundaries of the zone of pollution. Changes in part, or all of the microbenthic community due to anthropogenic pollution (domestic, agricultural and industrial) are also similar to changes in the meio- and macrobenthic communities (e.g. Small 1973, Curds 1973, Pearson 1975, Madoni & Ghetti 1981, Wyatt & Pearson 1982, Heip et al. 1985, Hul 1987 & 1988, Stoessel 1989). Those benthic species that are highly sensitive or are opportunistic in organically enriched sediments have been used as indicator organisms to identify the degree of pollution (e.g. mussels, nematode/copepod ratio, polychaetes, protozoans; Raffaelli & Mason 1981, Raffaelli 1987, Grabacka 1985, Dale 1987, Warwick et al. 1988, Manru et al. 1988, Gray et al. 1988, Warwick et al. 1990). There has not been a comprehensive study of the effects of fish waste and OTC on the microbenthic community for a number of reasons. Firstly, aquaculture has not been a large industry worldwide until the last two decades (see reviews by Phillips et al. 1985, Weston 1986, Windsor & Hutchinson 1990, Munro 1990, Iwama 1990, Svealv 1991). Only recently has more attention been directed toward calculating the different components and quantities of the pollutants from aquaculture facilities (i.e. fluxes and mass balances; Enell 1987, Korman 1989, Aure & Stigebrandt 1990, Hall et al. 1990, Holby & Hall 1991). Secondly, macrofauna have been traditionally used in pollution monitoring due mainly to the practical advantages of handling larger organisms in the field and laboratory, and because the taxonomy and general biology is more widely known (Gray et al. 1988). Thirdly, the main drawback involved in using microfauna (and meiofauna) in pollution studies are the taxonomic difficulties of the many different species. Finally, the logistics of quantitative biological surveys require that soft-bodied organisms (e.g. protozoans) be examined live, or fixed and preserved either in the sediment or immediately after extraction  8  (Heip et al. 1988). While the sediment microbial population can be easily preserved and counted (Montagna 1982), quantitative extraction, identification, fixation and preservation of benthic protozoans (also referred to as 'ciliates'), is an ongoing problem (Schwinghamer 1981, Alongi 1986). However, the advantages in using the microfauna are that smaller sediment samples are needed (i.e. field sampling is less labour intensive), and shorter generation times can result in potentially faster responses to pollution. Thus, the benthic microfauna can serve as a more sensitive indicator of pollution from mariculture sites (e.g. Finaly et al. 1979). Benthic microcosm tanks were used to assess the changes in the microfaunal community that may result from pollution of the sediments with fish waste and oxytetracycline. Although microcosm tanks may not provide a totally realistic benthic community structure because they may uncouple the physical and biological processes (Pritchard & Bourquin 1984, Federle et al. 1986, Findlay et al. 1990a), they have still been used extensively to study the benthos (e.g. Hargrave 1972, Kelly & Nixon 1984, Alongi & Tenore 1985, Nedwell & Lawson 1990, Hansen & Blackburn 1991, Sundback et al. 1991). Controlled conditions are always required to determine a definitive cause and effect relationship between the organisms and pollutants involved, whereas this relationship can only be inferred from field studies (Underwood & Peterson 1988, Lasserre 1990, Pilson 1990). Directions for possible future studies and interpretations of general interest can thus be obtained from the use of benthic microcosms in this study. Any experimental results obtained from a laboratory setting must eventually be applied towards understanding interactions in the field.  9  1.2 OBJECTIVES Following enrichment of the sediments with fish waste and oxytetracycline, the objectives of this initial study were:  (1) To evaluate the physical-chemical changes to the benthos such as (i) oxygen levels of seawater due to the biological oxygen demand of the sediments, (ii) sediment depth profiles of the reduction-oxidation potentials, and (iii) carbon: nitrogen ratios of the sediments. (2) To follow the decrease of oxytetracycline in the sediments over time. (3) To determine possible changes in the abundance of microfauna in the sediments (bacterial and protozoan populations).  1.3 MATERIALS AND METHODS  1.3.1 EXPERIMENTAL DESIGN A controlled experimental ecosystem approach was developed in which a series of microcosms (< 1 m 3 volume, Lasserre 1990) were used to create a benthic environment enriched in fish waste. Seawater and tank facilities were located at West Vancouver Laboratories, Department of Fisheries and Oceans, West Vancouver, British Columbia, Canada. The duration of this experiment was 28 days, from July to August 1990. Sediment was obtained with a Shipek grab sampler (3.3 L), near Spanish Banks/U.B.C., from 30 to 40 ft depth during low tide. This ensured that the sediment and benthic community was from a depth similar to that expected from net pens (e.g. Weston 1986), and any changes to the benthos could be understood better in relation to natural field conditions. Three to four grabs were combined to yield enough sediment for the microcosm tanks (4 to 6 cm depth). A period of two weeks was allowed for the biological and chemical properties of the benthic system to be re-established (Federle et al. 1986).  10  Ten, 20 L cylindrical tanks were used (Fig. 1.2), with a header tank supplying a continuous flow of seawater to each tank at 0.5 L.min -1 . This was the maximum rate of flow that would not disturb the sediments in the tanks. Drainage was by a central pipe in the centre of each tank, with the outflow near the water surface. Water temperature and salinity ranged from 13 to 16°C and 27 to 29 0/. throughout the study period, respectively. Based on 9 tanks, 3 treatments were designed with 3 replicates per treatment, all with the same levels of fish waste (feces and feed) but with differing levels of oxytetracycline (OTC). The treatments consisted of a relatively low and high concentration of OTC, with a control series (no OTC added) for comparison.  1.3.2 APPLICATION OF FISH WASTE AND OXYTETRACYCLINE  Fish feces and excess feed were collected from 10,000 L outdoor tanks at West Vancouver Laboratories, West Vancouver, B.C. with the assistance of Mr. S. Mattice. The antibiotic oxytetracycline was provided by Argent Laboratories, Richmond, B.C., in the form of Oxysol 440, which contains 100 g of OTC per 227 g Oxysol 440. A treatment level of 11.5 g Oxysol 440 per kg of fish feed was prescribed (Dr. E.L. Dahl, D.V.M., pers. comm.). The fish waste was applied in all tanks to a depth of 2 cm, to resemble mild to heavy loading beneath fish farms (Jacobsen & Berglind 1988). The fish waste was measured using a 10 L container and poured through a funnel and 1.5" diameter plastic hose, with the end of the hose held just above the sediment surface. To avoid disruption of the established reduction-oxidation characteristics, care was taken to avoid disturbing the sediment as the fish waste was evenly and carefully layered on the sediment surface. The Control treatments consisted of only fish waste, while the low concentration treatment of OTC (Treatment A) contained 5.75 g of Oxysol 440 (or 2.53 g OTC) in the fish waste. The high concentration treatment of OTC (Treatment B) contained 23 g Oxysol 440 (or 10.13 g OTC). The Oxysol 440 was added directly to the fish waste and stirred well to  11  Figure 1.2. The arrangement of the ten, 20 litre cylindrical benthic microcosm tanks with black plastic covers, side profile. A header tank supplies a common source of seawater to each microcosm.  12  ensure complete mixing before addition to the sediments. The replicate tanks for each treatment were randomly chosen and dark covers placed over all of the tanks to minimize (i) autotrophic activity due to poor and variable indoor lighting conditions, and (ii) epiphytic growth along the sides of the tanks. The design of this experiment thus monitored the recovery of the benthic environment following organic enrichment and chemotherapeutic treatment.  1.3.3 SAMPLING DESIGN  Initial samples were taken before the fish waste and oxytetracycline were added (on day 0). Subsequent samples were taken on the 2nd, 4th, 7th, 14th, 21st and 28th day following organic enrichment of the benthos. To avoid biased samples due to edge effects, sampling was completed along a series of 7 transect lines, radiating outwards from the centre of the tank and arranged equi-distance from each other (Fig. 1.3). Four cores were taken per tank on each sampling day for (i) redox measurements, (ii) bacterial analysis, (iii) total carbon and nitrogen and OTC, and (iv) microfauna (ciliates). These four cores were taken from one transect line, with the transect line and the order of the cores along the transect randomly determined for that day and maintained consistently for all the tanks. Reduction-Oxidation Profiles (Eh)  A reduction-oxidation (redox) depth profile was taken by inserting a 3 ml syringe (i.d. 0.85 cm, area sampled 0.567 cm 2) into the sediment to obtain Eh readings from an undisturbed core sample (Fig. 1.4). The rubber plunger ring was removed and fitted with a thin glass tube, to which was attached a short piece of rubber tubing with an autoclave clamp. Silicone sealant applied around the glass and rubber ensured an airtight seal and the rubber ring was refitted onto the syringe. A series of 1 mm diameter holes were drilled through one side of the syringe, at 0.5 cm intervals. Before the syringe corer was inserted into the sediment, parafilmTM was tightly wrapped around the corer to cover the holes, and the clamp was opened. After the corer was pushed into the sediment, the clamp was closed  13  Transect Lines for Sampling  38 cm  Figure 1.3. A schematic of the 7 transect lines used for randomly sampling the benthos, with the central drainage pipe. The sediment cores would be taken along the transect line.  14  3 mm i.d. Surgical Tubing  Autoclave Clamp  Glass Tube  17-71-^ Rubber Plunger Ring iE37 •-^• lmm Hole • • • • • • • •  Luer End of Syringe Removed^  •  +  0.5 cm Spacing  —  Figure 1.4. A 3 cc syringe-corer, modified to obtain Eh depth profiles at every 0.5 cm.  15  and the resulting vacuum within the syringe acted to maintain the sediment in position within the corer as it was withdrawn from the sediment. To prevent the sediment from slipping out of the corer during Eh measurements, another piece of wax paper was wrapped around the bottom end of the corer. The tip of the electrode was then inserted through the wax paper and holes of the syringe to obtain the Eh reading. Readings were taken at 0.5 cm above the sediment, then at the sediment-water interface, and at 0.5 cm depth intervals thereafter. Although the Eh profile has been shown to change with 5. 1 mm intervals (Reimers et al. 1984, Revsbech et al. 1989), the size of the core and electrode tip-width did not allow measurements at smaller intervals without excessive disruption of the core. The redox potentials were measured with an Accumet Model 320 pH meter fitted with a Platinum Combination Electrode (Fisher #13-620-82, electrode tip approximately 1 cm long, 0.8 mm in width), and calibrated with ZoBell's solution prior to, and during, measurement of the redox potentials (ZoBell 1946). This solution was used primarily to prevent electrode poisoning and secondarily to clean the tip (Whitfield 1969), and it was composed of 0.0033 M potassium ferrocyanide to 0.0033 M ferricyanide in 0.10 M potassium chloride (e.g. ZoBell 1946, Pearson & Stanley 1979, Brown et al. 1987). Dissolved Oxygen Water overlying the sediments was measured for dissolved oxygen by colourimetric analysis (Parsons et al. 1984a). This measurement of oxygen can be taken as a rough indication of the changes occurring in the sediment chemistry due to the biological oxygen demand (BOD) of the sediments, as the exchange of gases (oxygen and carbon dioxide) occurs between the sediment surface and water column (Pearson & Rosenberg 1978; Brown  et al. 1987). This measurement was expressed as the percent saturation of dissolved oxygen in seawater. Protozoans  Another 3 ml corer was used to collect the sediment, using a PercollTm-sorbitol gel (Sigma Chemical Co., St. Louis, MO) and centrifugation technique (Schwinghamer 1981,  16  Alongi 1986). This relatively fast technique has been shown to have better than a 90% extraction efficiency from muddy and sandy sediments for most microfauna. Due to the tendency of Percollim-sorbitol to gel in the presence of formalin or Lugol's solution (personal observations, Alongi 1990), the Percollml-sorbitol was drained by filtration through 8 gm filters at low pressures (1/2 atm), and the organic matter remaining on the filter was then washed into 20 ml scintillation vials with a 2% formalin-seawater solution (Bullough 1962). Bacteria The core for the bacterial sample was taken with a 1 ml syringe (i.d. 0.4 cm, area sampled 0.126 cm 2 , volume 0.628 cm 3 ), with the rubber stopper ring modified as for the Eh corers. The sediment was preserved with 3.6 ml of 0.22 Am filtered artificial seawater and 0.25 ml of 37% formaldehyde (final concentration of 2.1% formaldehyde), and it was analyzed for bacteria after two weeks, as outlined in Appendix A. Carbon:Nitrogen Ratios A separate core from a 3 ml syringe (with the rubber stopper ring modified accordingly) was taken and immediately frozen at -12°C. The core was then thawed and mixed well, and a small subsample was taken (<20 mg) for analysis of total organic carbon and nitrogen. The remainder of the sample was used for analysis of oxytetracycline. The subsample of sediment was dried at 60°C for 3 days, ground to a fine powder, and then Vanadium Pentoxide V205 (BDH Ltd., Poole England) was added to the sediment samples before being subject to an autoanalyzer (Carlo Erba, Model NA 1500) for determination of total carbon and nitrogen. Oxytetracycline (OTC)  The remaining sediment from the previous core was used for analysis of OTC, following a highly sensitive method (Jacobsen & Berglind 1988) outlined in Appendix 2. An external standard was calculated by dissolving 100 mg of OTC in 10 ml of methanol (2x10 -4 M). Analysis of samples was on a Hewlett Packard system using a 1084A liquid  17  chromatograph, 1030B variable-wavlength UV detector operated at a wavelength of 350 nm and connected to a 79850A LC terminal. The volumes of the samples that were analyzed had to be increased to 200 pi to obtain quantifiable readings.  1.3.4 DATA ANALYSIS  A repeated measures ANOVA was used in the analysis of data followed by a Tukey Multiple Range Test when appropriate, using the statistical package SYSTAT (Ver 5.0, Evanston, IL). When the assumptions of equality of variance and normality were violated and transformations were unable to correct the problems, the non-parametric KolmogorovSmirnov Two-Sample Test was used.  1.4 RESULTS  1.4.1 SEDIMENT COMPOSITION  Grain sizes of the sediment were determined by sieve analysis (as in Harrison 1981). A  sample of sediment was dried at 80°C for 5 days before being sieved through a nested  series of Canada Standard sieves (The W.S. Tyler Co. Ltd., St. Catharines, Ont.) of 595, 355, 180, 75 and 53 Am mesh diameter. The sediment retained on each screen was weighed and the weight of the subsieve material was estimated by the difference (Fig. 1.5). Based on the Udden-Wentworth scale, the sediment composition may be described as being of fine to medium sand (75 - 355 Am) with a large silt (or mud) component (<54m) (Buchanan 1984).  1.4.2 VISUAL OBSERVATIONS  The sediment surface was initially a grey colour and devoid of benthic epifauna. By day 2, a smooth white microbial film had begun to develop on the sediment surfaces of the Control tanks, whereas the sediment surfaces of the tanks in Treatments A and B were  18  40 35 -  o g) o  30 25  LS 20 -4a) c) (1)  -  15  (i) 10  -  5 0 <53^53-75^75-180^180-355^355-595^>595  Sediment Size Range (urn)  Figure 1.5. Grain size composition by relative percentages of sediment obtained from 30 to 40 ft depth near Spanish Banks/U.B.C. The major constituents were the fine to medium sand at 31% (from 180-355 Am), followed by the very fine sand at 29% (from 75-180 Am), and the coarse silt/mud at 26% (< 53 Am).  19  beginning to turn a dark brown colour. Numerous macrofauna, not quantified but mainly comprised of large polychaetes and a few brittlestars, appeared on the sediment surfaces of the tanks in all treatments. By day 4, the white microbial mat covered approximately three-fourths of the sediment surface of the tanks in the Control treatment. Only small patches of this microbial film had begun to develop in the tanks of Treatment A, and the bacterial layer was even less developed in the tanks of Treatment B. All of the macrofauna had disappeared from the sediment surfaces, and were assumed to have re-entered the sediment. An attempt was made to clean the tank walls of epiphytic growth that had appeared at this time, but due to the small volume of the tank and the physical agitation of the water column, the white microbial film developing on the sediment surfaces was easily disrupted, and further efforts to clean the tank walls were not made. By day 7, the microbial film had declined in all tanks, only partially covering the sediment surfaces of tanks 2 and 3 in the Control treatment and with very little appearing in tank 1. This patchy distribution of the white microbial layer was also prevalent in the tanks of Treatments A and B. The sediment surfaces in tank 1 of the Control remained a dark brown by day 14, while tanks 2 and 3 were again covered by a white film. This microbial mat was also well developed in all tanks of Treatment A, but was only partially present in tank 2 of Treatment B. The sediment surfaces of tanks 1 and 3 of Treatment B were greybrown in colour, but the white mat began to be re-established on the surfaces of these tanks by day 21. The colour of the sediment to a depth of 4 cm in Treatments A and B was a very dark brown to black, with a black layer in the top 2 cm. The colour of the sediment in the Control was a light brown at depth, but also with a 2 cm thick black surface layer. After day 21, the microbial layer began to disappear from the sediment surfaces of all tanks, but remained for a longer period of time and disappeared slower in Treatments A and B. By the final day of sampling, day 28, the sediment surfaces of all tanks were a light to dark brown colouration. Patches of the white microbial mat remained in some of the  20  tanks of Treatments A and B, but only in areas of the sediment disturbed by previous sampling.  1.4.3 REDOX PROFILES  (Eh)  The reduction-oxidation (redox) potential is a quantitative measure of the energy of the electron escaping tendency of a reversible system. This can be measured with an electrode system as Eh in millivolts (mV, hydrogen scale), E being the potential difference between the standard hydrogen electrode and the system in which the redox potential is being measured (Zobell 1946, Whitfield 1969). The redox profiles exhibited by the tanks of the Control treatment are all similar in pattern (Fig. 1.6). Organic enrichment of the sediment caused strongly reducing conditions by day 2 at the sediment surfaces and at depth, as indicated by Eh values < +100 mV. This point is the lower boundary of free oxygen in the oxidized zone (Jorgensen & Fenchel 1974) below which the sediment is reducing (Fenchel 1969, Jone 1979, 1981, Harrison 1981). Redox conditions were most reducing on day 4 at -160 mV and -318 mV for replicate tanks 1 and 2 respectively, and on day 21 at -316 mV for tank 3; thereafter the sediment became less reducing overall. Reducing conditions initially increased with sediment depth before stabilizing, but after day 14 in tanks 2 and 3 the redox profile began to become less reducing at depth - this condition never occurred in tank 1. A redox potential discontinuity (RPD) layer was always present between 0 to 1 cm beneath the sediment surface, a zone where an abrupt change in Eh values occurs with depth. Redox values at the sediment surfaces at day 0 in tanks 2 and 3 of Treatment A indicated that oxidizing conditions were present, with the RPD layer extending down to 2 cm and becoming very reducing at this depth (Fig. 1.7). By day 2 after the addition of fish waste and oxytetracycline, the sediment surfaces had become reducing. Redox profiles were similar to those in the Control treatment, with a low Eh of -285 on day 4 measured in tank 1, -235 on day 28 in tank 2, and -200 on day 21 in tank 3. The sediment became less  21 Eh (mV) N  N  0^0 0 0 1.0^NI^0^0^ 0^0^N OD^V.^0^et^CO 7^7^ I^I  N  0 N  N  0 CNI  I^I^I^I^I^I^t^I.^I^I^1^I^I^I^5^I^J^I  I^I  1.0  8 N  0.5  :CI.. • - 11  0.0  -  Replicate 1  6  0.5 1.0  Day 0 not available • Day 2 ^ Day 4 ^ Day 7 O Day 14 • Day 21 • Day 28  ;t? 0 V  70 • • 4 Elr' ,4 4  1.5 2.0  ‘  ,•■ k7^v  2.5 3.0  ib  3.5 -  s  4.0  N  O ° N  4.  N  0^0.^0^2 cv^o^0^c.,^2^• 8^0^ ^•^0^•^0^(NN^N 7^7^I^I  ■^I ^.II.I 1.0  N  ^I^1.1,1■I■1■1  0.5 0.0 0.5  --  7  0  ---------  0  v4  a  1.0 1.5  Day 0 not available • Day 2 7 Day 4 Day 7 0 Day 14 • Day 21 Day 28  .,  0  2.0  a'  2.5  0•4  .  ,  • ,nia^•  3.0  0•  3.5  Replicate 2  4.0 C^o^o^0^0^  0  0^0^0^0^0^0 CO^et^0^CD^N^0^0^ 0^0^N^LO^0^et^CO^N N^N^N N^N^NCO^ea.^0^et-^CO^ I^I^I^I^  7^I^I  1^i^,^I^1^I^I^i^,^I^,^i^,^I^.1  1.0 0.5 0.0 0.5 1.0 1.5 2.0 2.5 .3.0 3.5 4.0  •  - - ----------- -----•  -------^ - -----  0  Replicate 3 --.2, •  -  44 't 0  • 41 0 v 04 • is  •  Day 0 not available • Day 2 ^ Day 4 • Day 7 O Day 14 • Day 21 • Day 28  Figure 1.6. Reduction-oxidation time-series depth profiles of the replicate tanks for the Control Treatment in 1990. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment surface.  ▪ 22  n  Eh (mV)  O N^N^N^  .^  0 0  N^ V^O  ^  I^I^I  I^I  O b  O  O  N  N  O  N  O M  O M  1^1^1^1^1^.1^•^1^1  1.0 0.5  -a  0.0  _..p• C7 -  Replicate 1 Day 0 not available • Day 2 ^ Day 4 ^ Day 7 ^ Day 14 • Day 21 p Day 28  0.5 1.0  v,.......-.^•^o rri off, V\  1.5 2.0 2.5 3.0 3.5 4.0 O  N  N  N N  o^N^ O .-^m  O^ N^  I^I^I^I  0 1  O  o o m  N^0 10  8 N  N  8N  N  N  1.0 0.5 0.0 0.5 1.0  Day 0 Day 2 Day 4 Day 7 Day 14 Day 21 Day 28  1.5 2.0 2.5 3.0 3.5 4.0 0  n 1.0  0  m  a^  0^0 0 N^N  I^I  t0^  N^00^0^O^o  I^  I  I  m  0  0 O  0 N  0  N  0 N  ^i^l^l^I^1^i•^i^i^i^•l^1  0.5 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 -  Replicate 3 o Day 0 • Day 2 V Day 4 V Day 7 ^ Day 14 • Day 21 L Day 28  4.0  Figure 1.7. Reduction-oxidation time-series depth profiles of the replicate tanks for Treatment A in 1990, a low dosage of oxytetracycline. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment.  23  reducing after day 21 in tanks 1 and 3, but remained highly reducing on day 28 in tank 2. In comparison to the Control treatment, the sediment surfaces of Treatment A did not become reducing as quickly and recovered earlier; the sediment remained reducing at depth for a longer period of time and was also less reducing overall. Redox potentials on day 0 were^+200 mV (Fig. 1.8), much greater than in the tanks of the Control and Treatment A, with reducing conditions reached by 1 cm depth. After the addition of fish waste and a higher dosage of oxytetracycline, the redox profile of the sediments closely resembled the profiles in Treatment A. Overall, the sediments progressed to less reducing conditions, with the lowest surface Eh of -120 mV on day 7 in tank 1, and -120 and -170 mV on day 4 in tanks 2 and 3 respectively. This was also when the overall redox of the sediments was the lowest, at -260 mV in tank 3. Although the sediments in this treatment remained reducing for a longer period, and were less reducing overall than in the Control treatment (similar to Treatment A), the sediment surface recovered sufficiently to become oxidized, which never occurred in either the Control or Treatment A.  1.4.4 DISSOLVED OXYGEN  Changes in the levels of dissolved oxygen of seawater were not significantly different among the three treatments (ANOVA and Tukey test, P > 0.05, Fig. 1.9). Dissolved oxygen levels were initially 97% saturation of seawater, and declined rapidly by day 7, to fluctuate between 88% to 78% saturation of seawater for the remainder of the study. Analyzed on a daily basis, only the decrease in dissolved oxygen on day 2 for the Control was significantly different from Treatments A and B (Tukey test, P <0.05).  1.4.5 CARBON:NITROGEN RATIOS  The carbon:nitrogen (C:N) ratios of all treatments varied in range from 13.5 to 19 (Fig. 1.10). The C:N ratios decreased in value from 17-24 to 13.5-15 by day 2 after  24 Eh (mV) 0  N  0 03  0 0 0 0 *^0^ N  .-^CO^  0.5  g^0^CO0  0  0  0 0 0 ^ CO 0 N  I^I^I^I^I^I^I^1^I^I^I^I^I  I  1.0  0  -  0.0 •-?- 0.5 -0-1  1.0 '  0^1.5  -  4i:^2.0  -  2.5 'V kl)^3.0  v)  -  3.5 4.0 ^ 0 tet  Ft'  N /^I  I^I  1.0  N  8  I^I  0  0 CO  0  O  I^I^I^I  0  0  I  0^0 O I  0  I^I^I  0.5 0.0 0.5 Q.  Replicate 2 0 Day 0 • Day 2 Day 4 Day 7 ^ Day 14 ■ Day 21 o Day 28  1' 0 -  A^1.5  -  2.0 2.5 0)^3.0 3.5 4.0 ^ 0  N  0 0 0 *^0^l0^N^0^0^0^0 00^ 0^0  N^N^N^  I  U  a.  7^7^1CO^*^0^ ^1  *^03  0^0^0^0^0 CO  0  N^N^N^■,)  I^I.^I^I^1^I^,^I^,^I,^1^■^I^i^,i^,^i^,^I  1.0 0.5  ^  -  0.0 0.5 10  -  N  cz^1.5 -  TT:^2.0 g  2.5 -  .b l'^3.0 ril  -  3.5 4.0 ^  Figure 1.8. Reduction-oxidation time-series depth profiles of the replicate tanks for Treatment B in 1990, a high dosage of oxytetracycline. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began at 0.5 cm above the sediment.  25  1 00 95  -  90-85 : 80 : 75 -70 :  o Control • Treatment A  65 --:  v Treatment B  60 --55 50  1^I^I^I^I^1^1'1'1^1'1'1'1 , 1 0^2^4 6 8 10 12 14 16 18 20 22 24 26 28 30 Time (Days)  1  Figure 1.9. Comparison of the dissolved oxygen levels in seawater between the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3.  26  38 — 36 — 34  o Control • Treatment A v Treatment B  c 32  o  "-«-1^30 cz X 28 co 26 tali 0  —  s. 24 — ...) ..-‘^22 —  o 20 ,c)  s.^18 — ccs t.) 16 — 14 — 12 — 10 — 8  I  ^^^ 1 1 1^1'^1'^1'^1'^1^1^1^1^1^f^1^1  0^2^4^6^8  10 12 14 16 18 20 22 24 26 28 30  Time (Days)  Figure 1.10. Carbon:nitrogen ratios of the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3.  27  addition of fish waste, before increasing slightly by day 28. None of the treatments were significantly different from each other (Kolmogorov-Smirnov test, P < 0.05). For comparison to previous studies, the total carbon content of the sediments is presented as well (Table 1), and shows that the carbon content of the sediments increased after the addition of the fish waste.  1.4.6 OXYTETRACYCLINE  Levels of OTC declined quickly after addition to the sediments, from 260 ppm and 1080 ppm to < 1 ppm and 120 ppm by day 2, in Treatments A and B respectively (Fig. 1.11). OTC levels continued to decline slowly in both treatments for the remainder of the study period. After day 14, trace levels of OTC remained in the sediments of Treatment A, but could no longer be detected. OTC concentrations were significantly different between the two treatments (P >0.05, Kolmogorov-Smirnov test).  1.4.7 MICROBIAL ABUNDANCE  Following addition of fish waste to the Control tanks, bacterial abundance decreased from 9x109 cells. g-1 to 5.5x109 cells. g-1 sediment on day 4, before increasing to 22x10 9 cells•g-1 on day 21 and declining again on the final day of the experiment (Fig. 1.12). A relatively low dosage of oxytetracycline with the fish waste to Treatment A caused an initial rise in bacterial numbers to 12x10 9 •g-1 sediment by day 2, followed by an immediate decrease in abundance on day 4 as in the Control tanks. Changes in bacterial abundance were similar to the Control treatment, increasing to the highest level on day 21 (to 20x109 .g-1 ) and declining again. The change in bacterial abundance between Treatment A was not significantly different from the Control (P >0.05, Tukey test). After the high dosage of OTC in Treatment B, bacterial cells increased rapidly to peak at 26x10 9 •g-1 sediment on day 4, and dropping to about 14x10 9 • -1 on day 7, remaining near this level  28  1200 1100 1000  a)  900 300  o Treatment A • Treatment B  c.) 250 200 a., 150  1 00 50 --0  -  1!Ir^  I^'^I^'^I^1^'^I  0^2^4^6 8 10 12 14 16 18 20 22 24 26 28 30  Time (Days)  Figure 1.11. The concentrations of oxytetracycline (ppm) for Treatment A (2.53 g of OTC) and Treatment B (10.13 g of OTC). Plotted values are the Mean ± 1 Standard Deviation, n=3.  29  40  °-)  30  -  o Control • Treatment A v Treatment B -  sa)  cn  20  -  o  °  U'  10  0  I^I^I^I^I^I^1^1^I^I^I^I^1,1  0^2^4^6^8 10 12 14 16 18 20 22 24 26 28 30 Time (Days)  Figure 1.12. Bacterial abundance (cells x 10 9 .g -1 sediment dry weight) between the Control (fish waste added), Treatment A (fish waste and 2.53 g of OTC added), and Treatment B (fish waste and 10.13 g of OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3.  30  for the remainder of the study. The high dosage of OTC resulted in significant changes in bacterial abundance between Treatment A and the Control treatment (P <0.05, Tukey test). Bacterial were initially small, coccoid and short rod shaped cells. Chains composed of larger rods occurred during the first several days of the experiment, but were no longer noticeable by the end of the experimental period.  1.4.8 PROTOZOAN ABUNDANCE  Although the PercollTM-sorbital gel was vacuumed off, the remaining gel on the filter still coagulated with the 2% formalin-seawater preservative solution, rendering all samples unidentifiable.  1.5 DISCUSSION  1.5.1 REDOX MEASUREMENTS  The development of reducing conditions in the sediments by the addition of fish waste was clearly demonstrated by the changes in the reduction-oxidation profiles (Eh) in the tanks of all 3 treatments. The redox potentials < +100 mV at the surface and at depth, concurrent with the shift in the RPD layer to the surface, were indicative of reducing conditions being rapidly reached (Fig. 1.6 - 1.8). As oxygen is depleted within a short time in the surface zones of the sediments, a series of anaerobic oxidation processes occur, in the general sequence of nitrate reduction, sulphate reduction and methane formation (Zehnder & Stumm 1988, Legal & Fauque 1988). The energy released during the redox reactions can be measured (Eh), and the intensity of the reducing conditions can thus be related to the redox potential (Fig. 1.13). Microbially mediated reduction reactions thus occurred below redox potentials of -100 mV to -200 mV (e.g. Battersby & Brown 1982, Stanley et al. 1981), in several genera of sulphate reducing bacteria (e.g. Desulfovibrio, Desulfotomaculum, Desulfobacter, Desulfonema; Postgate 1979, Nedwell 1982, Parsons et  31  Redox Potential^-AG°(pH 7) [mV^ [kJ/mot el [CH 20] C O2  - 600  -20  [H]  D  -400 \ -----:'` NADH -320 NA D FADH FAD-'  -200  red. Cyt b )+ 3 0 ox .  -420  H2  H* CH 4 CO Fe"  HS  s of -  Fe 3 *  +20 +40  red.^+200 Cyt ox.-1 +230' ENH4  red. Cyt  0  +60  NO; ( Mn  +600 + 800 +1000  2.  +80  Mn`' 7 0 ( N2 810^NO (- H2° L 02  +10 0  +120  Figure 1.13. Electron free energy diagram for the biologically mediated redox sequences of the degradation of organic carbon, CH2O (acting as the electron donor). Note the relationship between the change in energy (G°), redox processes and the Eh (mV) (from Zehnder & Stumm 1988).  32  al. 1984b, Widdell 1988, Fauque et al. 1991, Bak & Pfennig 1991). This was reinforced by visual inspection of the sediment cores, which revealed that the upper 1 to 2 cm of the sediment surfaces were black, becoming grey-brown at depth. The black colouration was likely due to the presence of ferrous sulphide (FeS), and the FeOOH-FeS zone is associated with the brown and black colour change in the sediments (Jorgensen 1989, Jorgensen & Revsbech 1989). The values of the redox potentials measured in this study were within the ranges found in sediments beneath fish farms in a Scottish sea loch by Brown et al. (1987) (as low as -185 mV at 4 cm depth), and down to -200 mV within 2 cm of the sediment surface by Lumb & Fowler (1989). The sediment was less reducing in tank 1 of the Control treatment, compared to tanks 2 and 3, due to the inadvertent mixing of the sediments and fish waste at the time of addition (day 0). This mixing aerated the sediment and created a less reducing environment.  1.5.2 MICROBIAL MAT  The white microbial mat that developed on the sediment surfaces was composed of the non-phototrophic sulphide oxidizing filamentous bacteria, primarily Beggiatoa spp.  (e.g. Jorgensen & Fenchel 1974, Jorgensen 1977a, Nelson & Castenholz 1981, Jorgensen & Revsbech 1983 & 1989, Nelson et al. 1986, Kuenen 1989, Lumb & Fowler 1989) oxidizing the large amounts of H2S produced in the sediments during sulphate reduction. The presence of anoxygenic phototrophic bacteria is doubtful, as light was excluded from the tanks by dark covers except when sampling (Caumette 1989). The rapid development of the microbial mat was another indication of how quickly the sediments had become anoxic and reducing. The decline of the white microbial mat on day 7 was caused by wiping down the tank walls on day 4, which resulted in mild aeration of the sediment surfaces. But anoxic and reducing conditions redeveloped quickly as the microbial layer rapidly reformed.  33  Based solely on morphology, the small cocci and short rod shaped bacterial cells that were prevalent throughout the study are characteristic of sulphate reducing bacteria (Widdel 1988, Fauque et al. 1991), while the larger rods that comprised chain and colonies are characteristic of the chemoautotrophic white sulphur bacteria, Beggiatoa spp. (Nelson & Castenholz 1981, Jorgensen & Revsbech 1983). Benthic infauna surfaced from the sediments after day 2 to escape sulphide poisoning in the sediments (Jorgensen 1977b & 1980), and had presumably returned to the benthos by day 4, although burrows were not observed. Previous observations from field studies have noted that the presence of the white microbial mat did not exclude all epibenthic meio- and macrofauna, with sea anemones and geoducks present beneath fish pens (Aquametrix 1988). Under extremely reducing conditions, even these organisms will eventually die (Jorgensen 1980).  1.5.3 DISSOLVED OXYGEN All three treatments exhibited the same pattern of changes in the dissolved oxygen of seawater due to the BOD (Fig. 1.9), which has been consistently found for bottom water from sediments loaded with fish waste (Hall & Holby 1986, Brown et al. 1987, Gowen & Bradbury 1987, Parsons et al. 1990). There was a significant decrease in the dissolved oxygen levels of the Control treatment on day 2 compared to Treatments A and B. This may be attributed to the greater reducing conditions that were reached more rapidly by the sediments of the Control treatment, perhaps as the microbial population was not inhibited in some way by the OTC as in Treatments A and B.  1.5.4 CARBON:NITROGEN RATIOS  The carbon content of the sediments prior to addition of fish waste varied between 1 to 2% (Table 1). This was within the range expected for estuarine sediments along Vancouver Harbour (Johnson 1990). Enrichment with fish waste increased carbon content,  34  TABLE 1. Total Carbon content (percent) of the sediments in the Control, Treatment A and Treatment B. Values presented are the Mean % ± 1 Standard Deviation (SD). Day 0 2 4 7 14 21 28  Control SD Mean % 0.255 1.529 1.522 0.380 0.372 1.808 0.272 2.075 1.523 0.239 1.596 0.276 0.126 1.555  Treatment A SD Mean % 0.672 1.517 1.749 0.801 2.689 1.218 0.449 1.543 2.118 0.619 1.806 0.704 1.646 0.252  Treatment B SD Mean % 1.849 0.909 1.956 0.382 2.112 0.330 1.839 0.295 1.213 0.437 0.291 1.526 2.282 1.113  35  with values in the range of 1.5 to 2.7%. In comparison, the carbon composition of sediments found by Brown et al. (1987) directly beneath fish farms was as high as 9.35%, but averaged 4% composition 15-1400 m away from the farm sites. However, the sediments of this study were characteristically low in organic matter due to the quartz origin of the sediments and tidal action (Carolyn Jones, pers. comm.). Comparisons of the total carbon content of the sediments between this study and others should be made cautiously  (e.g. Brown et al. 1987). In field studies, phytodetritus would have accumulated in the sediments throughout the year, in addition to the fish waste released from the culture operations (e.g. Aquametrix 1988). Only fish waste was added to the sediments in this study, and insufficient phytodetritus would have accumulated in the microcosm tanks after only a few weeks for the carbon content of these sediments to resemble conditions in the field. The decrease in the C:N ratios of all treatments reflected the lower total carbon and nitrogen of the fish waste (Fig. 1.10). OTC did not affect the degradation of fish waste, as the C:N ratios did not differ between treatments. The C:N ratios were similar to those found by Samuelsen et al. (1988), of 10 and 14 at fish farm sites in Norway. The C:N ratios at the end of the study remained approximately the same as when the fish waste was initially added, suggesting that any changes in C:N ratios would require a longer duration than 4 weeks to become noticeable due to biological degradation processes by benthic fauna. Fish wastes are readily degradable by meio- and macrofauna (Weston 1986 & 1990, Tsutsumi 1987), and are thus assumed not to be refractory.  1.5.5 OXYTETRACYCLINE  The sediments with oxytetracycline added to the fish waste appeared to experience less reducing conditions (Fig. 1.6 - 1.8), although the sediments remained reducing overall for a greater period of time at depth, with oxidizing conditions reached at the sediment surfaces by the end of this study. Sediments with fish waste and OTC have been known to  36  remain reducing for a longer period of time (Samuelsen et al. 1988, Jacobsen & Berglind 1988, Bjorklund et al. 1990), but this differential effect at the sediment surface and at depth has not been previously recorded. This increase in the Eh of the Control treatment to a less reducing environment suggests that there was a shift in the microbially mediated degradation processes, away from sulphate reduction. The large decreases of OTC in Treatments A and B (Fig. 1.11), were due partly to the OTC being washed out of the fish waste during addition to the sediments, rather than being degraded by bacterial action (Samuelsen et al. 1988, Samuelsen 1989, Jacobsen & Berglind 1988, Husevag et al. 1991). But as discussed in Appendix 2, the oxytetracycline was also underestimated in the sediments because of problems experienced with the methodology. Even though analysis of the OTC samples was incomplete, the small concentrations of OTC would likely have persisted for a long period of time,^419 days under conditions of sediment anoxia and low turbulence in the water column (Bjorklund et al. 1990). Samuelsen et al. (1988) theorized that the development of the white sulphur oxidizing bacterial plaques on the sediment surfaces may have decreased diffusion of the oxytetracycline into the water column. As oxytetracycline is a hydrophilic compound with an octanol/water partition coefficient of 0.025 (Chopra 1985), it should tend not be incorporated into fish waste, and accumulate in the sediments (unlike the behaviour of hydrophobic organic pollutants, review by Jaffe 1991). This is offset by the preferred accumulation of organic pollutants in muddy and fine sediments that are high in organic carbon (Hiraizumi et al. 1979), while high amounts of dissolved organic matter (DOM) enhances the solubility of pollutants in the aquatic environment (Jaffe 1991). This suggests that the accumulation and persistance of OTC in the sediments beneath a fish farm would occur under high rates of sedimentation. Chemical degradation of OTC by photochemical decomposition can occur (Oka et al. 1989), but would be unlikely at the water depths of the sediments beneath fish farms where light is greatly reduced. Redox reactions and biodegradation by microorganisms of OTC  37  under highly reducing conditions are not known (Lunestad & Goksoyr 1990), and the disappearance of OTC over time due to diffusion and washing out from the sediments is currently the accepted explanation.  1.5.6 MICROBIAL ABUNDANCE  Total bacterial counts were within the range found for sediments beneath fish farms, 0.3 - 1.4x109 •g-1 (Enger et al. 1989), 1.4 - 6.3x109 •g-1 (Samuelsen et al. 1988), 1.4 2.2x109 .g-1 (Torsvik et al. 1988), and 3.7 - 10.1x109 .g-1 (Carr & Goulder 1990). Total cell counts in this study were to 2 to 10 times higher (e.g. 26x109 .g-1 in Treatment B) than in the previous studies which had not utilized ultrasonication to release all the bacteria attached to the sediment. The large standard deviations present (Fig. 1.12) were due to a combination of patchiness, sampling and subsampling variability that often occurs with benthic analysis (Venrick 1971, Montagna 1982). From the strongly reducing conditions present in the sediments (Fig. 1.6 - 1.8), the large increases in bacterial abundance on Days 2 and 4 of Treatment B (Fig. 1.12) were likely due to the sulphate (and nitrate) reducing fractions of the microbial population (Battersby & Brown 1982). Increases in the abundance and activity of sulphate reducing bacteria have been associated with increases in redox potentials (Jones 1979, Aller & Yingst 1980, Jorgensen 1989). Sulphate reduction in marine benthos can be the principal pathway controlling the degradation of organic matter (e.g. Jorgensen & Fenchel 1974, Pearson 1982, Battersby & Brown 1982, Thode-Anderson & Jorgensen 1989, Sampou & Oviatt 1991). Stanley et al. (1981) examined sediments loaded with organic waste from a pulp mill in Scotland, and concluded that while total aerobic/anaerobic heterotrophic bacteria and nitrate reducing bacterial cells did not increase in abundance, sulphate reducing bacteria were 2 to 3 times higher than in unpolluted regions. Samuelsen et al. (1988) also found sulphate reducing bacteria to be 2 to 4 orders of magnitude more at fish farm sites (104 and 106), compared to unpolluted sites (10 2).  38  The cell counts in all of the treatments (Fig. 1.12) did not begin to converge until the sediments began to become more oxidized by day 28. The overall increase in bacterial abundance on day 21 in the Control and Treatment A was due mainly to Beggiatoa spp., as day 21 corresponded to the maximal development of the white microbial mat.  1.5.7 POTENTIAL EFFECTS OF OTC ON MICROFAUNA  The smaller dose of OTC (1.27 •' 1 fish waste) resulted in a small increase in bacterial cell counts on day 2 in Treatment A (12x109 bacteria•g -1 sediment, Fig. 1.11), while the larger dose of OTC (5.07 g•1 -1 fish waste) resulted in a large increase in cell counts on day 2 and 4 in Treatment B (to 26x10 9 bacteria•g 4 sediment). Samuelsen et al. (1988) also found total bacterial counts to be 1.4x10 7 • -1 at a fish farm site abandoned for a year and without chemotherapy for 18 months, and 1.9x10 7 •g -1 at another fish farm site during a disease outbreak treated with antibiotics. These cell counts were higher in comparison to a control site where the bacterial counts were only 4.3x10 5 .g -1 . The large increases in microbial abundance during the first few days of the experiment in Treatment B, the slower development of reducing conditions within the sediments and of the white microbial mat (Section 1.4.2), and the persistence of the white microbial mat at the end of the experiment can be explained by the inhibition of benthic bacterivorous protozoans and the development of OTC resistance. Although only bacterial abundance and the redox state of the sediments was conclusively established in this preliminary investigation, the possibility of OTC resistance developing and the inhibition of the microbial and protozoan communities (and interactions between these microfauna) should also be considered. The cell counts were higher in Treatment B with the higher dosage of OTC, than in Treatment A and the lower dosage of OTC, despite the presence of the white microbial mat in the Control treatment and Treatment A. This would suggest that the increases in cell numbers were in another fraction of the anaerobic bacterial community, perhaps due to a  39  combination of OTC resistance developing and the potential inhibition of protozoan grazing.  Development of OTC Resistance The large increases in bacterial abundance in Treatment B, the overall increase in bacterial abundance despite the presence of OTC in Treatment A, and the stable bacterial population from day 6 to 28 in Treatment B suggested that there was a potential for OTC resistant strains to have developed among the microbial population. The frequent and long term use of OTC, has resulted in a majority of bacteria being resistant to this antibiotic (Levy 1984 & 1988), either from the sediments or in animals near fish farms (Aoki et al. 1974 & 1981, Austin 1985, Bjorklund et al. 1990 & 1991, Grave et al. 1990, Husevag et  al. 1991, Nygaard et al. 1992). The bacteria carry and transfer this resistance through plasmids (e.g. Chopra 1985, Terzaghi & O'Hara 1990, Al-Masaudi & Russell 1991), assumed to be mainly through conjugation and possibly transduction in aquatic environments (e.g. Trevors et al. 1987, Fernandez-Astorga et al. 1992). Torsvik et al. (1988) isolated 2 distinct strains of antibiotic resistant bacteria from sediments beneath a fish farm, with one strain susceptible to low concentrations of OTC (2.4 tig•m1 -1 ), and the other resistant to high concentrations of OTC (300 Ag•m1 -1 ). This double peak of resistance to OTC has also been reported in Aeromonas salmonicida (causing furunculosis), at 1.26 and 80 Ag•m1 -1 OTC (Inglis & Richards 1991). There is thus a possibility for the microbial community to have developed resistance in Treatments A and B. The minimum inhibitory concentration of OTC required to elicit an antibiotic resistant response in the microbial community was not measured. After the majority of the added OTC in the fish waste was washed out after addition to the sediments (Appendix B), the presence and activity of the OTC was probably diminished even further by complex formation with magnesium and calcium ions in seawater (Lunestad & Goksoyr 1990). The active form of OTC in the sediments and seawater can be as little as 5 %, with the OTC  40  bound mostly in the form of an OTC-magnesium complex, and some as an OTC-calcium complex. The microbial population could have been inhibited by the OTC or by the grazing pressures exerted by the bacterivorous protozoans.  Inhibition of Protozoan Predation The loss of all benthic ciliate samples precluded examination of protozoan abundance. However, the oscillations in bacterial abundance (Fig. 1.12) are characteristic of grazing effects by bactivorous protozoa (e.g. Tietjen 1980, Berninger et al. 1991, Weisse & Scheffel-Moser 1991). Protozoa are abundant in sewage (e.g. Pike & Curds 1971, Kinner & Curds 1987, Manru et al. 1988, Kosciuszko & Prajer 1990, Esteban et al. 1991), and inhibition of protozoa by antibiotics has been shown to increase total bacterial abundance (Mallory et al. 1983, Wiggins et al. 1987, Wiggins & Alexander 1988). The potential effects of OTC on protozoans and their grazing pressure should also be considered. The effects of OTC on benthic ciliates (Class Kaiyorelictea) indicated a chemotactic response (personal observations), with the ciliates attempting to move away from the antibiotic. At high dosages of OTC (e.g. 4 drops of 0.25 M OTC, or 1.24x10 4 mg OTC.1 -  1 seawater) these highly contractile ciliates exhibited a toxic reaction, swelling and bursting. In Treatment B, a sufficiently high and active amount of OTC could have been initially present to have inhibited the bactivorous ciliate population in the sediments. This could have been either by an immediate toxic effect, or a temporary cessation of proliferation followed by resumed growth at a decreased rate (Nilsson 1989). Tremaine & Mills (1987) determined that protozoans were inhibited by as little as 25 mg•1 -1 cycloheximide , with large protozoans inhibited within 24 hrs, and smaller protozoans active until 48 hrs. The dosages of OTC used in the Treatment A and B were initially 1.27 g•1 -1 and 5.07 g•1 -1 fish waste, respectively, perhaps sufficient to inhibit the protozoan populations. The toxic effect  41  of OTC on protozoans resulting in the organism bursting is not known, and requires further research. OTC is able to both enter eukaryotic cells and act on the mitochondria, and by implication OTC is thus able to enter protozoans (eukaryotes) as well. Oxytetracycline inhibits protein synthesis at the 70S ribosome (Levy 1984, Chopra 1985), and the mode of action in the mitochondria of ciliates may be analagous to the action of chloramphenicol (CAP, Chopra 1985) which also inhibits protein synthesis at the 70S ribosome. Studies on the ciliate Tetrahymena have shown that the action of CAP is confined to the mitochondria (Nilsson 1989), resulting in a decrease of mitochondrial DNA and volume, inhibition in protein synthesis and the synthesis of mitochondrial ribosomes. Recovery can be rapid, 6 hrs after CAP is removed, and different strains of Tetrahymena can tolerate higher levels of CAP and achieve cell doublings (Nilsson 1989). The mitochondrion in protozoans use 02 as a terminal electron acceptor to oxidize reducing compounds passing down the electron transport chain to produce energy (Finlay 1990). However, protozoans in reducing sediments that rely on anaerobic metabolic pathways for the release of energy do not have mitochondria, but contain endo- and ectosymbiotic bacteria and hydrogenosomes (Fenchel  et al. 1977). The symbiotic bacteria are methanogens, utilizing the H2 gas produced from the catabolism of pyruvate in the hydrogenosomes to produce methane that is expelled from the ciliate. The hydrogenosomes function like mitochondria, and yield 02, CO2, acetate and energy used by the protozoan (Finlay 1990). The action of OTC on the symbiotic bacteria and hydrogenosomes of anaerobic ciliates is not known, and additional research is required in this area. Although most of the OTC was rapidly washed out of the sediments (with low amounts of OTC persisting in the sediments, Fig. 1.11), the effects of the higher concentrations of OTC on the protozoan community could have extended over the first few days (day 2 and 4). During this period of ciliate inhibition, the microbial community would have responded rapidly by increasing in abundance. After day 4 the lower amounts of OTC  42  present (50 ppm) may not have been enough to inhibit the bactivorous ciliate population any further, resulting in the bacterial community being grazed down. The lower dosage of OTC in Treatment A may have only been enough to have had a small inhibitory effect on any ciliates present, which would account for the small rise in bacterial abundance on day 2 (and with the bacteria displaying OTC resistance). Similarly, Pratt & Cairns (1985) also found that a toxicant (cadmium) reduced numbers of photosynthetic and bactivorousdetritivores ciliates at low concentrations, and high toxicant levels eliminated photosynthetic species and reduced bactivorous-detritivore protozoan populations by one-half.  The potential effects of OTC within and between the microfaunal communities could have occurred as follows. Protozoan grazing on the sulphate reducing bacterial community could have been inhibited by a high dose of OTC. Concurrently, the greater increase in bacterial abundance indicated that growth of the anaerobic bacterial population was not inhibited by the OTC, perhaps due to the development of OTC resistance. Although bacterial abundance did increase, the slower increase in the reducing conditions of the sediments treated with OTC (Treatment B), along with the slower development of the white microbial mat at the end of the study showed that there was initially a lack of H2S available from the anaerobic bacterial community (i.e. sulphate reduction was not the dominant process in the degradation of fish waste). If some sulphate reduction had occurred, the H2S could have been bound in the sediments as FeS and FeS 2 (Jorgensen et al. 1990, Bak & Pfennig 1991, Kristensen et al. 1991) making it unavailable to the sulphur oxidizing bacteria, or the sulphur oxidizing bacteria may not have been able to use the H2S due to possible inhibition by OTC. The increase in bacterial abundance in Treatment B, but lack of available H2S for the sulphur oxidizing bacteria, suggests that while the sulphate reducing bacteria increased in abundance the metabolic rates of the bacterial community were inhibited, again through the lack of grazing pressure. Grazing by protozoans on the microbial community can stimulate bacterial production (Kemp 1990), which could have  43  occurred in the Control treatment, as indicated by the highly reducing conditions that were quickly reached and the rapid appearance of the white microbial mat. As well, any increases in the bacterial population could have been grazed down and never observed. Bacterial metabolic rates may not always be coupled to growth processes, and the increase in cell numbers on day 2 and 4 of Treatment B may not necessarily mean an increase in rates of bacterial activity (Hanson 1980, Pearson 1982). When the OTC washed out of the fish waste and sediments after the first day (and with the recovery of the protozoan community due to the potential inhibitory effects of OTC), along with the possibility of reduced grazing pressure on the sulphate reducing population but stimulated metabolic rates, sufficient H2S could have been produced for the sulphur oxidizing bacterial community to have become established, and with increased redox conditions in the sediments. However, these theories are only speculative, and cannot be conclusively demonstrated due to the lack of data on ciliate abundance and the inadequacy of studies on the effects of OTC on protozoan physiology and OTC resistance among the microbial community in this study. The greater reducing conditions at the end of the study, with the continued presence of the white microbial mat implied that H2S was still available from sulphate reduction (sulphate reducing bacteria remained active). In the Control treatment, the disappearance of the white microbial mat indicates that sulphate reduction was no longer the primary route in the degradation of fish waste (H2S was no longer available), and perhaps there was a subsequent switch to methanogenesis. In marine systems with high rates of organic matter deposition, sulphate can be depleted to the extent that methanogenesis becomes a significant pathway in anaerobic carbon catabolism (Capone & Kiene 1988). The composition of sedimentary gas at fish farms has been found to be composed of mainly methane (64%) and H2S (22%) (Kaspar et al. 1988). This is confirmed by Aure & Stigebrandt (1990) who estimated that in sediments strongly impacted by fish waste and experiencing reducing conditions from -200 to -320 mV, up to 70% of the biochemical decomposition should be  44  by methanogenesis, and the remaining 30% through sulphate reduction. Methane production would require a sedimentation rate greater than 8 g C•m 2 •d -1 beneath net pens (Iwama 1990). Nitrate reduction in the sediment is negligible under these conditions (Kaspar et al. 1990). Heterotrophic growth and activity would need to be measured in order to confirm these conjectures.  1.5.8 ENVIRONMENTAL CONCERNS  The impact of oxytetracycline on the microbenthos may result in two potential problems, the enhancement of disease transmission in the microbial community, and decrease in the mineralization rate of detrital carbon. Disease Transmission The persistence of low amounts of OTC in sediments ( 419 days), may cause a long-term antibiotic resistance response (Chopra 1985, Austin 1985), making it impossible to control bacterial diseases of cultured fish (Toranzo et al. 1984). Antibiotic resistance by bacteria has often been correlated to fish farms undergoing antimicrobial therapy, and resistance to the tetracycline antibiotics reached 69.1% among bacterial strains isolated from catfish farms with cross-resistance to the tetracyclines, chlorapmphenicol, kanamycin, ampicillin and nitrofurantoin commonly occurring (McPhearson et al. 1991). Nygaard et al. (1992) found that among bacteria isolated from sediments beneath fish farms, OTC resistance increased from 5% to 16% after 12 months. The presence of other fish pathogenic bacteria (see Sindermann 1984, Egidius 1984, Austin & Allen-Austin 1985, Cahill 1990) in the sediments also raises the possibility of horizontal gene transfer of antibiotic resistance between different pathogenic bacteria (Husevag 1991). With fish pathogenic bacteria remaining viable in fish farm sediments for long periods of time (Vibrio salmonicida causing vibriosis, Enger et al. 1989) and developing resistance to antibiotics (Husevag et al. 1991), there is a possibility for the reinfection of fish through degassing of the sediments returning the bacterium into the water column (Levings 1992). There is also  45  a potential for the development of cross resistance between oxytetracycline, oxalinic acid and furazolidone in bacteria, creating a greater risk for passing this resistance on between different bacterium (Trevors et al. 1986, Lunestad & Goksoyr 1990, Inglis & Richards 1991, Nygaard et al. 1992). Coliform bacteria such as Escherichia coli and Klebsiella spp. are present at fish culture sites and sewage contaminated waters (McPhearson et al. 1991, Qureshi & Qureshi 1992), with 10% of the coliform strains resistant to 8 types of antibiotics. Increased cross resistance in a greater portion of human pathogenic bacteria could result with increased usage of antibiotics, increasing a risk for human infections that cannot be treated by the usual antibiotics. The inability to treat bacterial diseases would also greatly curb the potential for cross culturing species (e.g. salmon and shellfish culture), particularly in the presence of human pathogenic bacteria. Preliminary results indicate that there is an overall increase in antibiotic resistance of bacterial isolates (Family Vibrionacea that cause gastroenteritis, septicemia, meningitis and skin/eye/ear infections in humans), from cultured fish and surrounding waters of B.C. coastal fish farms (Dr. M.T. Kelly, unpublished results). These bacterial isolates exhibited cross-resistance to the drugs ampicillin, tetracycline and cotrimoxazole. There was also increased resistance by V.  anguillarum to ampicillin, the causative agent of vibriosis in farmed salmon. Harvesting of wild fauna (fish and shellfish) around fish farms is another possible route for human infections. The consumption of raw cultured and wild shellfish which concentrate V.  vulnificus by filtration (e.g. Kelly & Dinuzzo 1985, Kelly & Stroh 1988), is a common cause of gastroenteritis (DuPont 1986). A large outbreak of gastroenteritis in Singapore in 1979 was traced to shellfish imported from the Phillipines, where Manila Bay is one of the major aquaculture sites for shellfish but is also one of the most polluted water bodies in Southeast Asia (Eng et al. 1989). The presence of antibiotics and resistant bacteria have been isolated from wild fish and shellfish which were feeding near the culture sites during medicated feed treatments (Bjorklund et al. 1990, Samuelsen et al. 1992).  46  The experimental evidence of Toranzo et al. (1984), suggests that plasmid gene transfer between fish and the aquatic environment is possible. They determined antibiotic resistance patterns and resistant plasmids of bacteria isolated from the water in hatchery tanks and the skin of rainbow trout. Of 170 bacterial isolates (belonging to 8 bacterial groups) from the cultured fish, 87.6% of the isolates were resistant to at least 1 drug, and the majority of these strains (85.2%) were multiresistant. The plasmid coding for resistance to chloramphenicol was found in Vibrio spp., Citrobacter spp., and Enterobacter spp., and resistance to sulfadiazine, tetracycline, nitrofurantoin and ampicillin were also cotransferred some of the time. If the contaminated hatcheries were not adequately cleaned prior to transferring in uninfected rainbow trout, there would be a great potential for water-borne contamination of the new fish. While there is concern about the potential inability to treat human pathogenic diseases contracted from cultured and wild fish and shellfish due to antibiotic resistance through the increased use of drugs in aquaculture, this has never been substantiated in the natural environment.  Mineralization of Detrital Carbon The addition of antibiotics to the microbenthos could affect the microbial and bactivorous protozoan populations, which in turn would affect the mineralization of detrital carbon. Providing that the microbial population displays resistance to the antibiotics used, elimination of the protozoan population could allow the bacteria to rapidly multiply and increase the degradation of organic compounds (Wiggins et al. 1987). However, the results from this study appeared to indicate that the presence of protozoans may have actually stimulated bacterial growth and activity, while bacteria increased in the absence of protozoans but probably not in activity (Section 1.5.7). From feeding experiments with isopods, Smock & Harlowe (1983) showed that the nutritional quality of detritus was decreased when microbial growth was inhibited by antibiotics. Inhibition of bacterivorous protozoans may be undesirable because of their role in nutrient cycling (Lighthart 1969,  47  Barsdate et al. 1974, Stout 1981, Kemp 1990). Benthic protozoan grazing can contribute indirectly to the mineralization of detritus, possibly by facilitating nutrient availability and the rate of nutrient turnover by lowering bacterial biomass (Pratt & Cairns 1985, Wiggins & Alexander 1988). This could in turn stimulate the growth rate of bacteria and increase the uptake of nonlimiting nutrients (Stout 1980). The half-life of E. coli in sewage waste treatment systems have been shown to decrease from 16 hrs to 1.8 hrs, while the uptake of dissolved phosphorous was also found to be 4 times greater by bacteria being grazed on by protozoans (Stout 1980). This was attributed to a higher proportion of younger, more rapidly dividing bacterial cells, compared to a more static population. However, protozoans can serve directly as a source of carbon in the flow of energy through the trophic web, by concentrating nutrients within themselves and making themselves available to other heterotrophs (e.g. Fenchel & Jorgensen 1977, Tenore & Coull 1980, Meyer-Reil & Faubel 1980, Stout 1980, Lopez & Levinton 1987, Sanders 1987, Stoecker & Capuzzo 1990). The importance of the microfauna (including the meio- and macrofauna) in the degradation of organic matter has been demonstrated (Small 1973, Curds 1973, Tietjen 1980, Mann 1988, Plante et al. 1989, Toerien et al. 1990). The input of the microfauna to the detrital food web may depend on the types of sediment, organic matter and trophic webs present (e.g. Tenore et al. 1982, Rieper 1985, Hansen et al. 1987, Alongi 1988, Kemp 1987 & 1988, Jumars et al. 1989, Kuuppo-Leinikki 1990, Riddle et al. 1990, Nilsson et al. 1991). While the extent of the contribution of the microfauna in the degradation of organic carbon and nutrient cycling in this study cannot be determined, the importance of the fish waste in enhancing growth of the microbial community, and of the bactivorous protozoans as a potential grazing control is apparent.  The addition of fish waste to the sediments created a highly anoxic and reducing environment within a day (this study). This was an indication of the consumption of free oxygen from the sediments in the degradation of the fish waste, and a subsequent shift in  48  anaerobic microbial metabolism from manganese, nitrate and iron reduction to sulphate reduction. The OTC was rapidly washed out of the fish waste when added to the sediments, with small quantities of the antibiotic persisting until the end of the study. Although the dissolved oxygen content of seawater decreased due to BOD from the fish waste, the fluctuations in the dissolved oxygen levels between treatments were not significantly different. Carbon:nitrogen ratios decreased with the addition of fish waste, reflecting the greater nitrogen to carbon content of the fish waste. Bacterial abundance peaked at 26x109 •g-1 sediment by day 4 in Treatment B, and was significantly different from the Control and Treatment A. The rapid rise in cell counts during the initial few days may have been due to the potential inhibition of the bactivorous protozoan population and development of OTC resistance in the anaerobic bacterial community. The amount of OTC in Treatment A was significantly less than in Treatment B, thus affecting any grazing on the microbial community by ciliates to a comparatively lesser degree. While bacteria increased in the absence of protozoans due to the OTC, metabolic rates may not have been affected, potentially slowing mineralization of the fish waste. While the presence of protozoans may have stimulated growth and activity in the sulphate reducing bacterial community, only the decrease of OTC and redox state of the sediments, and changes in bacterial abundance can be conclusively demonstrated. The possible effects of OTC on the microbial and protozoan communities must remain speculative until further research has been completed.  To this point, this thesis has dealt with only a portion of the microfaunal community. The next chapter will present the results of a pollution experiment in which an attempt was made to simulate the sedimentation rates of fish waste and OTC beneath a fish farm. Oscillations in the abundance of microfauna (ciliates, bacteria), diatoms, and meiofauna (nematodes, copepods) are presented.  49  2. THE Eli ECTS OF FISH WASTE AND OXYTETRACYCLINE ON THE MICRO- AND MEIOFAUNA.  2.1 INTRODUCTION  Intensive cultivation of fish, crustaceans and molluscs generates large amounts of solid wastes (uneaten feed and feces) that can accumulate on the sediment bottom (e.g. Weston 1986, Gowen & Bradbury 1987, Iwama 1991). Ecological assessment of the impact of aquacultural activities on the benthos has been restricted in the past to the macrofauna or macrobenthic invertebrates (Weston 1986). (Macro referring to those organisms which are retained on a 0.5 mm sieve mesh; Eleftheriou & Holme 1984). The larger size of the invertebrates allows them to be more easily observed and sampled. In zones of heavy organic enrichment beneath and immediately surrounding fish farms, species diversity generally decreases, with an increase in biomass and abundance from opportunistic species (Weston 1986). The dominant macrofauna (or infauna, as they burrow within the sediments) is commonly the small polychaete, Capitella capitata, from 8x10 3 to over 25x10 3 individuals•m -2 (Mattson & Linden 1983, Tsutsumi 1987, Brown et al. 1987, Weston 1990, Aquametrix 1990 & 1992). However, one of the disadvantages to the use of macrofauna (especially polychaetes) as a biological indicator of organic pollution, is the relatively long interval required to observe large changes in abundance. This is due to the longer reproductive cycles of the polychaetes (from 4 to 6 weeks) and environmental monitoring can take several months to a year (Pearson et al. 1982, Aquametrix 1990 & 1992, Tsutsumi 1990a & 1990b). Another drawback is that the sieve mesh sizes used in isolating macrofauna from sediments are often too large, and do not sample the smaller fraction of the macrofaunal community, missing the micro- and meiofauna (Bachelete 1990). This leads to potentially inaccurate estimates of abundance, and erroneous reports that polychaetes are the only infauna present (Kaspar et al. 1985; isolating macrofauna from  50  sediments beneath a mussel farm using a 1.68 mm mesh sieve). Using a 0.5 mm mesh sieve, Weston (1990) found that nematodes, and C. capitata comprised over 99% of the total number of individuals beneath a salmon farm in Puget Sound (Washington, U.S.A.). Biological monitoring with macrofauna thus requires long time intervals, which can hamper the identification of polluted sediments, and the rapidity of preventative or restorative management measures (Weston 1986). These actions may involve periodically rotating culture sites, submersible mixers to disperse the fish waste, and the collection of dead fish and fish waste by the use of funnels beneath net pens (Gowen & Bradbury 1987). Other methods include the use of extruded rather than pressed feeds, which remain more stable in water and sink slower allowing the fish a greater opportunity to ingest the feed, and more efficient feeding techniques such as hand feeding to minimize wastage (Seymour & Bergheim 1991). In addition, the inherent weakness in sampling the macrobenthos (i.e. sieving techniques that miss smaller organisms) assures that comprehensive and accurate studies of the benthic community cannot be obtained. Sampling of the microfaunal community (and perhaps meiofauna as well) can provide a solution to these limitations. While identifying the microfauna requires a greater amount of time and expertise (as discussed in Chapter 1), the rapid response of the microfauna to a community stress (within a few days) makes them a potential biological monitoring tool in outlining management objectives in minimizing organic enrichment of sediments at fish farms. The microfauna will include the microbial and protozoan communities, while the meiofauna community includes nematodes and harpacticoid copepods. In addition benthic diatoms will also be enumerated. While some of the meiofauna surrounding aquaculture sites have been sampled briefly before (Weston 1990), a comprehensive microfaunal survey should include the meiofauna due to possible micro- or meiofaunal interactions (Fenchel 1978, Tietjen 1980, Alongi 1988). In this experiment, fish waste and oxytetracycline was applied over 20 days in 200 L microcosm tanks, to simulate organic enrichment of the benthos at fish  51  farms. The rapid changes in the microfaunal community (and the meio- and macrofauna to a limited extent) were examined.  2.2 OBJECTIVES  The objectives of this study were:  (1)  To establish the responses of the microfauna (bacteria and protozoan communities) to fish waste and oxytetracycline.  (2)  To evaluate the physical-chemical changes to the benthos, including (i) oxygen levels in seawater due to the biological oxygen demand of the sediments, and (ii) sediment depth profiles of the reduction-oxidation potentials.  (3)^To follow the decrease of oxytetracycline over time.  2.3 MATERIALS AND METHODS  2.3.1 EXPERIMENTAL DESIGN  The experimental period was from July to August in 1991. There were three treatments, a Blank treatment without any fish waste or OTC added, a Control treatment with fish waste but no OTC added, and an OTC treatment with fish waste and OTC added. The Blank treatment consisted of three, 20 L sediment tanks, while three, 200 L sediment tanks were used for each of the Control and Oxytetracycline (OTC) treatments (for a total of six, 200 L tanks) (Fig. 2.1). The sediment was collected from Spanish Banks at low tide (Vancouver Harbour) and added to the tanks to an average depth of 5 cm. Sediment composition ranged from fine to coarse sand (Fig. 2.2). The sediment was acclimatized in the tanks for one month, to allow the biological, physical and chemical factors to reestablish. Water temperature and salinity during the study period varied from 11 to  52  Figure 2.1. The arrangement of the 200 L tanks in which fish waste and oxytetracycline were applied. The order of the tanks for the treatments was randomly chosen.  53  70 — 65 = 60 = 55 50 0  0 0  45= 40 =  0  35 30  U  25 = 20  (11  7  15 = 10  -  5 -0= <53^53-75^75-180 180-355 355-595 ^>595  Sediment Size Range (urn)  Figure 2.2. Grain size composition by relative percentages, of the sediment obtained from the intertidal zone during low tide at Spanish Banks/U.B.C. The majority of the sediment is comprised of fine to medium sand (65%), followed by a medium to coarse sand component (26%).  54  13.5°C and 26.5 to 29.5 04, , respectively. As the tanks were indoors with poor and 0  variable fluorescent lighting, covers were provided for all tanks to eliminate any differences due to autotrophic activity and epiphytic growth on the walls of the tanks. Due to the rapid increases in the bacterial population in the first phase of this study in July and August of 1990 (Chapter I) and of smaller ciliates (Fenchel 1982), the tanks were sampled every second day to ensure that a peak in microfaunal abundance would not be missed. The duration of the experiment was thus limited to 20 days (July 13 - August 2) due to the practical restraints in time required for sampling and analysis. Fish waste was applied to the Control and OTC treatment tanks every day, while only oxytetracycline was added to the OTC treatment tanks for 4 days (from Day 9 to Day 12 of the experiment).  2.3.2 THEORITICAL CALCULATION OF ORGANIC LOADING AND OTC  The discharge of excess fish feed and feces from an aquaculture facility can be calculated, based on the size of the cage or net pen and stocking densities, husbandry practices and the feed ingested and excreted (Iwama 1991, Hajen 1990, Henderson and Bromage 1988, Weston 1986). However, the accuracy of these estimates will depend on water quality conditions around the fish farm. A mean sedimentation rate was estimated to be 62 g dry wt•m -2 •d -1 , by suspending sediment traps directly beneath 7 salmon farms and 1 smolt farm in B.C. (Aquametrix 1988). Based on this figure and the sediment surface area of 0.385 m 2 for the 200 L tanks (Fig. 2.3), 23.9 g dry wt of fish waste was required to be applied to each 200 L tank per day. Oxytetracycline was contained in a coarse mixture by the trade name of Oxysol 440 (supplied by Argent Laboratories, Richmond, B.C.). Treatment levels of fish with Oxysol  440 is prescribed to be at 100 mg•kg -1 of biomass, or 11.5 g Oxysol 440-kg -1 fish feed (Dr. Dahl, D.V.M., pers. comm.). One gram of Oxysol 440 contains 0.44 g OTC. Although fish are usually treated for a period of 10 days with OTC (e.g. Bjorklund et al. 1990), the logistics of this study limited the application of OTC to only 4 days. From 11.5  Figure 2.3. A schematic of the division of grids used to randomly sample the sediments of the 200 L tanks. The sediment is divided into a system of 3 columns (A-C) and 5 rows (1-5), with a central drainage pipe and a 1.5" border which was not sampled to avoid edge effects.  56  g Oxysol 440•kg-1 fish feed and 0.0239 kg dry wt of fish waste•m -2 •d-1 , approximately 0.275 g Oxysol 440.m -2 .d -1 was required to be applied to the 200 L OTC treatment tanks (or 0.005 mg OTC•mg-1 dry wt of fish waste).  2.3.3 APPLICATION OF FISH WASTE AND OTC  Fish waste was collected from 10,000 L outdoor circular tanks, concentrated in 500 ml plastic bottles and frozen at -12°C until just before use. The dryweight to volume ratio of the fish waste was predetermined. After the fish waste was thawed, the appropriate amount was measured out into smaller containers, the OTC added if necessary, and the containers capped. The fish waste was only added to the 200 L tanks after sampling was completed each day. The containers of fish waste were lowered into the water above the sediment, and the cap opened. The fish waste was slowly and evenly distributed over the sediment surface, and care was taken to avoid disturbing the sediment. The flow of water to each tank was turned off to allow the fish waste time to settle to the sediment, and was turned on again after half an hour.  2.3.4 SAMPLING DESIGN  The 20 L sediment tanks were sampled following the grid system previously established (Chapter 1). The 200 L tanks were also randomly sampled based on a system of squares (Fig. 2.3) with a 1.5" border to avoid edge effects. The sides of the tanks were not wiped down to avoid disruption of the benthos. A single large sediment core was taken (i.d. 5 cm) from which smaller subcores could be drawn with the 1 ml and 3 ml syringe corers. (The cores for Eh measurements were taken directly from the 200 L tanks to ensure that the depth profile of the sediments remained undisturbed). Bacterial, dissolved oxygen, OTC and Eh samples were all collected and measured as previously discussed (Chapter 1, Appendix 1 and 2) with bacterial samples enumerated the same day. Dissolved oxygen and Eh were measured every fourth day.  57  Ciliate, nematode and copepod samples were taken from the same 3 ml core, and the total abundance recorded. Ciliates were extracted by the seawater-ice technique (Uhlig et al. 1973) using a 180gm sized mesh and collecting the ciliates in a petri dish. This technique was modified by adding 8% MgC12 to the crushed ice to enhance extraction by relaxing the ciliates (Dr. J. Berger and M. Adl, pers. comm., Kirby 1950). The ciliates were immediately preserved in Bouin's solution (Lee et al. 1985, Montagnes & Lynn 1987). The ciliate samples were later placed in 25 ml settling chambers, and enumerated after 24 hrs using an inverted microscope. The entire microscope field was scanned and counted, and transects were only necessary when the smaller ciliates (<50 Am) were very abundant. The sediment was then washed into a petri dish and 10% ethanol was added to anesthetize the nematodes and copepods. After 10 minutes, nematodes and copepods were collected by the washing and decantation technique into a 250 m mesh sieve (McIntyre & Warwick 1984). Three washes ensured that all organisms had been extracted from the sand (personal observations). The samples were preserved in 5% formalin buffered with borax and stained with Rose Bengal. Each sample was enumerated in a Bogorov tray, and the entire sample counted under a dissecting microscope. Diatoms were collected with a separate 3 ml corer and the sediment was immediately placed in a 20 ml scintillation vial, filled with filtered artificial seawater and Lugol's solution was added. Each sediment sample was diluted 10 times, and three, 3 ml subsamples were taken using a 5 ml automatic pipette as the sediment was mixed with a stirring bar. Each subsample was further diluted 15 times, and a 3 ml volume from each subsample was placed in a 5 ml settling chamber for enumeration beneath an inverted microscope. Data analysis was as previously described (Chapter 1).  58  2.4 RESULTS  2.4.1 VISUAL OBSERVATIONS  The surface sediment was light brown in all tanks, and sufficient fish waste had accumulated by day 4 in the 200 L tanks to become visible as a fine, dark layer of organic matter. Nudibranchs were present on the sediment surfaces of all tanks, a common species found along the coast of the Pacific Northwest (Thick Horned Aeolid, Hermissenda crassicornis) that feed on coelenterates (Kozloff 1987). The nudibranchs were probably  introduced from the seawater pipe intake system. A burrow from a marine invertebrate worm (unidentified) was present in tank 1 of the Control treatment which was avoided during sampling. The physical-chemical and biological characteristics of the sediment around the burrow will be different due to the introduction of oxygen into the sediment and feeding activities (e.g. Findlay et al. 1990b). A small fish, the Pacific sandlance, Ammodytes hexapterus Pallas (approximately 10 cm long) was in tank 3 of the Control  treatment. The sandlance was observed to spend the majority of its time in the sediment. A. hexapterus has been documented to feed on plankton, especially small crustaceans  (Clemens & Wilby 1961). It was not until day 8 (beginning of OTC addition) and day 9 that sporadic patches of the white microbial layer began to develop on the sediment surfaces. This white layer was smooth in appearance, while the accumulated fish waste was flocculent in nature (dark green to black in colour). By day 12 at the end of the oxytetracycline treatment, all of the tanks in the OTC treatment had a well developed white microbial layer on the sediment surface, underlain with dark brown sediment. In comparison, the tanks for the Control treatment had a less well developed, patchier distribution of the white microbial layer. At this time small benthic polychaetes (Phylum Annelida) had appeared on the sediment surfaces of the Control and OTC treatment tanks. The polychaetes ranged in size from 0.4 to 1.5 cm in length, with an average density of 1,079 individuals•m 2 (standard deviation of  59  377 individuals). A benthic worm (0.6 cm long) was isolated and tentatively identified as a sipunculid (Dr. C. Lalli) a detritus feeder. The nudibranchs had begun to avoid the sediment surfaces and had moved to the walls of the tanks. On day 16 the Control and OTC treatment tanks were characterized by the absence of all nudibranchs, and the white microbial layer in the OTC treatment tanks had begun to change from being smooth to spotted in areas (resembling small tufts). The smooth white microbial layer in the Control tanks was now well developed. The sediment surface of the Blank treatment tanks remained brown with nudibranchs present throughout the study period. By day 18, the white microbial layer of the Control tanks had also begun to appear spotted, and the nudibranchs had reappeared on the walls of the Control and OTC tanks. A nudibranch in tank 3 of the OTC treatment was observed to be laying eggs, and egg patches were present in all of the 200 L tanks. A small, younger nudibranch was observed in tank 3 of the OTC treatment. Sedimentary conditions had not changed for any of the tanks by the last day of the experiment, day 20. The burrow in tank 1 and the sandlance in tank 3 of the Control treatment remained active throughout the study. When the water supply was shut off for more than half an hour, the sandlance would emerge from the sediment until the water was turned on again.  2.4.2 REDOX PROFILES  (Eh)  As previously discussed (Chapter 1) the Eh measurements are related to the microbially mediated reducing or oxidizing conditions occurring in the sediments. The redox profiles for all of the tanks in the Blank treatment indicated that the sediment surfaces (and 0.5 cm above the surface) remained positive and highly oxidizing throughout the entire study period at approximately +250 mV (Fig. 2.4). The redox potential discontinuity (RPD) layer remained broad and deep, down to 2.5 cm depth throughout the study. The redox potentials were positive at most depths in all tanks, and was oxidizing ( > +100 my)  60  O  N  0^ N_^8^0 7^7 I^I  1 .0  Eh (mV) 0 0^:11  8N  0  N  I^I^I  O  0^0 N^c0  aa  ^  O  r")^,0")  N  I^.^I^■  I  0.5  E  0.0 0.5  Replicate 1  1 .0  a)  0 • ^ ^ ^ ■  1.5 2.0 2.5  a) 0)  3.0 3.5  Day 0 Day 4 Day 8 Day 12 Day 16 Day 20  4.0  8  0  7  0^0  0  1.0  0 ♦  co  0  8  X  0^0 8 r.., ,,,,  N  ,,-, Pn  I  1  ■  I  8♦  I  0.5 0.0 0.5 1.0  a)  Replicate 2  -  Day 0 not available • Day 4 7 Day 8 ^ Day 12 ^ Day 16 • Day 20  1.5 2.0 2.5 3.0 -  a)  0)  3.5 4.0 0^0^0 0^te,^ IC^0 ^0  CO  at  0  I^I  [  1.0 0.5 0.0 -  E  0.5 1.0  "at CZI  a)  E a)  0)  -  1.5 2.0  -  2.5 3.0  -  3.5 -  Replicate 3 Day 0 not available • Day 4 77 Day 8 ^ Day 12 ^ Day 16 • Day 20  4.0  Figure 2.4. Reduction-oxidation time-series depth profiles of the replicate tanks for the Blank Treatment, followed over 20 days in which no fish waste or oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began 0.5 cm above the sediment surface.  61  in tank 2 at all depths for the duration of the study. Although the Eh below 2 cm depth became reducing in tanks 1 and 3, a pattern to the changes in the redox profiles was not apparent. The sediments in the replicate tanks of the Control treatment were initially similar to those of the Blank treatment, with oxidizing conditions on day 4 and 8 (Fig. 2.5). But the sediments gradually became more reducing as fish waste was continually added to the sediments. This was strongly apparent by day 12 and 16, where the Eh readings at the surface and 0.5 cm above the surface were strongly reducing and negative. Reducing conditions of -200 mV in tanks 1 and 3, and -40 mV in tank 2 were reached at the end of the study by day 20. The sediments in all of the tanks were more reducing near the surface than at depth at this time (as in the prior study), although this reverse RPD layer remained deep (down to 3 cm). The changes in the Eh of the sediments for the OTC treatment (Fig. 2.6) followed the same pattern in the Control treatment. The surface sediments were initially well oxygenated, from the highly positive Eh values ( > +300 mV, Fig. ) and were -160 mV in tanks 1 and 3, and -120 mV at the sediment surface of tank 2 by day 20. The reverse RPD layer was also present, as the Eh near the sediment surface was more reducing than at depth by day 16 for all tanks. The addition of OTC did not make any difference in the between this treatment and the Control.  2.4.3 DISSOLVED OXYGEN  The oscillation in dissolved oxygen levels over the experimental period followed a similar pattern for all three treatments (Fig. 2.7), decreasing immediately and peaking again on day 12 before dropping again. The percent saturation of oxygen in seawater varied between 78% and 92% for all 3 treatments. Dissolved oxygen levels in the Blank treatment were significantly different from the Control treatment (P=0.009, Tukey Test) but not with  62 Eh (mV) C  o4^  7^7  2  ▪  I^I^I  1.0  0  0  8  0  C ro  0  8  C  0  o  0  C  esc  0  0  I^I  0.5 0.0 •  0.5 -  .4 1.0  a  1.5  -  2.02.5-  :45 •  3.0 -  •  3.5 4.0 ^ 0^0  0  so.  0^0 CO  "t  0^0  •  Fe,  to  I,^I  1.0 0.5  E  0.0  n.  1.0  Replicate 2  1.5 -  Day 0 not available • Day 4 Day 8 Day 12 ^ Day 16 ■ Day 20  C  E V)  0.5  2.0 2.5 3.0 3.5 4.0  0^ 0^0^0^0^0 0^0^eV^00^0^1.^0^IN^tO 0^47^CV^0^0^ CV^CC^N^h^•") •-•^CO^1.^0^V'^CO^  0^0^0^  t^I,^I^ItI^I^II^I^I^I^I,^I^I  1 .0  I  0.5 0.0 0.5 1.0  Replicate 3  -  1.5 2.0 -  E  2.5 -  a.> 3.0 3.5 -  -  a  Day 0 not available • Day 4 7 Day 8 ^ Day 12 ^ Day 16 ■ Day 20  4.0 ^  Figure 2.5. Reduction-oxidation time-series depth profiles of the replicate tanks for the Control Treatment, followed over 20 days in which fish waste but no oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began 0.5 cm above the sediment surface.  ^  ▪ 63 Eh (my) 0 0  0^0  0^0 O  7  ,^I  1.0  Fr^  2  !=  0  0  0  0 CNI  ro  0  0  I^I^I^I^I^I^I^I^I  0.5 0.0 c.)^0.5 1.0 •  1.5 2.0  not • ^ ^ ^ ■  2.5 •  3.0 3.5 4.0 ^  8  ^03  I  1.0  0  t0^N  ^1.  I^I^I  -  c.) 0.5  -  E  c  1.0  0^..... •  ---  - ---  ---  -  --------  -^  _  c2  -  10'  -  ■^.^Day 0 ^not available / 7 0 • Day 4 /^ I •v Day 8 \,v^ • Day 12 0 Day 16 Replicate 2 ■ Day 20  2.5 -  up  N  N^  1I^I^I^1^I^I^I^It^l^l^l^I^I^I  al 1.5 • 2.0 v  •  F,  o^CC  0.5 0.0  Day 0 available Day 4^ Day 8 Day 12 Day 16 Day 20 Replicate 1  3.0 3.5 4.0 ^ 0^0 o t^c,^0 o cv^-^ a^o I^1^ ;  N  I,^I^I^I^I  1.0 0.5 0.0  c2)^  C  ^  0^0^0^0 0^mr;^cci'D pr;  I  ,^1^I^I  1  0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0  Day 0 not available -• • Day 4 ^ Day 8 ♦ Day 12 Da 16, Replicate 3 0 Day 20  •  Figure 2.6. Reduction-oxidation time-series depth profiles of the replicate tanks for the OTC Treatment, followed over 20 days in which fish waste and oxytetracycline was added. Values < +100 mV (dashed line) indicates reducing conditions. Measurements began 0.5 cm above the sediment surface.  64  100 -  95 V 90 cu tx ›., k 0 85 -7 -ci .  a)  0 0  w ..  80  -  75 -:  70  o Blank • Control v OTC Treated 1^1^1^i^I^[^I^I^1^I^1 0^2^4^6^8^10^12^14^16^18^20 Time (Days)  Figure 2.7. Comparison of the dissolved oxygen levels in seawater between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation, n=3.  65  the OTC treatment. Dissolved oxygen levels did not differ between the Control and OTC treatments (P > 0.05).  2.4.4 OXYTETRACYCLINE  OTC was only detectable on day 10, at 13.82 ± 22.87 ppm. The initial concentration of OTC found in fish waste was 939 ± 68.7 ppm for a total of 0.12 g OTC (mean and standard deviation, n=3), but was 1100 ppm for 10.13 g OTC in Treatment B in the previous study (Chapter 1). The initial amount of OTC should have been much lower in this study or much higher in the previous study. This difference is due to the use of an external standard (OTC was injected directly on the HPLC column) in the previous study, compared to an internal standard (TC extracted from the sediments) used in this study for the calculation of the concentration of OTC. Consequently, the concentration of OTC in the previous study would be underestimated due to the use of the external standard. Residual amounts of OTC were detected in the sediments of this study for the remainder of the experimental period.  2.4.5 BACTERIAL ABUNDANCE  Total bacterial abundance of the Blank, Control and OTC treatments on day 0 was very similar (Fig. 2.8), between 6 - 7x10 9 • -1 sediment (corrected cell counts, Appendix A). Bacterial abundance for the Blank treatment remained approximately constant throughout the treatment period, while both the Control and OTC treatments exhibited a peak in abundance by day 8, 1.44x10 10 bacteria•g -1 and 1.42x10 10 bacteria•g -1 sediment respectively. Microbial abundance in the Control treatment appeared to peak 4 days earlier than in the OTC treatment, day 8 and day 12 respectively, but this trend was not readily apparent due to the large variances present. After peaking, cell counts decreased in the Control and OTC treatments, remaining constant in the Control treatment but continued to drop in the OTC treatment. Overall, bacterial abundance was significantly enhanced by the  66  20 19 18 17 16 15 14 • 13 12 ^ 7 11 rn  0  o  9 8 7 6 6 5 4 3 2 1 0 • •  -  o Blank • Control ^ OTC ^  0  2^4^6^8^10^12^14^16 Time (Days)  ^  18^20  Figure 2.8. Comparison of bacterial abundance (cell x 109 • -1 sediment dry weight) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation.  67  addition of fish waste in both the Control and OTC treatments when compared to the Blank treatment (log transformed data, Tukey Test, p< 0.05). Addition of oxytetracycline to the fish waste did not significantly affect microbial abundance between the Control and OTC treatments (P> 0.05).  2.4.6 MICROBIAL MAT  The development of the microbial mat occurred simultaneously in both the Control and OTC treatments. The bacterial community were initially small, individual rod and cocci shaped bacteria 1.5 Am in length. On day 6 in both Control and OTC tanks, larger rods began to appear in the sediment and were approximately 2.5Am in length. By day 8 when the smooth, white microbial layer began to be established, the microbial community was composed of both small cocci and large rods, with the rods forming many dense, clustered chains and colonies. After oxytetracycline was added to the OTC tanks for 4 days, there were less bacteria present overall after day 12, and fewer large rods and chains. The addition of oxytetracycline appeared to be correlated to the slower development of the large rod shaped bacterial chains in the microbial community. Visual observations of bacterial cells also seemed to indicate that the larger rods were associated with the white microbial film, as the smaller rods and cocci attached directly to sand grains while the larger chain forming rods were not associated directly with the sediment. It was also observed that fine layers of the bacterial film formed in the stagnant water column above the sediment surfaces of the cores after several hours. Examination of this microbial film indicated that it was indeed composed of only the large, chain forming rods. By day 12, the large rods had begun to decline, and the smaller cocci and rods began to predominate again. The morphological characteristics of the microbial community again changed on day 16 in the OTC treatment, when the smooth microbial film began to appear spotted (small tufts of growth). This spotted microbial film was not as well established in the Control treatment. Under the epifluorescent microscope, these tufts of  68  bacterial growth consisted of large, thick walled units that formed strands (approximately 10 gm wide and up to 2 mm long), which were filled with small rod shaped granules (3 to 8µm in length) that stained orange. Examination of the bacterial tufts with an inverted microscope showed that the tufts were capable of limited motion. By day 20 the white bacterial tufts were well established in both the Control and OTC tanks, but with an uneven distribution on the sediment surface.  2.4.7 PROTOZOAN ABUNDANCE The use of Bouin's Solution as a fixative caused extreme contraction in the benthic ciliates, which made accurate identification impossible. Only total ciliate abundance was estimated (Fig. 2.9a) and identification was based on the limited morphological characteristics visible through the inverted microscope. Overall, ciliate abundance in the Control and OTC treatments was significantly different from the Blank treatment (log transformed data, Tukey Test, P <0.05). While the addition of fish waste greatly increased ciliate abundance after 4 days in both the Control and OTC treatments, oxytetracycline did not cause any significant changes in ciliate abundance between these two treatments (P > 0.05), although the initial peak in ciliate abundance in the OTC treatment lagged behind the Control treatment by 2 days. A bimodal peak in abundance occurred during both of these treatments, with a greater initial peak in abundance of 35x10 3 individua1s•cm on day 6 and 26x10 3 individuals•cm -2 day 8 for the Control and OTC tanks, respectively.-2 The second peak in abundance was lower, with 14x10 3 and 22x10 3 individuals•cm -2 on day 12 in the Control and OTC treatments, respectively.  A bloom of heterotrophic microflagellates appeared on day 16, after the ciliate population declined on day 14 (Fig. 2.9b). The microflagellates consisted of small, circular cells (3-6 gm in diameter) that were observed to be feeding on the bacterial community within the fish waste. The microflagellates reached densities of 13x10 6 individuals•cm -2 in  69  60-  A o Blank • Control v OTC  40o^  -  30-  a) .0  z  20-  10  -  •---•---10---0 ♦^0 • • • • I^.^I^,^I^,^1^I^,^I^i^,^I^,^1^,^I^1 ,..  ^0  0^2^4^6^8^10^12^14^16^18^20  B • Control v OTC  4 2 0 0^2^4^6^8^10^12^14^16^18^20  Time (Days)  Figure 2.9. Comparison of protozoan abundance over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added); where A =ciliate abundance (numbers x 10 .cm-2 ); B=heterotrophic microflagellate abundance (numbers x 106•cm-2 ). Plotted values are the Mean ± 1 Standard Deviation.  70  the Control treatment, and 7.3x10 6 individuals•cm -2 in the OTC treatment, but the 2 treatments were not significantly different (T-Test, P >0.05). The microflagellates did not become noticeable until abundance increased significantly, and were thus not enumerated prior to day 14. In this organically poor sediment, there were initially few ciliates in the sediment tanks from day 0 to 4. The ciliates present ranged in length from 50 to > 300 Am, and increased numbers in the 50-150 Am size range appeared with the continued addition of fish waste. Some of the larger ciliates that appeared on day 6 belonged to class Karyorelictea (Order Protostomatida N. Ord. and Order Loxodida Jankowski) and subclass Haptoria (Lee  et al. 1985, Corliss 1979). Small ciliates (approximately 20-40 Am in length) appeared on day 4, and were the major contributing factor for the large increases in ciliate abundance. Rotifers appeared by day 10, but were not enumerated. Examination of live samples was accomplished by placing a drop of sand (obtained with an eyedropper) in 5 ml settling chambers and observed through the inverted microscope. Seven morphologically dissimilar benthic ciliates were readily distinguishable in the Control and OTC treatments up to day 14 and 16, but had declined to only 3 or 4 morphologically different ciliates at the end of the study period. There did not appear to be any differences in ciliate species diversity between the Control and OTC treatments. Benthic ciliate diversity in the Blank treatment always remained low, with never more than 3 or 4 distinct ciliates present.  2.4.8 DIATOM, NEMATODE AND COPEPOD ABUNDANCE While diatom abundance increased slightly at the end of the study period from 8x102 to 26x102 cells•cm -2 (Fig. 2.10), no significant difference was found between the 3 treatments (P >0.05, Tukey Test). All of the diatoms were of the pennate variety, and appeared colourless or with the cytoplasm greatly reduced. Only those cells with a majority of cytoplasm were counted.  71  36 — 32 — 28 —  c,2 1  co  0 Blank • Control V OTC  E 0 24  o --,  20  v) s..^16 a>  ,..0  Z  12 8  0^r^  1  1  0^2^4^6^8^10^12^14^16^18^20 Time (Days)  Figure 2.10. Comparison of diatom abundance (numbers x 10 2 •cm -2 ) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation.  72 Nematode densities were highly variable for all 3 treatments, with the highest abundance in the Blank treatment at 53 individuals•cm -2 on day 16 (Fig. 2.11). By the end of the study, nematode abundance declined in the Blank treatment and rose in the Control and OTC treatments, but was still not significantly different between treatments (P >0.05, Kolmogorov-Smirnov Test). Nematode sizes varied greatly from several hundred m to 1 cm in length. Values for harpacticoid copepod abundance were significantly different between all 3 treatments (P <0.05, Kolmogorov-Smirnov Test, Fig. 2.12). Densities initially decreased for all treatments, but remained highest in the Blank treatment at 23 individuals•cm -2 on day 16, while copepod abundance in the Control and OTC treatments remained at depressed levels until near the end of the treatment period, when abundance in all 3 treatments converged. Copepods with egg cases and nauplii were present throughout the study period.  2.5 DISCUSSION  2.5.1 REDOX MEASUREMENTS The sediment in the Blank treatment remained oxidized with very little change in redox conditions during the experimental period (Fig. 2.4), as expected when the fish waste was not added. In comparison, the sediments of the Control and OTC treatments steadily became more reducing with time at the surface and at depth by day 12 and 16 (Fig. 2.5 and 2.6). The highly reducing conditions, visual observations of the changes in colour of the sediment and the development of the white microbial mat confirmed that microbially mediated anaerobic redox reactions (i.e. sulphate reduction) were occurring (as in Chapter 1; e.g. Jones 1980, Nedwell 1984, Revsbech & Jorgensen 1986, Bak & Pfennig 1991). Up to 86% of the sulphate reduction has been shown to occur in the upper 1 cm of the sediment (Skyring & Bauld 1990). The redox conditions of the sediments in the OTC treatment did not appear to be different from that in the Control treatment, as the Eh of both treatments  73  70 65 60 55 -E 50 =  C■2  o Blank • Control 7 OTC  45 -2 40 = cl) 35  E  30 25 20 15 10 = 5 I^'^I^I^I^  I^'^1^'^1'  1^I  0^2^4^6^8^10^12^14^16 Time (Days)  ^  18  ^  20  Figure 2.11. Comparison of nematode abundance (numbers•cm -2 ) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean ± 1 Standard Deviation.  • • 74  o Blank • Control v OTC  24  E  20  a.) 16  •  z  12  I ^'  0^2^4^6^8^10^12^14^16^18^20  Time (Days)  Figure 2.12. Comparison of harpacticoid copepod abundance (numbers•cm -2) over 20 days, between the Blank Treatment (fish waste and OTC not added), Control Treatment (fish waste added but not OTC), and OTC Treatment (fish waste and OTC added). Plotted values are the Mean + 1 Standard Deviation.  75  approached -200 mV in value by the end of the experiment. These values were considerably less reducing than found in the Control and OTC treatments (Fig. 1.6, Chapter 1), approaching -320 mV. However, the pattern of changes to the redox profile between the 2 studies were the same, with the sediment becoming more reducing at the surface than at depth, and the RPD layer shifting closer towards the surface. The redox potentials found in this study were very similar, to the Eh of sediments beneath fish farms in Scottish sea lochs at -185 and -200 mV (Brown et al. 1987 and Lumb & Fowler 1989, respectively). This would suggest that with regard to redox potentials, the addition of fish wastes to the sediments in the microcosms closely matched the input of fish wastes from aquaculture sites to the benthic environment, and sulphate reduction remained the primary route in the biodegradation of fish waste. The fish waste that was cumulatively added in this study was approximately 1.24 kg•m-2 at the end of the experiment, with a total of 1.06 kg•m-2 in the first study. As the sediments were less reducing than previously found (the Eh at -200 mV instead of -320 mV in Chapter 1) and the white microbial mat was present throughout the experimental period, this would suggest that the sulphate in the fish waste (and thus the H2S produced by sulphate reduction) remained available throughout the experiment. The sedimentation rate of the fish waste was insufficient for methanogenesis to have occurred. With continued accumulation of the fish waste in the sediments, greater reducing conditions could eventually have been reached. Perhaps the sulphate was not exhausted in the fish waste due to a diminished metabolic rate of the anaerobic microbial community in response to the gradual addition of the fish waste, or the stimulatory effect of the grazing pressure exerted by the bacterivorous protozoan population was not as great. That the OTC did not appear to make any difference in the redox of the sediments between the Control and OTC treatments could have been due to an insufficient quantity of OTC to affect the microfauna.  76  2.5.2 MICROBIAL MAT  The morphological changes of the microbial population in the fish waste are remarkably similar to those described by Gonzalez & Biddanda (1990), during microbial transformation of isopod (Idotea granulosa) feces. The large colonial cells in the white microbial mat and filaments that formed by day 16 could have been due to the sulphur oxidizing bacteria Beggiatoa spp. For instance, the large cells were observed to be from 38 Am in length, while the tufted filaments were approximately 10 Am wide and up to 2 mm long with limited motion. These cellular dimensions are within the range described for Beggiatoa spp. (Jorgensen 1977a) in which cell sizes range from 4 to 26 Am long, with  filaments composed of hundreds of cells up to 1 cm long and 50 Am wide. As well, the filamentous strands were similar in appearance to the Beggiatoa cultures presented by Strohl & Larkin (1978) although the "gliding" motion was not observed. Other colourless sulphur oxidizing bacteria that form white films or veils are the Thiovulum spp., but these are large, spherical cells (8-16 Am wide; Wirsen & Jannasch 1978, Jorgensen & Revsbech 1983). While filamentous cyanobacteria have also been described from benthic environments (Kuenen et al. 1985, Caumette 1989), the limited amount of light in the tanks would preclude this possibility.  2.5.3 DISSOLVED OXYGEN  The biological oxygen demand (BOD) of the sediments should have resulted in the dissolved oxygen (DO) levels in the Blank treatment being higher than in the Control and OTC treatment tanks in which fish waste was placed. Yet the DO level in the Blank treatment was significantly lower than in the Control treatment (Tukey Test, P > 0.05, Fig. 2.7). This reversal in the expected change of DO was due to the different design of the microcosm tanks used in the Blank treatment and the Control and OTC treatments. For instance, the seawater supply for the 200 L Control and OTC treatment tanks was provided directly from pipes (Fig. 2.1), while the seawater supply for the Blank treatment tanks had  77  to be piped from a further distance and settled in a header tank before emptying into the 20 L tanks. Measurements of DO levels in seawater from the Blank treatment (66.46+1.87%) were significantly different from the Control and OTC treatments (78.72+0.56%; P <0.05, T-Test). Perhaps some BOD may have occurred in the header tank as a great deal of phytodetritus was observed to have accumulated in the header tank. If this 12% difference in the DO of seawater was accounted for in the Blank treatment tanks, then the dissolved oxygen levels would be higher overall in the Blank treatment as expected, although similar fluctuations in the DO levels of seawater for all treatments had occurred. The sharp decrease in the DO levels of seawater on day 4 following the addition of fish waste was not due to the BOD from the sediments, as this drop in DO also occurred for the Blank treatment which did not have any fish waste added. The DO levels between the Control and OTC treatments were not significantly different from each other. The oxygen demand from sediments impacted by fish waste can be much greater than from nonpolluted sites (0.74 g.d - 1 .m -2 compared to 0.10 g.d - l.m -2 , Aquametrix 1990), which should result in noticeable differences in the dissolved oxygen content of bottom seawater. As the differences in DO levels were minimal between the treatments in this study, it can be concluded that the microcosm systems were flushed with well oxygenated seawater at all times.  2.5.4 OXYTETRACYCLINE  The total amount of OTC that was added to the fish waste in this study was 0.48 g (spread over 4 days) or 0.73 g•1 -1 total, compared to the 2.53 g and 10.13 g that were added to Treatment A and B of the first study all at once, respectively. A yield of 13.82 ± 22.17 ppm OTC (mean ± 1 standard deviation, n=3) was obtained on day 10 from the OTC treatment, and the residual amounts of OTC detected throughout the rest of the study could have remained within the sediments for several months (up to 419 days, Bjorklund et al. 1990 & 1991). The majority of the OTC likely washed out of the fish waste during  78  application, with continued diffusion and leaching from the sediment (Samuelsen  et al.  1988). OTC was not detected in water samples taken from near the sediment surface. As discussed in Appendix B, OTC levels were underestimated due to difficulties with the extraction procedure. The redox profile of the OTC treatment appeared to be the same as in the Control treatment, again suggesting that the quantity of OTC applied to the sediments was insufficient to inhibit the bacterial or protozoan community.  2.5.5 MICROFAUNAL ABUNDANCE  Bacteria Total cell counts in sediments not impacted by fish waste remained constant at 5 7x10 9 1 -1 sediment (Blank treatment, Fig. 2.8), with elevated bacterial numbers up to 1.4x10 10 .g -1 in the Control and OTC treatments with fish waste. Microbial abundance from sediments polluted by fish waste in this study (1.4 - 2.6x10 10 1 -1 ), were several times higher than in previous studies (0.03 - 1.01x10 10 .g -1 ; Enger et 1988, Torsvik et  al. 1989, Samuelsen et al.  al. 1988, Carr & Goulder 1990), which had not used ultrasonication to  loosen attached sedimentary bacteria.  Ciliates Initial ciliate counts were fairly low, < 50 individuals•cm -2 , composed of the larger 50-300 gm ciliates. It is not unusual to find low numbers of ciliates in organically poor sediments (Parker 1981) which could have been the situation in this study, where the sand was from the intertidal zone was strongly impacted by wave and tidal action. The application of oxytetracycline did not affect ciliate abundance between the Control and OTC treatments, as both treatments exhibited bimodal peaks in abundance from a small ciliate, 20-40 gm in length (Fig. 2.9a). In both treatments, the first peak in abundance was the highest but 2 days apart, at 35x10 3 individuals•cm -2 on day 6 in the Control treatment, and  79  26x10 3 individuals•cm -2 on day 8 in the OTC treatment. The second peak in ciliate abundance occurred on day 12 in both treatments, 14x10 3 and 22x103 individuals•cm -2 in the Control and OTC treatments, respectively. As data on protozoan abundance in sediments beneath fish farms have not been previously collected, it is difficult to assess whether the ciliate numbers observed in this study were within the range for sediments polluted by fish waste. In comparison, ciliate numbers in different benthic habitats can range from a few hundred to several thousands individuals•cm -2 (e.g. Kemp 1990), which would suggest that the values for ciliate abundance observed were much greater than usually found. In exceptional cases, ciliate numbers can reach as high as 75x10 3 -cm -2 in organically enriched sediments (Finlay 1978). This is within the range of values found in this study. As both treatments exhibited the same oscillations in ciliate numbers (rapid increases and decreases in abundance) these changes could not be attributed to the OTC. The 2 day lag by the ciliates in the OTC treatment to reach the first maxima on day 8 (compared to day 6 in the Control treatment), was not due to the OTC either as this lag period occurred before the addition of OTC. The lag period was thus due to natural variations within the benthic ecosystem. As the concentration of OTC in this study was over 5 times less than in Treatment A of the previous study (in which the protozoan community was not considered to have been inhibited, Section 1.5.7), the amount of OTC in this study is thus considered to have been insufficient to have inhibited the ciliate community. The OTC also could have been diluted to a greater extent during the gradual addition of the OTC over 4 days, compared to the addition of OTC at one time. The low concentrations of OTC found in the sediments beneath fish farms after 12 months (up to 5 ppm; Bjorklund et al. 1990, Nygaard et al. 1992), thus may not be sufficient to inhibit bacterial or protozoan activity.  80  Ciliate Diversity In general, macro-, meio- and microfaunal populations exhibit similar changes in species diversity, numbers and biomass to different sources of organic pollution (e.g. Pearson & Rosenberg 1978, Weston 1986, Gowen & Bradbury 1987, Ritz et al. 1989). When protozoan communities are environmentally stressed (i.e. by a pollutant) there is an elimination in sensitive species and an increase in the numbers of tolerant species, leading to higher numbers and lower species diversity (Atlas 1984). Qualitative changes in this study showed that the larger ciliates ( > 250 Am) disappeared quickly with organic enrichment of the benthos, and the protozoan community soon became dominated by the smaller, much more numerous ciliates. While the diversity of the protozoans increased from a few morphologically distinct ciliates to several distinct ciliates during the first 2 weeks of sediment enrichment, this declined to only a few ciliate types again (although these ciliates were different than those initially present at the start of the experiment). Species numbers are higher in transition zones where deposition of nutrients are occuring (Pratt & Cairns 1985). Studies of protozoan colonization in nutrient enriched systems (Pratt et al. 1987) have shown that colonizing protozoans often "overshoot" eventual species equilibrium, where a species maximum can be reached within a few days followed by a decrease in species numbers after longer intervals. The seawater-ice technique for extracting benthic ciliates was not considered quantitative, as observations of drops of unpreserved sediment would often show some types of ciliates that were not present in the preserved samples. While nearly all benthic ciliates can be extracted with this technique (Fenchel 1969), extraction efficiences have been found to be low with underestimations of ciliate abundance by over 90% (Alongi 1986). It is possible that the number of ciliate species was underestimated, as it would have been easy to miss the rare species. While the diversity of ciliate species has been shown to decrease in response to agricultural and domestic pollution with an associated increase in numbers (Hul 1987 & 1988, Grabacka 1988, Kosciuszko & Prajer 1990), a higher ciliate diversity, abundance and  81  biomass was found in the sediments of a Scottish Loch polluted by effluent discharge from a pulp and paper mill (Wyatt & Pearson 1982). This greater ciliate diversity was attributed to not only the increase in available prey (with the sulphur bactera serving as the major food source), but also to the structural changes in the sediment surface layers from the cellulose fibre, that provided a greater number of potential niches available for colonization.  Heterotrophic Microflagellates Microflagellate abundance remained relatively low in the sediments throughout the study period (Fig. 2.9b, < 1x105 individuals-cm-2), until after the ciliate population declined on day 14. A bloom of the heterotrophic microflagellates occurred in the Control and OTC treatments, to a maximum of 12.5x10 6 individuals-cm-2 on day 18, and were absent in the Blank treatment. The addition of oxytetracycline did not affect microflagellate abundance between the Control and OTC treatments. Comparison with flagellate abundance in other types of sediment indicates that the values from this study were high (Kemp 1990) but not exceptionally so (Fenchel 1975, Kemp 1988).  2.5.6 MICROFAUNAL INTERACTIONS  The dynamics of protozoa associated with the degradation of fish waste indicated a predator-prey relationship with bacteria (Gonzalez & Biddanda 1990). There was a positive correlation between the initial peaks in ciliate and bacterial numbers (Fig. 2.8 and 2.9a) with ciliate abundance in the Control and OTC treatments occurring 2 and 4 days, respectively, before the peak in bacterial abundance. This suggested that bacterial numbers did not peak as quickly in the previous study (Chapter 1) partly due to this grazing pressure. Other causes may have been the differences in the application of fish waste and OTC, which was added gradually in this study instead of all at once; or that the fish waste was stored frozen prior to use in this study, but was added to the sediments without freezing in the previous experiment. The greater amount, and the unfrozen state of the fish waste  82  would have allowed a greater proportion of the bacterial and protozoan community to have been initially present in the first study (Chapter 1), which could have resulted in a different microfaunal community developing with different responses to the benthic ecosystem and OTC than in this study. While ciliate numbers always decreased after reaching a maxima, bacterial abundance never peaked again and was likely continually grazed back down (Fenchel 1982). The second, smaller peak in ciliate numbers may reflect a decrease in bacterial prey availability. Although bacterial abundance remained depressed, bacterial activity could have been high, which has been shown to be stimulated by macro-, meio- and microfaunal grazing (e.g. Barsdate et al. 1974, Fenchel & Jorgensen 1977, Gerlach 1978, Finlay 1978, Tietjen 1980, Montagna 1984). A combination of the fish waste (as a supply of organic matter) and the grazing effects exerted by the bacterivorous ciliates appear to be the controlling factors among the microfauna (i.e. a substrate, substrate and grazing, and grazing control; Weisse & Scheffel-Moser 1991, Bak & Nieuwland 1989). The rapid response of the bacterial and ciliate populations also implies that growth and productivity of the microfauna in sediments organically enriched by fish waste can be high.  While benthic ciliate bacterivory was potentially important in controlling bacterial production in sediments enriched by fish waste, most studies have indicated that of the 5 to 50% of benthic bacterial production that could be consumed by protozoans, most grazing is attributed to microflagellates (e.g. Barsdate et al. 1974, Fenchel 1975 & 1986, Kemp 1988 & 1990, Patterson & Fenchel 1990). As a microflagellate bloom occurred quickly after the ciliate bloom ended on day 14 (Fig. 2.9b), the microflagellates thus became the primary grazers of bacteria. Kemp (1990) suggests that microflagellates should ingest a large fraction of bacterial production only when the relative abundance of microflagellates to bacteria is on the order of 1:1000. Comparison of bacterial and microflagellate abundance can be easily accomplished by transforming bacterial abundance (Fig. 2.8) to cells•cm-2. The mean sediment dry weight for the 1 ml bacterial cores is 0.697 g (0.154 g standard  83  deviation) and the surface area is 0.126 cm 2 . For instance, bacterial abundance at 6 14x109 •g-1 sediment can also be presented as 3.32 to 7.74x10 10 •cm-2 . As the microflagellate:bacteria ratio of 1:1000 is met (8 - 12x10 6 flagellates.cm-2 : 3.32 7.74x10 10 cells•cm-2 ), there is a potential capacity for these heterotrophic microflagellates to be important bacterivores. Existing data suggests that microflagellates are more likely to be important in organically enriched sediments, while ciliates may be more important in the interstitial spaces of sandy sediments (Kemp 1990). Interactions within the protozoan community can also occur, as ciliates can feed on flagellates (Fenchel & Finlay 1990), and at least 1 ciliate was observed to contain a few microflagellates (personal observations, this study).  Interpretation of how the microfauna interact also depends on sedimentary composition and the prevailing chemistry of the sediments (due to the bacterial activity resulting from the influx of organic materials). Oxygen does not penetrate as deeply in finer sediments (or in sandy sediments with a heavy organic load) and organic matter oxidizes slower (Fenchel 1967), with ciliates and protozoans adsorbing to the finer particles. Microbial activity is greatest near the surfaces, and the strongly reducing nature of the sediments also restricts protozoan populations to the surface layers of the sediment  (e.g. Fenchel 1969, Marty 1981, Joint et al. 1982, Novitsky & Karl 1986). Benthic protozoan abundance has been correlated to organic enrichment (and the resulting bacterial activity) and other physical factors such as the Eh and temperature of the sediments (Wyatt & Pearson 1982). It is interesting to note that the decline in ciliate abundance (after day  14, Fig. 2.9a) occurred when the sediments reached anoxic, reducing conditions by day 16 (Fig. 2.6 and 2.6). The small, 20 gm ciliates were thus dependent on an aerobic environment, if they were sensitive to reducing conditions. Perhaps ciliate abundance was . correlated to both the increase in bacterial abundance and the development of strongly reducing conditions. As the characteristics of the sand used in the experiments outlined in  84  Chapters 1 and 2 were mud/silt and sand, respectively, the initial microfaunal community may have been different. However, the heavy organic loading and resulting reducing conditions in the sediments would have restricted the bacterial and protozoan communities to the surface of the sediments in both studies.  2.5.7 DIATOMS  Diatoms (mainly of the pennate variety) increased in abundance from 8x10 3 cells•cm -2 to 14 - 24x10 3 cells•cm -2 (Fig. 2.10), but there was no difference in abundance between the three treatments. While nutrients (inorganic phosphorous) released from cage fish farming activites in rivers and lakes have stimulated the growth of benthic diatom communities downstream of the fish farms (Carr & Goulder 1990, Stirling & Dey 1990), high concentrations of pollutants will generally reduce diatom species diversity and abundance (Atlas 1984, Kwandrans 1988). Lumb & Fowler (1989) found that diatoms were present in sediments away from fish farms, but declined and disappeared approximately 15 m away from the aquaculture sites. A gradient from diatoms to a white microbial mat (Beggiatoa spp.) occurred at this point, corresponding to the zone where the redox potential at the sediment surface reached < 0 mV. The appearance of nearly all of the diatoms in this study were colourless or with greatly reduced cytoplasms. Most of the diatoms could have been dormant or dead, as light was excluded from the microcosms. However, Wyatt & Pearson (1982) also reported the presence of living diatoms in similar conditions from sediments enriched by cellulose fibres, and in the presence of hydrogen sulphide. The viability of diatoms in highly reducing sediments is partly attributed to an ability to tolerate reduced levels of light for long periods, and perhaps to function as facultative or obligate saprophytes (Lackey 1961), or live heterotrophically in the dark on dissolved glucose or lactate (Fenchel 1969). While diatoms may be able to survive under these conditions, active cells were not observed in live sediment samples, or undergoing division. The diatoms could have been mostly neritic  85  species, having settled out of the photic zone (Lackey 1961, Wyatt & Pearson 1982). The absence or low abundance of diatoms under polluted conditions (e.g. Lumb & Fowler 1989, Kwandrans 1988, Wyatt & Pearson 1982), would tend to confirm that entire algal communities can be eliminated if pollution becomes too extreme (Atlas 1984). There is thus a potential for error in this assessment of diatom abundance, as there is a strong possibility that dead cells were also counted. Recalling that the source of seawater for the microcosms was bottom water from Vancouver Harbour, and the outlet for the drainage pipe in the microcosm tanks was at the water surface instead of at the bottom, the slight increase in diatom abundance could have been due to an accumulation of cells rather than from an active and dividing diatom population. Possible trophic interactions could have occurred, with some species of the diatoms serving as prey for the ciliate and nematode community (Wyatt & Pearson 1982, Findlay 1982), but the poor condition of the diatoms would suggest that in this study, diatoms were not an important source of carbon.  2.5.8 MEIOFAUNA  Nematodes and harpacticoid copepods can respond quickly to pollution, forming the basis for the nematode/copepod ratio in organic pollution studies (e.g. Raffaelli & Mason 1981, Raffaelli 1987), with a subsequent decrease in species diversity (Hockin 1983). In general, nematodes would increase in abundance along an increasing gradient of organic pollution, but would decrease at the highest levels of enrichment. Mesobenthic (interstitial copepods) will decrease with increasing pollution due to a combination of the clogging of interstitial spaces, low oxygen and increasing sulphide concentrations. However, epi- and endobenthic (surface and shallow burrowing) species are not as affected by pollution, and tend to increase with increasing organic enrichment, being able to swim up to water with higher dissolved oxygen levels (reviewed by Hicks & Coull 1983, Heip et al. 1985, Raffaelli 1987).  86  The results from this study would tend to support the expected trends that would occur in the harpacticoid copepod and nematode communities with organic enrichment (Gee  et al. 1985, Widbom & Elmgren 1988). Nematode density in the Control and OTC treatments remained lower than in the Blank treatment (55 individuals•cm -2 on day 16, Fig. 2.11), but abundance began to rise at the end of the study and decrease in the Blank treatment. Copepod abundance in the treatments with fish waste decreased, with the highest numbers reached in the Blank treatment at 24 individuals•cm -2 on day 16 (Fig. 2.12) although all treatments exhibited similar values by day 20. Weston (1990) found that nematodes and polychaetes contributed to over 99% of the total number of individuals in sediments directly beneath a mariculture site in Puget Sound (Washington, U.S.A.), with surface and near-surface-dwelling crustacean species (amphipods, cumaceans, isopods and ostracods) becoming more abundant along a decreasing gradient in pollution away from the fish farm. The presence of crustacean nauplii (assumed to be harpacticoid copepod larvae as no other crustacean species were observed) and copepods with egg cases throughout the study, were an indication of the ability of the copepod community to survive organic enrichment of the sediments, at the present dosage levels of fish waste and OTC. However, as there were no significant differences found between the different treatments due to the large fluctuations in abundance, an absolute pattern cannot be established. The OTC levels in this study did not appear to affect copepod and nematode abundance, but high concentrations (0.25 M OTC) had an immediate and toxic effect on the nematodes (personal observations). Crustaceans were also affected, but survived for a longer period of time. The long-term effects of low concentrations of OTC on meiofaunal communities has never been examined. Studies on the effects of OTC on dauer larva (a dormant stage) of the nematode Caenorhabditis elegans, indicates that there is a moderate to strong inhibitory effect on pharyngeal pumping activity at concentrations of 5 mM, with impaired incorporation of methionine into protein, delaying longitudinal growth (Reape & Burnell 1991). Lower concentrations of OTC slowed growth, but the developmental  87  pathway of the nematodes eventually became normal once treatment with the antibiotic ceased. While the OTC concentrations in this study did not appear to reach toxic or strongly inhibitory concentrations, the concentrations of OTC initially present in the fish waste during addition to the sediments (and perhaps at aquaculture sites during treatment of diseases with medicated feed), could slow growth down for a short period of time. Large populations of nematodes can still develop under these conditions though, as Weston (1990) found high densities of nematodes in sediments enriched by fish waste (90,507 individuals•m -2 ). Rotifers were not enumerated, and their contribution to the marine meiofauna is not large (Fenchel 1978) and is more important in brackish and freshwater systems (Sanders et  al. 1989). Weglenska et al. (1987) established that the rotifer community was important, in the bottom waters beneath rainbow trout cage cultures of Lake Glebokie (Poland) feeding on bacteria and plankton. Perhaps the impact of rotifer activity in sediments polluted by fish waste should also be assessed for future studies.  2.5.9 INTERACTIONS BETWEEN MICRO- AND MEIOFAUNA  Bacteria, ciliates and diatoms have been implicated as food items for nematodes and copepods, while dead meiofauna can serve as a substrate for bacteria and diatoms (e.g. Fenchel 1978, Alongi & Tietjen 1980, Montagna 1984, Rieper 1985, Eskin & Coull 1987, Stoecker & Capuzzo 1990). However, the importance of the microfauna as a potential food source for the meiofauna is debatable, as Montagna et al. (1983) found that the meiofauna did not respond to changes in potential food abundance. Alongi (1988) also suggested that protozoan and meiofaunal populations may not be tightly coupled to the dynamics of bacterial and microalgal communities in tropical intertidal habitats, and Wyatt & Pearson (1982) never observed the meiofauna to be feeding on the ciliate populations. Physical-chemical factors (temperature and Eh) have also been implicated in controlling meiofaunal abundance (Montagna et al. 1983), while Finlay (1980) decided that  88  interactions between the physical-chemical factors (Eh, pH, oxygen availability, daylight) and the biological factors (bacterial, ciliate, benthic chlorophyll-a and nematode populations) determined the distribution of the micro- and meiofauna (Heip et al. 1985). However, correlations between changes in the meiofauna, microfauna and Eh of the sediments were not apparent. The addition of fish waste to sediments increased bacterial and protozoan abundance, although the response of the meiofaunal community was less clear. While the meiofauna could have been feeding on the microfaunal community, potential interactions could not be concluded with certainty, as correlations between micro- and meiofaunal abundance were weak or not apparent. Trends in meiofaunal abundance may require a longer period of time to become apparent, as meiofauna will typically require a relatively greater period of time to respond to organic enrichment, due to longer life cycles (Hicks & Coull 1983, Heip et al.  1985). For instance, while the bacteria and protozoan populations greatly increased  during the experimental period of 20 days, the meiofauna may require several weeks to a few months to exhibit such large changes. The lower abundance of the nematode and copepod populations suggests that their impact on the microfauna will not be important until higher population densities are reached.  2.5.10 MACROFAUNA The presence of the nudibranchs and small polychaetes did not appear to disturb the sediment surfaces, as the redox profiles remained relatively consistent for all tanks of each treatment. The disappearance of the nudibranchs from the sediment surface may have been related to the development of anoxic, reducing conditions, or perhaps to reproductive activity. The presence of nudibranch egg patches would also suggest that prey items (coelenterates) were abundant (but the coelenterates were not collected or enumerated). Polychaete abundance did not approach the densities seen in the field (e.g. 1,079 individuals•m-2 compared to over 25,000•m -2 , Tsutsumi 1987) as the duration of this study  89  was too short for such large changes in polychaete abundance to occur. Perhaps over a longer period of time and with a greater amount of fish waste and potential prey (e.g. bacteria, ciliates, diatoms), the polychaete population could increase greatly. The impact of OTC on macroinvertebrates has not been considered before, but the high growth and productivity of macroinvertebrate populations (at least for C. capitella) in sediments around fish farms would suggest that toxic or inhibitory concentrations of OTC are not reached. It is reasonable to expect that OTC is present in the sediments beneath the fish farms during part or the entire year when the macrofaunal studies were taken, as the use of medicated feed during disease outbreaks at fish farms during the summer can occur more than once (Weston 1986), and low concentrations of OTC can persist for long periods of time in the benthic environment (e.g. Bjorklund et al. 1991). The activity of the sandlance should have resulted in bioturbation of the sediment, but this was not evident in any changes in the redox profiles. The sandlance was probably collected with the sediments from Spanish Banks, and its ability to survive implies that there was sufficient prey available (crustaceans, probably harpacticoid copepods). However, differences between copepod abundance of the 3 treatments were not significant. In addition, as macrofauna would require several months for large increases in abundance to become apparent (Mattson & Linden 1983), correlations should not be expected with the micro- and meiofauna, and the physical-chemical conditions over the experimental period of 20 days in this study.  2.5.11 PROTOZOANS IN BIOLOGICAL MONITORING  Ciliates have commonly been used in Europe and Asia as a biological tool in the evaluation of the degree of water pollution (saprobiological analysis, Sladecek 1973). Based on the dominating species present, several zones of pollution can be classified (e.g. Grabacka 1985 & 1988), although explaining the temporal variation in the frequency and abundance of individual species can be difficult. Finlay et al. (1981) suggested that instead of relying on subjective decisions of whether or not large fractions of the species present are  90  characteristic of the degree of pollution, classification techniques could be used to indicate the presence or absence of significant associations between known or defined protozoan communities in polluted areas (association analysis). The short-term response of ciliate populations (a few weeks) presents an ideal opportunity to utilize the changes in the ciliate communities as a rapid biological tool in identifying the organic pollution of sediments around aquaculture sites. Rather than waiting the required several months to identify severe pollution problems through the major structural changes in the meio- or macrofaunal communities, rapid and expedient measures can be taken to minimize the impact of fish waste on the benthos and environment. This study has shown that the use of ciliates as a biological monitoring tool, for identifying the pollution of sediments impacted by fish waste can be promising. However, a great deal of further research is required to identify the major components of the microfaunal community that exists beneath fish farms (e.g. bacterial and protozoan species and abundance), and potential interactions with the meio- and macrofauna.  91  2.6 SUMMARY  The main objectives of these 2 studies were to determine if and how the micro- and meiofauna would react to organic enrichment of the sediments by fish waste, and an additional response to the antibiotic oxytetracycline. These investigations have shown that:  (1)  The benthic bacterial population will greatly increase in magnitude, shortly after the addition of fish waste to sediments.  (2)  Protozoan populations will also increase rapidly along with the increase in bacterial abundance, perhaps mediated by the reduction-oxidation conditions in the sediments.  (3)^Interactions between the micro-, meio- and macrofauna may have occurred, although there were little or no correlations apparent. The duration of this study was considered to be insufficient to have allowed substantial changes in meio- and macrofaunal abundance to occur. (3)  Oxytetracycline was found to persist at low concentrations for relatively long periods of time, with minimal or no inhibitory effect on microbial communities (possibly due to the expression of oxytetracycline resistance). It is unlikely that high concentrations of OTC would occur in the sediments beneath fish farms, or would be found for long periods of time in the sediment due to the rapid dispersion of the majority of OTC in the marine environment.  (4)  The protozoan community did not appear to be affected by OTC levels that would be found at aquaculture sites treating fish with medicated feed. However, higher dosage levels of OTC may have inhibited bacterivorous protozoan grazing, resulting in an increase in bacterial abundance but decrease in metabolic activity, and overall slower mineralization of the fish waste.  92  (5)  ^  The rapid response of the microfauna suggests that protozoan communities would be promising as a biological tool in pollution assessment studies on the impact of fish  waste on the benthic environment. This in turn could lead to more rapid  identification of potential pollution problems and efficient management at fish farms to minimize these environmental problems.  2.7 RECOMMENDATIONS FOR FUTURE STUDIES  The effects of fish waste and oxytetracycline on the microfaunal (and meiofaunal) communities requires further work. Currently, research is strongly underway in studying the long-term behaviour and effects of oxytetracycline in the benthos and on the microbial community (e.g. Norway) and on macrofaunal communities. However, there is a lack of information on changes in meiofaunal and protozoan communities. While it was established that fish waste may increase microfaunal abundance, with a further potential effect by OTC, further research should concentrate on the following aspects to determine the interactions of the benthic communities in the degradation of fish waste.  ( 1)  If marine benthic protozoan populations are to be used as an indicator of organic enrichment (e.g. by association analysis), precise taxonomy is required, along with a technique that will provide an accurate assessment of abundance. Determination of short and long-term physiological and developmental responses of protozoans to differing concentrations of oxytetracycline is also needed. Measurements of bacterial growth, heterotrophic activity and OTC resistance, along with taxonomic studies should be taken to determine the distribution and abundance of potential prey items for the protozoan population.  93  (4)  A concurrent study of the meiofaunal community, with a complete assessment of abundance and species diversity to evaluate potential interactions with the microfaunal community.  (5)  The responses of the microfauna to organic enrichment by fish waste and OTC at a fish farm should be assessed, when physical and biological factors are not uncoupled (e.g. light, wave/tidal action, increased recruitment, macrofaunal interactions, etc.).  The field study should be long-term (over several months or years) to ensure that seasonal changes in the microfauna will not be misinterpreted. (6)^The physical-chemical and biological factors of the benthic communities should be monitored after cessation of activities at different culture sites, to determine optimal conditions for the recovery of the sediments.  While these recommendations would require a large amount of effort, these studies have been completed before in different areas and for different sources of organic enrichment. For instance, the ecology of the marine microbenthos has been previously established (e.g. Fenchel 1967 & 1969) and the ecological, biological and physical-chemical effects of cellulose degradation in the benthos have also been well documented (e.g. Pearson 1982 and references therein). Benthic trophic interactions and nutrient studies on mangrove litter in tropical coastal regions have also been completed (e.g. Alongi & Tenore 1985, Alongi 1988, Alongi et al. 1989, Alongi 1990). While the required research on the ecological, biological and physical-chemical questions associated with fish waste from aquaculture may require several years and involve many groups and individuals, this would not be an impossible task.  94  APPENDIX A EXTRACTION OF ATTACHED BACTERIA IN SEDIMENTS BY ULTRASONICATION  INTRODUCTION  The enumeration of bacteria from sediments by ultrasonication or homogenization have been shown to be the most effective methods in releasing the attached bacteria (e.g. Montagna 1982, Ellery & Schleyer 1984). The adhesion of bacteria to sediments are a combination of physicochemical processes (Van Loosdrecht et al. 1989, Kemp 1990). These processes involve electrostatic attraction such as Van der Waals forces, coulombic attraction (Krone 1978), hydrophobicity of cells (Fattom & Shilo 1984, Rosenberg & Kjelleberg 1986) as well as special appendages (e.g. pili or fibrils) or polymers produced by bacteria to attach to surfaces (e.g. Marshall et al. 1971, Weise & Rheinheimer 1978, Shilo 1989, Mir et al. 1991). Up to 33% of bacteria can be left behind on sand grains after homogenization (Newell & Fallon 1982) and Ellery & Schleyer (1984) demonstrated that ultrasonication can be significantly more effective than homogenization in separating attached sedimentary bacteria (using a Decon FS 100 ultrasonication bath sweeping through a frequency range of 40 to 50 kHz, power output between 100 and 200 W). Dye (1979 & 1983) found that in sediments > 200 p,m, sonication was very effective in removing attached bacteria, whereas in silty/clay sediments homogenization proved to be better in removing attached bacteria. As the major size fraction of the sediments used in this study were in the 180-355 Am range (Chapters 1 and 2), ultrasonication was the technique used in this study to release bacteria attached to sediments. The sampling and subsampling variability in counting bacteria attached to sediments can be quite large (Montagna 1982). But the effect of strengthening cells with formaldehyde and dispersing them with a deflocculent (e.g. 'Tween 80 or pyrophosphate) and ultrasound can cause bacteria to be  95  randomly distributed, with subsequently lower variances between subsamples. For this experiment a Branson B-220 ultrasonication bath, sweeping through a frequency of 50 to 60 kHz and 125 W (at 117 volts) was utilized. The optimum ultrasonication time required for this ultrasonicator type was assessed prior to the study, and was found to be 90 seconds for this type of sediment.  MATERIALS AND METHODS  A sand sample was taken with a 1 ml syringe-corer and placed in a 20 ml scintillation vial with 3.55 ml of filtered seawater (0.22 gm pore size), and 0.25 ml of 37% formaldehyde (2.1 % formaldehyde, final concentration). The preserved sand sample was placed in a 125 ml Erlenmeyer flask and shaken vigorously with 80 ml of sterile, filtered (0.22 gm pore size) distilled water with 'Tween 80 added (1x10 -4 %) (supplied by Sigma Chemical Co., St. Louis, MO), following the technique of Ellery and Schleyer (1984). Ultrasonication was performed for 30 seconds. The sand sample was shaken again and the solution was allowed to settle for 30 seconds to allow the heavier sand grains to settle out, which has been determined not to affect abundance estimates of bacteria (Montagna 1982). A 1 ml subsample was then extracted from the solution. Ultrasonication times of 0, 30, 60, 90, 120, 180, 240 and 300 seconds were applied in series to the sand sample, with 1 ml subsamples taken for bacterial enumeration after each sonication period. Acridine orange epifluorescence microscopy was used to obtain direct cell counts from the diluted samples (Parsons et al. 1984).  To ascertain the extent of bacterial destruction by ultrasonication, 80 ml of sterile, filtered distilled water with 'Tween 80 was added to another formalin preserved sand sample. The sample was vigorously shaken and the larger sand particles allowed to settle out after 30 seconds. Eight 2 ml aliquots of the bacterial suspension were removed and placed in individual test tubes, and ultrasonicated in series for the same time intervals as the  96  previous experiment. Direct counts were performed following the aforementioned acridine orange epifluorescence microscopy technique. Statistical analysis was performed with SYSTAT (Ver 5.0, Evanston, IL). Square root transformations were required to meet the assumptions of normality and homogeneity of variance (Bartlett's test, P 0.05).  RESULTS AND DISCUSSION The effect of ultrasonication with time on bacterial abundance (numbers•g 4 sediment dry weight) was highly significant (ANOVA, p 0.05). An initial abundance of 9.8x10 7 cells•g -1 (Fig. 1) was observed with 4x10 8 cells•g -1 present after 90 seconds of ultrasonication. Ultrasonication for 90 seconds or more yielded consistent results, Tukey Test, p z 0.05, with a plateau in bacterial abundance after 90 seconds. The optimal ultrasonication time for extraction of attached sedimentary bacteria in this study was observed to be 90 seconds, rather than 2 1/2 minutes in the study by Ellery and Schleyer (1984). This may reflect differences in the ultrasonicator models, sediment sizes and types used, and bacterial types present. However, in determining their optimal ultrasonication time, Ellery and Schleyer (1984) did not appear to rigorously test for the peak in bacterial abundance (as indicated by a lack of error bars). For example, their optimal sonication time could have been less, if substantial overlap occurred between the standard deviation of means.  The destructive effect of ultrasonication is apparent in Figure 2. There is an immediate, significant decrease in bacteria (p <0.05, ANOVA) with increased ultrasonication time (as in Ellery and Schleyer 1984) to 77.5%. There was no signicant change in the percent of bacterial abundance for the remaining sonication times. The greater variation present at 180 seconds corresponds to the point when large amounts of bacterial fragments began appearing in the samples. For an ultrasonication time of 90  97  5  ao S.  a)  E z  2  1  0 0^30^60^90^120 150 180 210 240 270 300 Sonication Time (Seconds)  Figure 1. Accumulation of bacteria attached to sediments following ultrasonication times of 30, 60, 90, 120, 180, 240 and 300 seconds, of a formalin preserved sand sample. Values presented are the Mean ± 1 Standard Deviation, n=5.  98  100 — --- 90 — 80 — co  a) tn  70—  0  60 — c.) 50 —  a 40 — 30 0^30^60^90^120 150 180 210 240 270 300 Sonication Time (Seconds)  Figure 2. The destructive effects on a bacterial suspension extracted from the sediment, following ultrasonication times of 30, 60, 90, 120, 180, 240 and 300 seconds. Values presented are the Mean ± 1 Standard Deviation, n=4.  99  seconds, all bacterial counts should therefore be increased by 143% (standard deviation of 26%) to obtain a more accurate estimate of bacterial abundance in the suspension.  Two samples had an increase in bacterial abundance after 2 minutes. Direct observation of sand grains at this time still showed attached bacteria. This suggests that a longer ultrasonication time is required to release those bacteria still attached to the sediment. Although 90 seconds was taken to be the optimal ultrasonication time, it is apparent that another source of error may be causing an underestimation of bacterial abundance. To determine the fraction of bacteria (%) still attached to the sediment, sand samples were sonicated for 90 seconds (in a 1:200 dilution), 1 ml of the suspension removed for counting and the rest of the suspension discarded. The sand sample was rinsed twice and the rinses discarded each time, sonicated again (in a 1:100 dilution) for 60 seconds, 1 ml of the suspension removed for counting and the remaining suspension discarded. This procedure was repeated twice more for 60 seconds each time (Fig. 3) for a total sonication time of 270 seconds. There was an exponential decrease in bacterial abundance with increasing ultrasonication time, with approximately 65% of the bacteria present released after the first 90 seconds of ultrasonication. An additional 1 minute treatment released only 22% more bacteria, with the remaining 13% extracted during the next 2 minutes of sonication. This exponential release of bacteria would seem to indicate that there are portions of the microbial population that require greater ultrasonication times to detach from the sediments. This may be due to the presence of different types of adhesion processes used (e.g. Van Loosdrecht et al. 1989). For example, as bacteria in marine environments have negative charges on their surface membranes, in distilled water (low ionic strength) Van der Waal's attraction between bacteria and particles are weakened considerably (Krone 1978, Van Loosdrecht et al. 1989). This would allow the fraction of the sedimentary microbial population utilizing Van der Waal's forces to be removed quickly, while the bacteria that  ▪ 100  100  0.) 7:1 Q.)  rn  a O Co  10  7  0  E Log Y = 2.185 — 0.007X  1  2  •  r = 0.896  a) U  a)  U cu  4.  0.1 — 1^  '  ^  I^'^'^I^I^'^I^'^1  0^30^60^90^120 150 180 210 240 270 300 330 360  Sonication Time (Seconds)  Figure 3. The percentage of bacteria remaining on sediments, following ultrasonication periods of 90, 150, 210 and 270 seconds. The relationship is linearized by log transformation and fitted with a first order regression and 95% confidence limits, n=16.  101  are more firmly attached by appendages or polymers may require a longer ultrasonication time to be displaced. The initial events in the attachment of non-motile and motile bacteria primarily due to physical-chemical forces (Marshall et al. 1971, Scheraga et al. 1979). To relate this exponential decline in bacterial abundance with increasing ultrasonication time, log transformation of percent bacteria was required to linearize the curve (Fig. 3: r2 =0.896, highly significant, p .0.001):  LOG (Y) = 2.185 - 0.007(X)  This equation can be used to correct for the additional underestimation of bacteria not released during the optimal ultrasonication period of 90 seconds. For instance, with 90 seconds of ultrasonication approximately 3.90x108 bacteria•g-1 dry weight of sediment are released (Fig. 1). Destruction by ultrasonication is 1.68x10 8 bacterial-1 sediment (Fig. 2). Using the above equation (Fig. 3) 35.89% of the bacteria are still attached to the sediment at 90 seconds of ultrasonication. The cells released at 90 seconds (including those cells destroyed by sonication, for a total of 5.58x10 8 bacterial-1 sediment) only account for 64.11 % of the overall bacterial abundance. A further correction is therefore required, resulting in an additional 3.12x10 8 bacteria•g-1 sediment. The total bacterial count should therefore be 8.70x10 8 bacteria• -1 sediment, over twice the abundance of the initial bacterial counts at 90 seconds. It is not certain how prolonged ultrasonication may degrade those bacteria remaining attached to the sediment, and this correction does not account for that. It may be possible that an underestimation of bacterial abundance still exists. In comparison to the study by Ellery and Schleyer (1984) the second source for the underestimation of bacterial abundance was not calculated. This formula is valid only within the range tested for (Ricker 1984, Zar 1984), in this instance the sonication periods from 90 to 330 seconds.  102  The increase in bacterial abundance after an additional minute of sonication (after the initial 90 seconds) is not apparent in Fig. 1. It may be that the rate of destruction of cells by ultrasonication (from the microbial population released from the sediment in the first 90 seconds) is equivalent to the rate of release of bacteria after an additional one minute of ultrasonication. However, these rates of destruction and release have not been measured.  Although Ellery and Schleyer (1984) found a strong trend of decreased bacterial abundance with ultrasonication time, their results do not consider that the bacterial suspension they used was originally extracted from a homogenized sand sample. Bacterial cells may have been already damaged by 2 minutes of homogenization, thus more susceptible to destruction by ultrasonication. Losses caused by damage during extraction may also differ between sediment types as well, and these effects should be evaluated. Ellery and Schleyer (1984) sampled surface sediment from a lagoon, while the sediment in this study came from a 20 m depth in Vancouver Harbour (ranging from mud to fine sand). Examination of the sediment following 5 minutes of sonication shows negligible amounts of bacteria attached as in Ellery and Schleyer (1984), suggesting that for the sediment types examined for their study and in this one, 5 minutes of ultrasonication is sufficient to loosen the majority of the sedimentary bacteria.  The reasons for differences in optimum ultrasonication times may be due to extraction techniques, differences in sediment types as well as components of the microbial population, and types of ultrasonicators used. The use of ultrasonication in extraction of attached sedimentary bacteria to find the optimal ultrasonication time (and correction factor due to subsequent destruction) should be determined for each study.  103  APPENDIX B EXTRACTION OF OXYTETRACYCLINE FROM SEDIMENTS BY HIGH-PERFORMANCE LIQUID CHROMATOGRAPHY. A CALIBRATION AND RECOVERY STUDY.  INTRODUCTION  In aquaculture, oxytetracycline (OTC) is a widely used broad-spectrum antibacterial drug, active against Gram-positive and Gram-negative microbes of both aerobic and anaerobic species (Neu and Caldwell 1978). The primary effect of tetracyclines is thought to be inhibition of protein synthesis, at the 30S ribosomal site (Levy 1984). Oxytetracycline is usually administered orally to fish through medicated feed. However, this method of treatment results in a very low amount of OTC being taken in by the fish (Bjorklund et al. 1991). Reduced feeding by diseased fish due to decreased feed intake is due in part to the effects of the infection and to the unfavourable taste of the OTC medicated feed. Feed intake has been shown to be reduced to 61 % of normal (Hustvedt et  al. 1991), to complete refusal of medicated feed (Salte 1982) in healthy rainbow trout (Oncorhynchus mykiss) when OTC was present in the feed pellets. When treating farmed fish, the decrease in feed intake is compensated for by increasing the dosage of OTC per kilogram of biomass. Concentrations range from 100 mg OTC/kg fish (about 200 mg OTC/kg feed, Dr. E.L. Dahl, Argent Laboratories Ltd., Richmond, B.C.) to 4000-16000 mg OTC/kg feed (Grondel et al. 1987). This obviously results in a greater proportion of the uneaten feed pellets and oxytetracycline collecting on the benthos. Of the OTC ingested by the fish, the majority of the OTC remains unabsorbed. Cravedi et al. (1987) estimated a bioavailability of 7-9% for rainbow trout in freshwater, while Rogstad et al. (1991) found a maximum absorption of 2.6%, 72 hours after administration of medicated feed to rainbow trout. Grondel et al. (1987) found absorption levels to be only 0.6% for carp after 25 days (Cyprinus carpio L). The OTC is excreted to  104  the environment in the feces (Rogstad et al. 1991) and also accumulates in the benthos. The low bioavailability within the fish may be due to the high affinity of OTC to form complexes with di- and trivalent cations (Lunestad and Goksoyr 1990, Grondel et al. 1987), with the OTC binding strongest to the bones and tissues of fish (Rogstad et al. 1991). The time required for complete elimination of OTC from fish is temperature dependant (Jacobsen 1989). Times for rainbow trout range from 60 days for temperatures > 10°C to 100 days between 7-10°C (Salte 1982, Salte and Liestol 1983). More recently, withdrawal times of 27 days at 16°C to 135 days at 5°C have been suggested for rainbow trout (Bjorklund and Bylund 1990). In comparison, Aoyama et al. (1991) could not detect any OTC in chinook salmon muscle tissue 35 days after treatment of medicated feed (at water temperatures of 7-10°C). This means that OTC can still be released in fish feces for a long period following treatment, and is bound to particles and sediments beneath the fish farm (e.g. Samuelsen 1989, Bjorklund et al. 1990). It has not been until recently that the degradation of oxytetracycline in the environment has been examined (Samuelsen et al. 1988, Samuelsen 1989, Jacobsen and Berglind 1988, Bjorklund et al. 1990, Bjorklund and Bylund 1991). The half-life of OTC in sediments ranges from 9 to 419 days depending on anoxic conditions, with a rapid halflife in seawater (5 to 16 days). The objective of this study is to examine the effects of fish waste and oxytetracycline on the microbenthic community. The methodology for the extraction of OTC from sediments was adapted from Jacobsen and Berglind (1988). A calibration and recovery study of this technique was simultaneously run, to evaluate the sensitivity and replicability of the method.  105  MATERIALS AND METHODS CHEMICALS  Oxytetracycline hydrochloride (OTC) and tetracycline hydrochloride (TC) were obtained from Sigma Chemical Co. (St. Louis, MO). Oxalic acid was from Fisher Scientic Co. (Fair Lawn, NJ) while methanol and acetonitrile (Omnisolv, HPLC grade) was obtained from BDH (Toronto, Ont., Canada). For the Na2-EDTA/Mcllvaine buffer, ethylenediaminetetraacetic acid disodium salt (NA2-EDTA) was from BDH (Toronto, Ont.), citric acid monohydrate from Fisher Scientific Co., and the sodium phosphate dibasic from Matheson Coleman & Bell (Ohio). Pure OTC is not used for treatment of the food pellets, but is supplied as Oxysol 440 by Argent Chemical Laboratories Inc. (Richmond, B.C., Canada) where 227 g of Oxysol 440 contains 100 g OTC.  APPARATUS  The HPLC system that was utilized for the calibration and recovery study, and for the second year of this study was a Shimadzu SIL-9A autoinjector coupled to a Shimadzu SPD-6A UV Spectrophotometer Detector and a Beckman 110A pump. Integration of the areas under the curves was performed by a Shimadzu C-R3A Chromatopac. The analytical column was a 254 x 4.6 mm I.D. RP-8 column, packed with Adsorbosphere-5 pm particle size. The prefiltering element was a direct connect universal guard column cartridge holder fitted with guard column cartridges packed with C8 5/2 Adsorbosphere (supplied by Alltech Associates Inc., Deerfield, IL).  CHROMATOGRAPHY  The mobile phase consisted of a methanol-acetonitrile-0.01 M aqueous oxalic acid solution (1:1.5:2.5) (v/v/v), pH 2 at a flow rate of 1 ml•min -1 . This solution was degassed under a vacuum prior to use, to eliminate bubbles from entering the column. Distilled and de-ionized water was used to make all solutions, which were then filtered (0.22 /AM  106  Millipore filters). HPLC grade methanol and acetonitrile were also filtered before use. The wavelength of the UV detector was set at 365 nm, rather than 350 nm suggested by Oka et al. (1985) and Jacobsen and Berglind (1988), as this wavelength has been found to be more sensitive to OTC (McErlane, pers. comm.). Pure OTC was used to make a stock solution of 500 Ag.m1 -1 OTC, by dissolving 5 mg in 10 ml of methanol. Differing volumes of serial dilutions were made with methanol to give concentrations of 0.5, 1.0, 4.0, 10, 20 and 30 Ag•m1 -1 (adapted from Aoyama et al. 1991). Pure TC was used as an internal standard, for the calibration curves and in the recovery study. A stock solution of 200 Ag.m1 -1 TC was made by dissolving 2.0 mg in 10 ml methanol, and a 10 Ag.m1 -1 working solution was obtained by diluting 0.5 ml of the TC stock solution to 10 ml with methanol. Enough of the internal standard TC was added to the samples for a final concentration of 1.0 Ag•g -1 sediment (Aoyama et al. 1991). The recipe for the Mcllvaine buffer (Oka et al. 1986) required for the extraction of OTC from sediments was provided by Aoyama (pers. comm.), and was made by dissolving 28.41 g dibasic sodium phosphate in distilled water in a 1 L flask and diluting to volume. After dissolving 21.01 g citric acid monohydrate in distilled water and diluting to volume, the 1 L citric acid solution was mixed with 625 ml of the sodium phosphate solution in a 2 L flask, pH of 4.0. The 0.1 M Na2-EDTA/Mcllvaine buffer was prepared by adding 60.49 g EDTA to the 1.625 mL Mcllvaine mixture.  CALIBRATION CURVE AND RECOVERY STUDY  A calibration curve was needed to calibrate the UV detector and determine if the tetracycline internal standard and oxytetracycline can be accurately measured over the concentrations that would be present in the samples. If the OTC and TC are not interacting with each other or other compounds which may be present, there should be a linear relationship (Bjorklund 1988, Bjorklund et al. 1990). For an internal standard  107  concentration of 1.0 gg•g -1 sediment 0.5m1 TC was added to 5 g sediment, and 0.5 ml of the following OTC standard solutions was also added to give concentrations of 0.05, 0.1, 0.4, 1, 2 and 3 ppm sediment. The calibration curve was made by plotting the ratios of the peak ares of OTC to the peak areas of the internal standard against the known concentrations of OTC. To determine the efficiency of the extraction technique for this type of sediment, a recovery study was made by adding 0.5 ml of 1, 10, and 20 Ag•m1 -1 of the OTC stock solutions to 5 g of dried sediment, but without an internal standard added. The area of the curves for the OTC extracted from these samples were plotted against the area of the curves for 0.5 ml of the unextracted OTC standard solutions, (adapted from Aoyama et al. 1991).  EXTRACTION PROCEDURE  The sediment sample was removed from the scintillation vial, the volume was measured in a 10 ml graduated cylinder and dry weight calculated based on the volume to dry weight ratio. Tetracycline was then added as the internal standard, to a concentration of 1 Ag.g4 sediment (adapted from Aoyama et a/. 1991). The sample was then rinsed into a 50 ml polypropylene centrifuge tube with 15 ml of the 0.1 M Na2-EDTA/Mcllvaine buffer, and homogenized with a variable speed Tri-R stirrer (Model S63C, 12000 rpm, 115 V, Tri-R Instruments Inc., Rockville Centre, N.Y.), at 5000 rpm for 1 minute (adapted from Jacobsen and Berglind 1988). A 5 ml aliquot of Na2-EDTA/Mcllvaine buffer was added and the sediment homogenized for 1 minute, after which another 5 ml of the Na2EDTA/McIlvaine buffer was added and the sediment homogenized for another minute. Two ml of the Na2-EDTA/Mcllvaine buffer was used to rinse the stirring rod, and the 2 ml buffer was also collected in the 50 ml centrifuge tube. The sample was then centrifuged for 15 minutes at 2500 rpm (Refrigerated Centrifuge model PRJ, International Equipment Co.). After centrifugation, the supernatant was decanted into a 15 ml filtration tower, and filtered through a 0.45 Millipore filter. The  108  filter tower was then rinsed and filtered with 1 1/2 ml Na2-EDTA/Mcllvaine buffer. After a 60 ml syringe was used to collect the filtered supernatant, the flask was rinsed with an additional 1 1/2 ml Na2-EDTA/Mcllvaine buffer, and the buffer rinse was also collected in the 60 ml syringe. An Alitech C18 900 mg Maxiclean cartridge was activated with 10 ml of methanol and 10 ml of distilled water, and the syringe attached to the female end of the cartridge. The supernatant was drawn through the cartridge by vacuum filtration at the recommended rate of 1.5 drops per second, maximizing the retention of OTC by the cartridge. The cartridge was then washed with 20 ml distilled water, and the OTC was eluted by passing 10 ml 0.01 M methanolic oxalic acid solution through the cartridge and collecting the methanolic-oxalic acid in a 20 ml scintillation vial. All samples were stored at -12°C for approximately a week, until ready for analysis.  RESULTS AND DISCUSSION The peak for OTC eluted from the column at 7.49 minutes (standard deviation =0.05 minutes, n=6). The recovery of OTC from the sediments is exceptionally poor following this technique. Quantities below 1 ppm did not yield a recognizable peak, while at 1 ppm of OTC the efficiency of recovery was only 3.28%. Generally, there were no quantifiable peaks, although one OTC sample at 3 ppm did show a 0.68% recovery. For tetracycline, the recovery was even less from the sediment, with no peak recognized even at a concentration of 4 ppm. The calibration did not yield a linear relationship, but this can be attributable to the poor extraction efficiency. Although small peaks were present up to 1 ppm (increasing in peak height), the peak height remained approximately constant to the highest concentration of 3 ppm. This low extraction efficiency is unusual as Bjorklund et al. (1990) using a similar technique in extracting OTC from sediments had an 88.6% efficiency, and a lower extraction efficiency of 72.2% for tetracycline. Loss of OTC in the extraction process  109  could have occurred due to a number of causes. During filtration of the OTC-Na2McIlvaine-EDTA supernatant through the C18 cartridges, the flow rate is critical to optimize extraction of OTC, and the columns should not go dry during the procedure (Aoyama, pers.comm.). White et al. (1974) emphasized the presence of small amount of EDTA in the mobile phase prevent tetracycline antibiotics from complexing with the metal tubing of the HPLC colunmn (the formation of hydrogen bonds with active sites on the silica surfaces, The Supelco Reporter 1985). Ciarlone et al. (1990) noted the binding of tetracycline antibiotics to borosilicate glass and polypropylene labware, although for OTC and TC this was not a significant loss. Rapid degradation of OTC occurs when in solution. Cold storage is required, and even at -20°C some degradation still occurs (Aoyama and McErlane, pers. comm.). Dark storage is also necessary, as tetracyclines photodecompose easily (Samuelsen 1989, Oka et al. 1989). All samples were kept in dark, cold storage except during the extraction process, and the loss of OTC due to degradation is expected to be minimal (although this loss was not measured). It is not known if OTC may have been retained on the 0.45 pm filter, when filtering the supernatant following centrifugation. The filter was assumed to collect only particulate matter, allowing the OTC in solution to pass through. The large loss in extraction efficiency is still not clear, but may have occurred with the Mcllvaine buffer which is one of the critical steps in extracting OTC (Oka et al. 1985). An attempt to raise the pH of the Na2-EDTA/Mcllvaine buffer to 4 with HC1 could have destabilized the oxytetracycline (Onji 1984). An excessive amount of HCl may have been added, as the acidic buffer (pH 4.0) used to calibrate the pH meter was too acidic. Jacobsen and Berglind (1988) did not show the extraction efficiency of OTC from sediments, after adapting the technique from Oka et al. 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