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2013 The esr ponse of marine benthic fauna exposed to continuous and pulsed exposures of contaminated sediment: a study of avoidance and toxicity Daniel Ward University of Wollongong
Recommended Citation Ward, Daniel, The er sponse of marine benthic fauna exposed to continuous and pulsed exposures of contaminated sediment: a study of avoidance and toxicity, Doctor of Philosophy thesis, School of Chemistry, University of Wollongong, 2013. http://ro.uow.edu.au/ theses/4075
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THE RESPONSE OF MARINE BENTHIC FAUNA EXPOSED TO CONTINUOUS AND PULSED EXPOSURES OF CONTAMINATED SEDIMENT: A STUDY OF AVOIDANCE AND TOXICITY
A thesis submitted in fulfilment of the requirements for the award of the degree of
DOCTOR OF PHILOSOPHY
from
The University of Wollongong
By
Daniel Ward†,‡
Supervisors: Dr. Dianne F. Jolley† & Dr. Stuart L. Simpson‡
† School of Chemistry, University of Wollongong, Wollongong, NSW 2522 Australia.
‡ Centre for Environmental Contaminants Research, CSIRO Land and Water, Lucas Heights, NSW 2234 Australia.
September 2013
Certification
I, Daniel J. Ward, declare that this thesis, submitted in fulfilment of the requirements for the award of Doctor of Philosophy, in the School of Chemistry, University of Wollongong, is wholly my own work unless otherwise referenced or acknowledged. The document has not been submitted for qualifications at any other academic institution.
Daniel J. Ward September 2013
“Desire is the key to motivation, but it’s determination and commitment to an unrelenting pursuit of your goal – a commitment to excellence – that will enable you to attain the success you seek.” - Mario Andretti
Acknowledgements
I was 8 years old when I decided that I wanted a career in environmental science. I had no idea that the dream of becoming a scientist would take me on such a fulfilling journey - a journey that would not have been possible without the support, guidance and encouragement of so many wonderful people.
First and foremost, I cannot express enough the appreciation I have for my supervisors
Dr Dianne Jolley and Dr Stuart Simpson. None of this would have been possible without you and I thank you for mentoring me and giving me the opportunity to work with you. You have been my inspiration to become the best scientist I can be. Thank you for sharing your knowledge and experience with me; for motivating and encouraging me; for you advice, support and words of wisdom; and for your guidance during the early stages of my career.
A huge thank you also goes out to the staff at CSIRO for the advice and technical assistance that I have received throughout this project. To Dr Jenny Stauber and Dr
Graeme Batley for taking the time out of your busy day to provide me with invaluable feedback on many occasions. To Merrin Adams and Monique Binet for your assistance with maintaining a constant supply of algae to feed my critters. To Dr Anthony
Chariton for the numerous discussions on statistics we had and the advice you gave me. To Chad Jarolimek for your assistance with the sometimes temperamental ICP.
I would like to specially thank David Spadaro and Ian Hamilton for their constant support and assistance in the laboratory - the skills I gained through working with you have been invaluable and have made this project possible.
Thanks also to my friends and family. I have probably never really expressed just how grateful I am to have you all in my life. You have provided not only words of encouragement and motivation, but also a lot of fun and laughter along the way. You have certainly been an integral part of my mental wellbeing.
Many thanks to my parents, John and Irene, and my brother Matt for your love, support and motivation. Mum and Dad, you allowed me to dream and then did everything you could to help make that dream come true. You have always been so supportive and your unconditional love and belief in me gave me the courage to challenge myself and always strive to do my best. Without this, I would never have been able to complete my PhD, and for that I am eternally grateful.
And last but not least, I must thank my soon-to-be wife, Lauren. You have always been there to celebrate the highs and provide support during the lows of PhD life. You amaze me every day with your patience, compassion and understanding which has allowed me to complete this thesis. Thank you for your ongoing encouragement, motivation and love.
List of publications
Manuscripts
Ward, D. J., Simpson, S. L., Jolley, D. F. (2013). Slow avoidance response to contaminated sediments elicits sublethal toxicity to benthic invertebrates. Environmental Science and Technology 47(11): 5947-5953
Ward, D. J., Simpson, S. L. and Jolley, D. F. (2013). Avoidance of contaminated sediments by an amphipod (Melita plumulosa), harpacticoid copepod (Nitocra spinipes), and a snail (Phallomedusa solida). Environmental Toxicology and Chemistry 32(3): 644-652.
Ward, D. J., Perez-Landa, V., Spadaro, D. A., Simpson, S. L. and D. F. Jolley (2011). An assessment of three harpacticoid copepod species for use in ecotoxicological testing. Archives of Environmental Contamination and Toxicology 61(3): 414-425.
Presentations
Ward, D.J., Jolley, D.F., Simpson, S.L. (2011). The response of marine epibenthic fauna to short exposures to contaminated sediment: a study of avoidance and toxicity (ORAL). EnviroTox 2011, Society of Environmental Toxicology and Chemistry (SETAC) and the Royal Australian Chemical Institute (RACI), Darwin, Northern Territory, Australia, 17th – 20th April 2011.
Ward, D.J., Jolley, D.F., Simpson, S.L. (2010). The response of Melita plumulosa to continuous and pulsed exposures to contaminated sediment: a study of avoidance and toxicity (ORAL). 20th Annual Meeting: Society of Environmental Toxicology and Chemistry (SETAC) Europe, Seville, Spain, 23rd – 27th May 2010.
Ward, D.J., Perez-Landa, V., Jolley, D.F., Simpson S.L. (2008). An evaluation of the life cycles and suitability of three Australian benthic copepods for use in whole-sediment toxicity tests (POSTER). 5th Society of Environmental Toxicology and Chemistry (SETAC) World Congress, Sydney, New South Wales, Australia, 3rd - 7th August 2008.
Ward. D.J. (2008). An evaluation of the life cycles and suitability of three Australian benthic copepods for use in whole-sediment toxicity tests (ORAL). 16th Annual Royal Australian Chemical Institute (RACI) Analytical and Environmental Division R&D Topics, Sydney, New South Wales, Australia, 6th - 9th December 2008.
Ward. D.J. (2007). Toxicity of contaminants to the life cycles of benthic copepods (ORAL). 15th Annual Royal Australian Chemical Institute (RACI) Analytical and Environmental Division R&D Topics, Adelaide, South Australia, Australia, 9th - 12th December 2007.
Abstract
This study aimed to develop environmentally relevant chronic toxicity test methods to build on routine sediment quality assessment techniques currently in use. A series of experiments were used to assess the suitability of new Australian benthic invertebrate species for toxicity testing and investigate the use of chronic and behavioural endpoints to enable the development of more robust sediment quality guidelines.
Many researchers are evaluating representative benthic organisms for use in robust yet rapid toxicity tests to assess the sublethal and lethal effects of sediment contaminants. Four species of harpacticoid copepods, identified as Nitocra spinipes,
Tisbe tenuimana, Robertgurneya hopkinsi and Halectinosoma sp., were assessed as potential test species. The influence of diet on life cycle progression (development) was assessed and the use of a mixed tri-algal diet was found to be superior to the use of commercial fish food alone. Water-only bioassays showed that the times required to cause 50% lethality (LT50) were 24 (22-27) h at 50 µg Cu/L for T. tenuimana; 114
(100-131) h and 36 (32-40) h for 200, and 400 Cu/L, respectively, for N. spinipes, and
119 (71-201) h and 7 (4-10) h for 200 and 800 µg Cu/L, respectively, for R. hopkinsi.
96-h lethal effects thresholds were also determined for N. spinipes exposed to Cd, Cu,
Zn, ammonia and phenol in water-only exposures. Species were ranked based on ease of handling, culturing, rate of maturity, food selectivity and sensitivity to copper.
N. spinipes was found to be the most suitable species to use in sediment bioassays as it was robust, easily cultured and was sensitive to dissolved copper.
The avoidance response of M. plumulosa, N. spinipes and the snail Phallomedusa solida when exposed to contaminated sediments was investigated. Test vessels were
Page | i designed to allow the assessment of organisms moving between test sediments. Each species was observed to disperse evenly between test chambers that contained reference sediment. In the presence of contaminated sediment, test species avoided the contaminated sediment as early as 6, 6, and 24 h following exposure for
N. spinipes, P. solida and M. plumulosa, respectively. Avoidance was generally greater for sediments which elicited greater 10-d lethality. Each species was observed to have the ability to respond to chemical cues in the environment and inhabit sediment that provided the best opportunity for survival. Rapid screening methods to assess sediment toxicity could be developed using avoidance as an endpoint.
Acute and chronic toxicity associated with short, intermittent exposure to four field collected contaminated sediments was assessed for M. plumulosa and N. spinipes.
Increasing the duration of exposure caused a decrease in survival of M. plumulosa and
N. spinipes during 10-d bioassays. In addition, reproduction decreased following exposure to contaminated sediment. For M. plumulosa, reduced fecundity appeared to occur from exposure to contaminated sediment and reproductive effects occurred follwoing shorter exposures than needed to sense and avoid contaminant exposure.
Thus, while avoidance behaviours may prevent acute lethality, slow responses may not prevent sublethal effects.
This study indicates that sediment toxicity methods which utilise static continuous exposures may over-estimate the toxicity that would occur at a field location. However, by preventing organisms from avoiding unfavourable sediments, these methods provide a precautionary assessment of possible effects, which is usually the aim of most assessments frameworks.
Page | ii Table of contents
ABSTRACT...... i
CHAPTER 1 INTRODUCTION ...... 1 1.1 Common sediment contaminants ...... 2 1.1.1 Organic contaminants in the environment ...... 3 1.1.2 Metal contaminants in the environment ...... 5
1.2 Factors influencing the risk posed by metal contaminants ...... 8 1.2.1 Speciation and bioavailability ...... 8 1.2.2 Sediment redox chemistry ...... 9 1.2.2.2 Influence of redox chemistry on metal partitioning in sediments ...... 11 1.2.3 Effect of metals on organism health ...... 12 1.2.4 Dietary metal exposure and toxicity ...... 14 1.3 Sediment toxicity test methods ...... 15 1.4 Meiofauna ...... 18 1.4.1 Meiofaunal distribution, abundance and community structure...... 19 1.5 Copepods ...... 19 1.5.1 Ecology ...... 20 1.5.2 Development (life cycle) and reproduction of harpacticoid copepods ...... 22 1.5.3 Harpacticoid copepods and ecotoxicology ...... 26 1.6 Behavioural toxicology ...... 27 1.6.1 Avoidance as a response to contaminated sediment ...... 29 1.7 Pulsed exposures to contaminants ...... 31 1.7.1 Contaminant Pulses ...... 32 1.7.2 The effect of contaminant pulses on the environment...... 33 1.8 Aims and objectives ...... 35
CHAPTER 2 GENERAL METHODS ...... 39 2.1 Sediment Collection ...... 39 2.1.1 Site description ...... 39 2.1.2 Sediment sampling ...... 42 2.2 Analytical Method ...... 42 2.2.1 General cleaning ...... 42 2.2.2 Reagents...... 42 2.2.3 pH measurements ...... 43 2.2.4 Metal analyses by inductively coupled plasma atomic emission spectroscopy (ICP-AES) ...... 43 2.2.5 Water content ...... 44 2.3 Preparation of laboratory spiked sediments...... 44
Page | iii 2.3.1 Copper spiked sediment ...... 44 2.3.2 Diesel spiked sediment ...... 45 2.4 Collection, culturing and handling of test species ...... 46 2.4.1 Amphipods ...... 46 2.4.2 Copepods ...... 47 2.4.3 Estuarine Snails ...... 49 2.5 General data analysis ...... 50
CHAPTER 3 AN ASSESSMENT OF THREE HARPACTICOID COPEPOD SPECIES FOR USE IN ECOTOXICOLOGICAL TESTING...... 52 3.1 Introduction ...... 53 3.2 Methods ...... 55 3.2.1 Water and sediments ...... 55 3.2.2 Copepod collection and culturing ...... 56 3.2.3 Influence of food type and quantity on copepod culturing ...... 58 3.2.4 Toxicity test procedures ...... 59 3.2.5 Whole-sediment exposures using N. spinipes ...... 60 3.2.6 Statistical Analyses ...... 61 3.2.7 Analytical Methods ...... 61 3.3 Results and Discussion ...... 63 3.3.1 Copepod species and culturing ...... 63 3.3.2 Effect of food type on reproduction and juvenile development ...... 64 3.3.3 Sensitivity to dissolved copper ...... 72 3.3.4 Selection of copepod species for routine toxicity tests ...... 74 3.3.5 Sensitivity of N. spinipes to dissolved contaminants ...... 76 3.3.6 Use of N. spinipes for assessing sediment toxicity ...... 77 3.4 Conclusion ...... 78
CHAPTER 4 AVOIDANCE OF CONTAMINATED SEDIMENTS BY AN AMPHIPOD (MELITA PLUMULOSA), HARPACTICOID COPEPOD (NITOCRA SPINIPES) AND SNAIL (PHALLOMEDUSA SOLIDA) ...... 81 4.1 Introduction ...... 82 4.1.1 Materials and Methods ...... 85 4.1.2 Test media ...... 85 4.1.3 Test species ...... 87 4.2 Amphipod toxicity testing ...... 88 4.2.1 Contaminant avoidance experimental design ...... 89 4.2.2 Contaminant avoidance assays ...... 91 4.2.3 General chemistry ...... 92 4.2.4 Data Analyses ...... 93 4.3 Results and Discussion ...... 94 4.3.1 Grain size and avoidance behaviour in uncontaminated sediment ...... 94 4.3.2 Time to response: dispersal and avoidance of uncontaminated sediment ...... 97
Page | iv 4.3.3 Avoidance of contaminated sediment ...... 102 4.3.4 Validation of sediment avoidance of M. plumulosa ...... 104 4.3.5 Influence of hazard magnitude on avoidance behaviour ...... 107 4.4 Conclusion ...... 110
CHAPTER 5 SLOW AVOIDANCE RESPONSE TO CONTAMINATED SEDIMENTS ELICITS SUBLETHAL TOXICITY TO BENTHIC INVERTEBRATES ...... 111 5.1 Introduction ...... 112 5.2 Materials and Methods ...... 114 5.2.1 General Chemistry...... 114 5.2.2 Test media...... 115 5.2.3 Test species...... 116 5.2.4 Toxicity test procedures...... 116 5.2.5 Sublethal bioassays with pulsed contaminant exposure...... 118 5.2.6 Amphipod gender-exposure tests...... 122 5.2.7 Data Analyses...... 122 5.3 Results and Discussion ...... 123 5.3.2 10-day lethality from continuous or pulsed exposures to contaminated sediment ...... 124 5.3.3 Sublethal toxicity from pulsed exposure to contaminated sediment ...... 128 5.3.4 Gender of M. plumulosa influencing reproduction...... 130 5.4 Implications ...... 133
CHAPTER 6 CONCLUSIONS ...... 134 6.2 Further research ...... 143
REFERENCES...... 146
APPENDIX I...... 175
APPENDIX II...... 177
APPENDIX III...... 178
MANUSCRIPTS...... 179
Page | v List of Figures
Figure 1.1 Key processes in the cycling of copper in the environment (vanLoon and Duffy 2005)...... 6 Figure 1.2 Transport and transformation of contaminants in sediments (Eggleton and Thomas 2004) ...... 6 Figure 1.3 Sequential utilisation of electron acceptors by microbes and the product of the reduction reactions changes the abundance of chemical species in the sediment. The high redox potential (Eh) of oxic surface sediments rapidly decreases with depth, indicating increasingly anoxic (reducing) conditions. (modified from Burdige 1993; Park and Jaffé 1996)...... 11 Figure 1.4 Concentration-effect curve of a) synthetic organic chemicals, b) essential metals and metalloids and c) nonessential metals and metalloids (Chapman and Wang 2000)...... 13 Figure 1.5 Copepod habitats: a schematic representation of the primary habitat of each of the ten copepod orders. A. Platycopioida. B. Misophrioida. C. Harpacticoida. D. Calanoida. E. Mormonilloida. F. Cyclopoida. G. Monstrilloida. H. Poecilostomatoida. I. Siphonostomatoida. J. Gelyelloida. [A-C, benthic; D-G, planktonic; H-I, parasitic/associated; J, groundwater]. From Huys and Boxshall (1991)...... 21 Figure 2.1 Three life stages of the amphipod Melita plumulosa, (a) a full grown adult, (b) a young adult (1-2 months old) and (c) a juvenile (7-14 days old)...... 47 Figure 2.2 Adult specimens of the copepods species (a) Nitocra spinipes (b) Tisbe tenuimana and (c) Robertgurneya hopkinsi...... 49 Figure 2.3 The dorsal (left) and ventral (right) view of the common estuarine gastropod Phallomedusa solida...... 50 Figure 3.1 Photographs of (i) Nitocra spinipes (Boeck, 1864) (~24day life cycle), (ii) Tisbe tenuimana (Giesbrecht, 1902) (~28-day life cycle), and (iii) Robertgurneya hopkinsi (Lang, 1965) (~35-day life cycle). The scale bar represents a length of 250 µm...... 63 Figure 3.2 The reproductive output of N. spinipes, T. tenuimana and R. hopkinsi when fed a tri-algal diet (a-c, respectively; at concentrations of 0, 1, 3 and 10 ×106 cells per feed for control, low, medium and high treatments, respectively) and powered fish food (d-f, respectively; at concentrations of 0, 0.15, 0.5 and 1.5 mg per feed for control, low, medium and high treatments, respectively). The number of nauplii and copepodids were counted following 7 days of feeding at the respective diet and rate (mean ± standard error, n=4)...... 67 Figure 3.3 Dissolved oxygen (DO, % saturation, ) and ammonia (mg/L, ■) concentrations measured in the test vials of the feeding experiment for N. spinipes...... 69 Figure 3.4 The survival of a) N. spinipes, b) T. tenuimana and c) R. hopkinsi (mean ± SE, n= 4) when exposed to dissolved copper for ≥ 48 h. Nominal copper concentrations were 0 (○), 50 (■), 200 ( ), 400 (●), 600 (□) and 800 (▲) µg/L...... 73 Figure 4.1 Schematic diagrams of the experimental vessels used in a) M. plumulosa and P. solida, and b) N. spinipes avoidance experiments. The temporary barrier is shown i) in place (blue cross-hatching) and ii) removed from the avoidance vessels...... 90 Figure 4.2 The distribution of (a) Melita plumulosa, (b) Nitocra spinipes and (c) Phallomedusa solida when exposed for 48 h to sediments of differing silt composition (>90%-, 50%- and 10% silt, respectively, mean ± SE, n = 3; n = 4 for N. spinipes). The filled bars represent control sedimentand the patterned bars represent...... 96 Figure 4.3 The observed avoidance response of organisms (mean ± SE, n = 3; n = 4 for N. spinipes) when exposed to contaminated sediments: a) Melita plumulosa to Sediment 5;
Page | vi b) Nitocra spinipes to Sediment 3 and Sediment 1; and c) Phallomedusa solida to Sediment 7 and Sediment 6. The bar on the left of each pair represents the seeded side. The filled bars represent control sediments and the striped bars represent contaminated sediment...... 99 Figure 4.4 The observed avoidance response of Melita plumulosa following 48-h exposure to field contaminated and laboratory spiked sediments (mean ± SE, n = 3). * denotes a significant avoidance of contaminated sediment. 10-d whole sediment toxicity decreases from left to right...... 108 Figure 5.1 Visual representation of exposure scenarios used for a) amphipod and b) copepod bioassays. Acute toxicity (survival) is also shown for organisms exposed to Sediment 1 - green, yellow and red represent non-toxic, moderately toxic and toxic, respectively. See Appendix III for a summary of acute and chronic toxicity results...... 119 Figure 5.2 The effect of increasing periods of exposure to contaminated Sediment 1 on survival and reproduction of M. plumulosa. The shaded region represents duration of exposure required to elicit contaminant avoidance (from Chapter 4, Ward et al. 2013). The error bars represent standard error (n=4)...... 126 Figure 5.3 The effect of short (48-h) exposure to contaminated sediment on the survival and reproductive output of M. plumulosa exposed to uncontaminated sediment (control) and field-collected contaminated sediments (Sediments 1 and 4; mean ±SE, n=4)...... 129 Figure 5.4 The effect of short exposure to contaminated sediment on the reproductive output of N. spinipes (mean±SE ) exposed to Sediment 1 for 24-h (n=4) and Sediment 2 for 48-h (n=5). Total offspring was counted after a 10-d period in uncontaminated sediment following prior exposure to contaminated sediment...... 130 Figure 5.5 Reproductive success observed for four gender-exposure scenarios generated by exposing male and female M. plumulosa to contaminated Sediment 1 (mean ±SE, n=4). A reduction in reproductive success was found to occur in treatments where females had been exposed to contaminated sediment prior to mating ...... 132 Figure 6.1 The cumulative effect of short exposures to contaminated sediment causes a significant reduction in survival of M. plumulosa. Similarly, as the total exposure time increases, offspring production decreases. Significant chronic effects occur following shorter exposure times than those that cause significant mortality. This indicates contaminant avoidance may prevent acute toxicity from occurring but may result in chronic effects if clean habitat cannot be found within a sufficient period of time...... 141
Page | vii List of Tables
Table 1.1 Examples of metal ion speciation in surface waters (modified from Campbell et al. 2006) (EDTA = ethylenediaminetetra-acetic acid, HA = humic acid, FA = fulvic acid)...... 9 Table 1.2 Phyla typically found within meiofaunal assemblages (Nybakken 2001)...... 20 Table 2.1 Physico-chemical properties of the Bonnet Bay sediment (modified from Simpson et al. 2004; Simpson and King 2005)...... 40 Table 2.2 Sediment chemistry data for the test sediments used in this study ...... 41 Table 3.1 Sediment properties and effects to reproductive output of N. spinipes caused by exposure to contaminated sediments...... 79 Table 4.1 Properties of uncontaminated sediments...... 94 Table 4.2 Sediment chemistry data for the test sediments used in this study ...... 98 Table 5.1 Properties of control and contaminated sediments used in sediment pulse experiments .. 124
Page | viii Chapter 1 Introduction
The introduction of contaminants into the environment is a growing problem which poses a threat to both the health and function of ecosystems and human populations
(Li et al. 2012; Singare et al. 2012). Background concentrations of organic and metal contaminants are generally low, however elevated concentrations occur as a result of both natural and anthropogenic sources/activities (Li et al. 2012; Ye et al. 2012; Li et al.
2013).
Anthropogenic contaminants enter waterways from a variety of sources which may include chemical spills, pesticides and agrochemical application, urban and industrial runoff, aerial deposition, sewage flow or industrial discharge (Brent and Herricks 1998;
Reinert et al. 2002). Ultimately, these contaminants accumulate in the sediments of aquatic environments where they pose a risk to organisms and ecosystem health and function (Linnik and Zubenko 2000; Ho et al. 2002; Roulier et al. 2008; Gilroy et al.
2012; Singare et al. 2012).
Our understanding of the fate and behaviour of aquatic contaminants has grown in recent years. At the same time, a greater awareness of the complexity of benthic environments and the range of adverse effects that result from contaminant exposure has been gained. This awareness is being driven by the development of new, innovative methods used to identify and measure the potential ecological risk posed by sediment-bound contaminants.
One of the emerging trends in the area of sediment ecotoxicology is the recognition of a need to develop rapid sediment assessment tools which are of high ecological
Page | 1 significance. There is a general consensus in the literature that meeting this 'need' requires a more diverse range of standard test species and an increased use of chronic toxicity endpoints (Simpson et al. 2005; Greenstein et al. 2008; Kennedy et al. 2009;
Ward et al. 2011). As a result, new organisms that have the potential to be used as model test species are being sought to develop chronic toxicity test methods. This change in the focus of sediment ecotoxicology research will lead to an improvement in the robustness of current sediment quality guidelines and a more holistic approach to environmental risk assessment.
1.1 Common sediment contaminants
Organic contaminants may occur from natural sources such as forest fires, oil seeps, volcanic eruptions and plant exudates (Haritash and Kaushik 2009). However, industrialisation has led to a significant increase in the use and production of organic compounds such as polyaromatic hydrocarbons (PAHs), pesticides, polychlorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDEs) biphenyls (PCBs) (Warren et al.
2003). Many fully synthetic organic compounds exist and these chemicals have been collectively termed hydrophobic organic contaminants (HOCs). The wide spread use of such chemicals has led to the growing concern over the contamination of aquatic environments with HOCs.
Metals are persistent contaminants, as they are not created or destroyed by either anthropogenic or biological processes. The surrounding environment has a major influence on the chemical species of a metal contaminant (Campbell et al. 2006), which in turn mediates the impact of the contaminant to an ecosystem. Given the number of different species and phases in which metals can exist within the sediment, it is
Page | 2 important to understand sediment chemistry so that regulations can be implemented to better identify and monitor potential risks that metal contaminated sediments may pose to the aquatic ecosystems (Simpson et al. 2005). For the purpose of this study,
‘metal’ will be used to refer to metals and metalloids that have been associated with contamination and are potentially toxic to biota (Duffus 2002).
Metals are liberated into in the environment through natural mineral weathering processes, however elevated metal concentrations may arise as a result of anthropogenic sources (Birch et al. 1996; Campbell et al. 2006). Anthropogenic sources of metals are of concern because they can contribute 10-100 times more metals to an area than natural sources (Payne and Price 1999). The major natural sources of metal contaminants include volcanic eruptions, forest fires and mineral weathering products
(Garrett 2000; Dias and Edwards 2003). This is in contrast to the anthropogenic sources which are linked to urbanisation and industrial development (Birch et al. 1996;
Li et al. 2012) and may come from numerous sources including but not limited to metal processing and manufacture, storm water runoff, power generation and chemical spills.
1.1.1 Organic contaminants in the environment
The hydrophobic nature of many organic compounds means that they are poorly soluble in water. However, once these compounds enter an aquatic system, they have a tendency to bind with dissolved and particulate organic matter and settle through the water column and settle in the sediment (Karickhoff et al. 1979; Baker et al. 1986;
Lara and Ernst 1990; Golding et al. 2008). HOCs strongly bind to the organic carbon fraction of sediments, and can remain in the sediment long after being released to the
Page | 3 environment (Ghosh et al. 2003). Many of these compounds which end up in aquatic environments are persistent, toxic and have the potential to bioaccumulate in exposed organisms (Golding et al. 2008). Organic contaminants differ from metal contaminants as they generally undergo several degradation processes after being released into the natural environment. In general, these processes are volatilisation, hydrolysis, chemical oxidation and microbial degradation (Wild and Jones 1995). Hence the fate of organic contaminants that enter natural water bodies will vary depending on the resistance of the compound to degradation processes as well as environmental conditions. For example, the persistence of PAHs is known to increase with increasing molecular weight (Haritash and Kaushik 2009).
Once within the sediment, benthic organisms are at risk of exposure, with bioaccumulation resulting in exposure to organisms higher up the food chain (Burgess et al. 2003; Vane et al. 2009). The relationship between total concentrations of HOCs in sediments and ecotoxic effects is poorly understood due to the heterogeneity of sediment physicochemical properties which can influence the partitioning of HOCs in sediments (Ghosh et al. 2003). It is well-established that the partitioning of HOCs in sediments not only affects the accumulation of the contaminants in the sediment profile, but also influences the bioavailability and accumulation of organic contaminants in exposed organisms (Di Toro et al. 1991). The partitioning to the dissolved phase is controlled by the equilibrium between dissolved organic contaminants and those in the solid phase (sorbed onto sediment particles), with particulate organic contaminants considered to be less chemically and biologically available (Huang et al. 2003; Gilroy et al. 2012). However, disturbances to the
Page | 4 sediment, including bioturbation, can increase the rate of desorption of HOCs and result in an increase in their bioavailability (Gunnarsson et al. 1999; Goedkoop and
Peterson 2003).
1.1.2 Metal contaminants in the environment
Metals enter the hydrosphere via three main routes: (i) precipitation from the atmosphere; (ii) breakdown of organic and mineral matter; and (iii) discharged as solid or liquid waste (vanLoon and Duffy 2005). An example of this is shown in Figure 1.1, which depicts the environmental cycling of copper. Once metals have entered the hydrosphere, they react with suspended materials, flocculate and precipitate, accumulating in the bottom sediments of water bodies (Linnik and Zubenko 2000)
(Figure 1.2). The adsorption of metal contaminants is facilitated by the formation of insoluble complexes with organic matter, iron and manganese (oxy)hydroxides, carbonates, sulfides or colloidal surfaces associated with sediments (Salomons and
Förstner 1984; Calmano et al. 1993; U.S. EPA 2005). This results in the sediment becoming enriched with metal contaminants (Pempkowiak et al. 1999; Linnik and
Zubenko 2000; Simpson et al. 2005). As metals accumulate, the sediment can also desorb metals back into the water column (Park and Jaffé 1996; Maher et al. 1999).
This can occur if there is a change in the chemical or physical properties of the surrounding environment. For example, under oxic conditions Fe(III) is the dominant iron species in aquatic environment, complexes of which are insoluble. However, the reduction of Fe(III) results in the formation of soluble Fe(II) species, therefore releasing iron and the adsorbed metals as free metal ions as shown by Equation 1.1 (vanLoon and Duffy 2005).
Page | 5 Figure 1.1 Key processes in the cycling of copper in the environment (vanLoon and Duffy 2005).
Water Dissolved contaminant Dissociation / degradation Binding or complexation Partitioning or biotransformation
DOC bound contaminant Particulate contaminant
Resuspension
Contaminant Settling Mobilisation Surficial Pore water
sediments transport O2
Anoxic subsurface Burial sediments
Figure 1.2 Transport and transformation of contaminants in sediments (Eggleton and Thomas 2004)
Page | 6 2+ + FeOOH(s) + Me (aq) → FeOO-Me(s) + H (1.1) ↓ Reduction (in deeper sub-oxic/anoxic sediments)
2+ Fe(II)(aq) + Me (aq)
Where Me = metals such as Ag, Cd, Cu, Ni, Pb, Zn, and Hg
In anoxic sediment, sulfide has a significant influence on metal accumulation. A number of metal ions react with sediment-bound sulfide or residual sulfide in the pore water and precipitate as metal sulfides (Equation 1.2) which have a low solubility
(Simpson et al. 2000). This is reflected by the copper binding constants with sulfide, log
KCuS = 36, and hydroxyl groups, log KCuOH = 6.5 (Chen and Mayer 1999).
2+ 2+ FeS(s) + Me (d) → MeS(s) + Fe (aq) (1.2)
Uptake of contaminants in the aquatic environment can occur via a number of different pathways (Bianchini and Bowles 2002; Eggleton and Thomas 2004; Marsden and Rainbow 2004). These pathways include the uptake of metals via active transport across a cell membrane and passive diffusion for dissolved phase metals, and ingestion of food and sediment with associated particulate phase metals (Eggleton and Thomas
2004; Campbell et al. 2006). In the case of sediment dwelling (benthic) organisms, metal accumulation may be the result of the combined exposure to dissolved (pore water, overlying water) or particulate (sediment ingestion, food) metals (Simpson et al.
2005; Simpson and King 2005). Dietary exposure is also possible, and likely for organisms that ingest significant amounts of sediment during their feeding activities
(e.g. polychaete worm) (Landrum et al. 2013).
Page | 7 1.2 Factors influencing the risk posed by metal contaminants
Many of the factors that influence contaminant bioavailability and the risk they pose to benthic organisms are similar for organic and metal contaminants. However due to metals being the major contaminant class with many of the sediments used in this research (thesis), the factors influencing the risks posed by metals in the environment will be described in greater detail.
1.2.1 Speciation and bioavailability
In natural ecosystems, the chemical forms in which a metal can occur is strongly influenced by the chemistry of the surrounding environment, and may be referred to as ‘speciation’. Speciation is defined as the specific form of an element based on isotopic composition, oxidation state and/or molecular structure (Campbell et al.
2006). Examples of some of the metal species found in surface waters can be seen in
Table 1.1. Metal speciation will influence the biological availability (bioavailability) of a particular contaminant and therefore influence the effect of the contaminant on the ecosystem. In other words, the toxicity of a metal is largely dependent on the bioavailability of the contaminant to an organism (Maher et al. 1999; Di Toro et al.
2005; Simpson and King 2005).
According to Simpson (2005), bioavailability of metals in sediment is controlled by the partitioning of metals between dissolved phases (in sediment pore waters and overlying water), and solid phase particulate matter associated with sediments
(complexed iron and manganese (oxy)hydroxides, particulate organic matter and
Page | 8 Table 1.1 Examples of metal ion speciation in surface waters (modified from Campbell et al. 2006) (EDTA = ethylenediaminetetra-acetic acid, HA = humic acid, FA = fulvic acid).
Species Examples
3+ Free metal ions Al (H2O)6 2+ Cu (H2O)6
2+ + - Hydroxo-complexes AlOH , Al(OH)2 , Al(OH)4 2+ + - FeOH , Fe(OH)2 , Fe(OH)4 0 + Cu(OH)2 , CuOH
2+ + Simple inorganic AlF , AlF2 + 0 - 0 complexes CdCl , CdCl2 , CdCl3 CuCO3 , 0 CuSO4 , Cu-EDTA, Cu-colloid, Cu-HA/FA
sulfide). In addition, organism physiology and behaviour will determine the extent of adverse effects of metals to the organism (Chen and Mayer 1999; Simpson 2005).
1.2.2 Sediment redox chemistry
The ability of metal contaminants to bind to sediment particles is influenced by the redox state of the sediment (Calmano et al. 1993). Several redox transition zones have been identified in the sediment profiles of a number of sedimentary facies (Figure 1.3).
Vertical stratification of the sediment results from microbial degradation of organic matter in addition to the limited ability of oxygen to penetrate the sediment (Alongi et al. 1996).
The penetration of dissolved oxygen from the overlying water into the surface sediments creates an oxic environment (Clark et al. 1998; Maher et al. 1999). The oxidised sediment layer does not extend deep into the profile, often having a thickness of only a few millimetres (Clark et al. 1998). However, oxygen penetration may be
Page | 9 increased due to sediment mixing from bioturbation and bioirrigation, boating, high energy water flow and other disturbance events (Alongi et al. 1996; Clark et al. 1998;
Maher et al. 1999).
Below the oxic zone, anoxic conditions prevail. However this zone can be subdivided into two categories: the suboxic zone (upper reduction zone) at the limit of oxygen penetration, and the deeper anoxic zone (lower reduction zone) (Clark et al. 1998).
Both the suboxic and anoxic redox zones are characterised by the occurrence of decomposing organic matter and sulfate reducing bacteria leading to the production of sulfides (Clark et al. 1998). However, the chemical composition of the suboxic zone differs from the deeper anoxic sediments as it characteristically has high concentrations of dissolved iron and manganese and low oxygen and sulfide concentrations as shown in Figure 1.3 (Alongi et al. 1996).
Micro-organisms in sediments control sediment redox potential due to the chemical reactions facilitated by microbial degradation of organic matter (Park and Jaffé 1996).
Electron acceptors used by sediment microbes include oxygen, nitrate, manganese and iron oxides, sulfate and carbon dioxide (Park and Jaffé 1996), listed in order of preferential sequestration (Figure 1.3). Oxygen penetration is generally restricted to the surface layer of sediment (Clark et al. 1998), and is dependent on sediment texture and the presence of biota in the sediment, which rapidly make use of available oxygen for respiration. In the suboxic layer, the reduction of iron and manganese
(oxy)hydroxides (e.g. FeOOH, MnO2) to form the dissolved Fe(II) and Mn(II) species,
Page | 10 Sedimentation Concentration
O2 H2O + 900 Aerobic bacteria O2 CO - Oxic <2cm 2 NO3 and fungi
- NO3 N2
O 2+ Denitrify ing Mn R CO2 bacteria G 4+ 2+
A Mn Mn N Manganese CO I reducing bacteria 2 Sub-oxic C 2-8 cm Fe3+ Fe2+ Fe2+ C Iron reducing 2- CO Potential Redox SO A bacteria 2 4 R 2- 2- B SO4 S
O Sulfate reducing CO2 N bacteria
S2- CO2 CH4 - 300 Anoxic Methanogenic CO2 >8 cm bacteria
Burial
Figure 1.3 Sequential utilisation of electron acceptors by microbes and the product of the reduction reactions changes the abundance of chemical species in the sediment. The high redox potential (Eh) of oxic surface sediments rapidly decreases with depth, indicating increasingly anoxic (reducing) conditions. (modified from Burdige 1993; Park and Jaffé 1996). dominates many chemical processes (Burdige 1993; Park and Jaffé 1996). The deeper anoxic sediments contain sulfate-reducing bacteria which result in a build-up of sulfide in this region. These processes change the abundance and distribution of the chemical species of metal contaminants within the sediment.
1.2.2.2 Influence of redox chemistry on metal partitioning in sediments
Bioavailability of metals in the aquatic environment is dependent on the partitioning behaviour of a metal contaminant and the binding strength of the metal to sediment particles (Di Toro 1990; Simpson 2005). Therefore, dissolved metals are typically considered to be more bioavailable than sediment bound contaminants which may only become available if ingested (Eggleton and Thomas 2004).
Page | 11 As stated, the speciation of metals in sediments may be strongly influenced by the sediment redox potential. In oxidised sediment, iron and manganese (oxy)hydroxides
(FeOOH, MnOOH) and particulate organic matter (POC) are prevalent binding phases for metal contaminants (Equations 1.3 and 1.4) (Saulnier and Mucci 2000; Fan et al.
2002).
2+ + + POC(H)(s) + Me (d) → POC-Me (s) (+H ) (1.3)
2+ + + FeOOH(s) + Me (d) → FeOO-Me (s) + H (1.4)
However, in anoxic sediment, sulfides are often the dominant metal-binding phase influencing the bioavailability of metals (Di Toro et al. 1990; Cooper and Morse 1998;
Eriksson and Sundelin 2002).
1.2.3 Effect of metals on organism health
The development of regulatory systems within an organism allows the concentrations of these metals to be maintained within narrow ranges (homeostasis). Figure 1.4 shows a generalised concentration-effect curve of both essential and nonessential elements. In the case of the essential element, it is evident that there is an optimum concentration range which is maintained by homeostasis to allow the organism to function. Deficiency occurs when internal body concentrations fall below the required optimum which may result in symptoms that mirror that of toxicity (body concentration is too high). In either case, the organisms health deteriorates and death can result. Homeostatic responses can be responsible for the detoxification of non- essential metals (Pb or Hg) and removal of essential metals before bioaccumulation leads to toxicity. This has been observed to occur with other substances.
Page | 12
Figure 1.4 Concentration-effect curve of a) synthetic organic chemicals, b) essential metals and metalloids and c) nonessential metals and metalloids (Chapman and Wang 2000).
For example, some organisms may convert sulfide (which can be toxic) to thiosulfate, sulfite and sulfate as a detoxifying mechanism (Taylor et al. 1999).
Organisms have evolved with metals being naturally present in the environment.
Because of this, there are many environmental repercussions that must be considered including essentiality, bioaccumulation and tolerance. Many metals are considered essential elements and are required by organisms to sustain life, including metals such as Cu, Co, Fe, Mn, Mo, Ni, Se, and Zn. Therefore it is essential that these be present in the ambient environment (Scheinberg 1991; Campbell et al. 2006). The bioaccumulation of these essential metals is a natural process required by living organisms for metabolism and growth (Campbell et al. 2006).
In contrast to metals, organic contaminants often have negligible background concentrations. As a result, HOCs are generally highly toxic as organisms have not evolved with detoxification mechanisms and a beneficial concentration range does not exist (Schwarzenbach et al. 1993; van der Hoeven and Gerritsen 1997; Schulz and Liess
2000; Naddy and Klaine 2001).
Page | 13 1.2.4 Dietary metal exposure and toxicity
Recent research has highlighted the importance of considering an organisms dietary exposure to metals when evaluating metal toxicity thresholds for benthic organisms
(Absil et al. 1996; Simpson et al. 2005) and for higher order aquatic species such as fish
(Woodward et al. 1995; Berntssen et al. 1999; Shaw and Handy 2006). For example, it has been demonstrated that fish which have been fed metal-contaminated benthic invertebrates may display symptoms of metal toxicity, such as reduced growth and survival (Woodward et al. 1995; Morris et al. 2003). This suggests that the dietary uptake of metals may be contributing to the observed chronic and/or acute toxicity.
For benthic organisms, contaminant exposure can occur via a number of pathways including overlying water, pore water, sediment ingestion and food sources (e.g. algae)
(Simpson and King 2005). It is essential that the sensitivity of benthic organisms to metal contamination be assessed to allow for the development of more robust sediment quality guidelines. An investigation into exposure pathways of benthic organisms indicates that sediment ingestion by benthic organisms can lead to toxicity
(Simpson and King 2005). Eriksson and Sundelin (2002) concluded that for the amphipod Monoporeia affinis, the contribution of metal exposure from metal accumulation via sediment and food ingestion was greater than that of pore water concentrations. These results were supported by studies conducted on the Australian amphipod species Melita plumulosa, an epibenthic deposit feeing species. Simpson and King (2005) and King et al. (2006b) demonstrated that causality in toxicity tests using this species could be explained by dietary uptake of sediment-bound copper.
These tests were conducted on oxidised sediments, where copper was bound to iron
Page | 14 and organic matter phases (POC-Cu+, FeOOCu+), rather than as copper sulfide (CuS) commonly associated with anoxic sediment.
There is still debate over the bioavailability of metal sulfides to sediment ingesting organisms, as the bioavailability of sediment-bound copper is related to both the digestive physiology of organisms and the geochemistry of the sediment (Chen and
Mayer 1999). Evidence indicates a reduction in the bioavailability of sulfide-bound copper due to reduced solubility in organism gut fluid extract taken from the gastrointestinal tract of a polychaete species and two species of holothuroid (Chen and
Mayer 1999). Similarly, a reduction in copper toxicity was observed in sediment containing elevated amounts of AVS (Lee and Lee 2005). This is in contrast to studies of bivalves which demonstrate that metals accumulate within soft tissues regardless of the concentration of AVS in the sediment (Griscom et al. 2000; Lee et al. 2000). Further investigation into the bioavailability of metal sulfides (associated with anoxic sediment) is required to determine if the assumptions of the AVS-SEM model are acceptable for organisms with sediment ingestion exposure routes (Simpson et al.
2012).
1.3 Sediment toxicity test methods
Sediment bioassays are commonly used as a tool to assess the biological effects that may result from exposure to contaminated sediment (Simpson et al. 2005). Unlike chemical analyses, sediment bioassays can provide additional insight into potential population and ecosystem changes resulting from contaminant exposure which can be useful for the management of coastal marine environments (Roper and Hickey 1994).
Furthermore, the use of sediment bioassays for toxicity testing provides a measure of
Page | 15 the bioavailability of contaminants by exposing test organisms directly to a sample of sediment. In this way, toxicity testing is seen as a powerful indicator of adverse ecological effects.
The methods used for toxicity testing can be grouped into two broad categories depending on the duration of exposure and the endpoint chosen as the measure of toxicity. Acute toxicity tests are shorter in duration and encompass a short period of the test organism’s life cycle and typically result in rapid and severe effects (such as lethality). Alternatively, a bioassay may be used to observe non-lethal effects due to contaminant exposure, including embryo development, growth, moulting, and reproduction (Hickie et al. 1995; Grimalt et al. 2001). These are called sublethal tests, and many assess potentially chronic impacts. Typically, chronic test procedures are often longer (often exceeding 20 days) and more labour intensive (U.S. Environmental
Protection Authority 2001; Egeler et al. 2010). For sediment bioassays to be adopted into routine risk assessment and incorporated into regulatory guidelines, they ideally need to be rapid, easy to conduct and inexpensive.
For a biological species to be suitable for toxicity testing, the species should ideally be sensitive to contaminants of concern, have a year-round availability from natural populations or laboratory cultures, be of ecological significance and a have a broad geographical distribution (ASTM 2003). A wide variety of test species have been identified as suitable for use in toxicity testing. For a selection of these, standard test methods have been developed and endorsed by regulatory bodies (refer to USEPA
1994; ASTM 1999; USEPA 2001; ASTM 2013). Currently many of these standard
Page | 16 methods focus on acute lethality and do not assess the impact of contamination on other aspects of an organism’s life cycle (chronic endpoints).
While acute toxicity test methods may have a time advantage over chronic toxicity tests, the use of rapid acute bioassays may not adequately assess the impacts of environmental contamination (Birch 2000; Grimalt et al. 2001; Finkelstein and Kern
2005; Scarlett et al. 2007; Greenstein et al. 2008). However, chronic endpoints are generally more sensitive, offer a better prediction of toxicity and can be utilised to assess impacts on community and population dynamics (Davison et al. 1997; Birch
2000; Diamond et al. 2006; Kennedy et al. 2009). Despite this, many researchers and regulatory guidelines rely on standard methods which confine test organisms enforcing a continuous static exposure aimed at assessing acute effects, usually mortality (ASTM
2003; Simpson et al. 2005). The relevance of these common ecotoxicology methods is widely debated in the literature, with chronic test methods often preferred over acute toxicity testing (Cairns 1992; Lopes et al. 2004). Chronic toxicity tests are considered to offer greater ecological relevance, protection at the population level, increased sensitivity, better prediction of toxicity, and the ability to model population effects
(Kennedy et al. 2009).
Recently there has been a push to incorporate chronic toxicity testing for regulatory purposes. This has resulted in the development of standard methods which employ chronic endpoints for sediment quality assessment (ISO 2010; OECD 2010). However, the increased test duration and labour intensive methods have prevented chronic test methods from being adopted for regulatory purposes (Abel 1980; Simpson et al. 2005).
As a result, there is a need to identify new species which are potential candidates for
Page | 17 toxicity testing and also have attributes that make chronic toxicity assessment more achievable. This requires development of rapid methods for assessing chronic effect resulting from contaminant exposure. Smaller organisms that frequently reside near the bottom of the food chain and have short life cycles (such as meiofauna) provide excellent opportunities for the development of fast, sensitive and ecologically relevant chronic toxicity tests. For this reason, the following sections provide a substantial review and discussion of meiofuana, with a focus on copepods, and their use for environmental effects assessment.
1.4 Meiofauna
Meiobenthic fauna or meiofauna is a term used to describe the range of organisms that live within the sediment interstitial waters or adhered to the surface of individual sediment grains (Nybakken 2001; Somerfield et al. 2005). These organisms were not observed until the 20th century but have since been found to occur in intertidal and subtidal environments of both marine and fresh waterways (Nybakken 2001). The classification of meiofauna is based on size, commonly accepted to comprise of organisms that pass through a 500 µm sieve but are retained on a 63 µm mesh
(Somerfield et al. 2005).
Meifauna play a significant role in the marine environment as a link between microfauna (algae and bacteria) and macrofauna (reviewed in Coull 1999). They are known to be an important food source for higher trophic levels such as polychaetes, prawns, birds and juvenile fish (Beier et al. 2004). In addition, meiofauna stimulate microbial activity thus increasing carbon mineralisation and/or the consumption of detritus by deposit feeding invertebrates (Schmid-Araya and Schmid 2000).
Page | 18 1.4.1 Meiofaunal distribution, abundance and community structure
It is estimated that there is an average of 106 meiofaunal organisms per square meter of sediment which would equate to a biomass of approximately 1-2 g/m2 (Coull and
Bell 1979; Giere 1993). The abundance of meiofauna does vary spatially with the greatest abundance found to occur in intertidal environments, declining as water depth increases (Nybakken 2001). Within these environments, it is generally observed that meiofauna abundance is greatest in organically enriched muds rather than clean sand (Coull 1999). Meiofauna seem to be restricted to the oxic fraction of the sediment profile where more than 95% of these organisms are found to occur (Coull and Bell 1979; Kovatch et al. 1999; Hagopian-Schlekat et al. 2001).
Meiofaunal communities may be comprised of a number of organisms coming from a variety of invertebrate phyla, however, those phyla which are typically represented by large bodied or sessile organisms are poorly represented in this class of aquatic organisms (Nybakken 2001). Meiofauna communities typically consist of nematodes, crustaceans, foraminifera, turbellarians, rotifers, gastrotrichs and annelids (Kovatch et al. 1999) but may contain species representative of any number of the phyla listed in
Table 1.2. While there is a diverse range of organisms that are found to occur as meiofauna, it has been commonly observed that at any one particular site there will be a dominance of only a few species (Coull 1999).
1.5 Copepods
Copepods represent one of the many under-represented meiofauna considered for bioassays. Copepods were so named based on their broad paddle like swimming legs,
Page | 19 Table 1.2 Phyla typically found within meiofaunal assemblages (Nybakken 2001).
Phylum of meiofauna
Ciliophora Platyhelminthes Nematoda Gastrotricha
Kinorhyncha Tardigrada Rotifera Priapulida
Loricifera Entoprocta Annalida Crustacea
Echinodermata Cnidaria Bryozoa Chordata
with the name coming from the Greek words ‘hope’ and ‘podos’ which means ‘oar- footed’ (Huys and Boxshall 1991; Pechenik 1996). Copepods are a group of organisms belonging to the phylum crustacea. They are generally small, being less than 2 mm in length (Pechenik 1996), although some parasitic species may grow larger, up to 250 mm in length (Huys and Boxshall 1991). These organisms are distributed widely, inhabiting a range of aquatic environments including marine waters, freshwater lakes and ponds and terrestrial environments (Figure 1.5) from water depths of 1 km in the
Philippine Trench to an altitude of more than 5 km in the Himalayan Mountains (Wells
1988; Huys and Boxshall 1991). Copepods also display a number life-history strategies with benthic, pelagic and parasitic species identified (Suárez-Morales et al. 2006).
1.5.1 Ecology
It is estimated that more than 12,000 identified copepod species exist (Ho 2001). Ten phylogenetic orders have been identified in the subclass of copepoda (Ho 2001). It is estimated that almost 99% of copepod species belong to the orders of Calanoida,
Page | 20
Figure 1.5 Copepod habitats: a schematic representation of the primary habitat of each of the ten copepod orders. A. Platycopioida. B. Misophrioida. C. Harpacticoida. D. Calanoida. E. Mormonilloida. F. Cyclopoida. G. Monstrilloida. H. Poecilostomatoida. I. Siphonostomatoida. J. Gelyelloida. [A-C, benthic; D-G, planktonic; H-I, parasitic/associated; J, groundwater]. From Huys and Boxshall (1991).
Cyclopoida, Harpacticoida, Poecilostomatoida and Siphonostomatoida, each containing more than 1,200 species (Ho 2001). Free-living copepods primarily exist in one of three orders; calanoida (mostly planktonic), harpacticoida (mostly benthic) and cyclopoida (includes both planktonic and benthic species) (Barnes 1963; Green 1968).
Copepods are recognised as the most abundant metazoans in the sea comprising 70% of the oceans biomass (Raisuddin et al. 2007). The abundance of copepods is so high that even their faecal pellets are of ecological importance providing an energy source for detritus feeders and possibly influencing nutrient cycling and sedimentation rates
(Huys and Boxshall 1991). Planktonic copepod species form a major component
Page | 21 oceanic zooplankton communities and are an important part of aquatic food chains
(Pechenik 1996; Støttrup 2003), particularly as a food source to juvenile fish (Hicks and
Coull 1983). In open marine waters, calanoid copepods dominate zooplankton communities and are the basis of the food chain for most fish larvae and planktivorous fish, forming the link between phytoplankton and higher order consumers
(Raisuddin et al. 2007). Species from the order Harpacticoida dominate copepod species in marine sediments (Suárez-Morales et al. 2006), often being the second most abundant taxa in meiobenthic communities, behind Nematodes (Pechenik 1996;
Suárez-Morales et al. 2006). Harpacticoid copepods also have a significant ecological role, again as a link between primary producers and higher order consumers, but in particular as a source of food to juvenile fish (Hicks and Coull 1983; Raisuddin et al.
2007), with some species, such as herring, continuing to feed on copepods during adulthood (Huys and Boxshall 1991).
1.5.2 Development (life cycle) and reproduction of harpacticoid copepods
In the order Harpacticoida, fertilised eggs are usually carried in a single egg sack, but may also be carried as paired egg sacks or carried for a short period of time then attached to the sediment (Dahms and Qian 2004). Juvenile copepods hatch from the eggs as nauplius larvae, a characteristic that is typical of several groups of crustacea
(Pechenik 1996). Although some variation may exist, the naupliar larvae are often sub- circular in shape, with a broad unsegmented body, for example the nauplii of the species Stenhelia palustris (Green 1968). The naupliar larvae are equipped with a single median eye and three pairs of appendages; antennules (uniramous), antennae
(biramous) and mandibles (biramous) (Barnes 1963). Development of the nauplii
Page | 22 generally results in a change in body form (due to segmentation and growth) and the acquisition of additional limbs (Barnes 1963; Dahms and Qian 2004). Harpacticoid copepods develop through six nauplius stages (N1-N6) followed by five copepodid stages (I-V), before taking the form of the mature adult as the sixth copepodid stage.
The transition between each nauplii/copepodid stage is generally marked by a single molt (Johnson and Olson 1948), therefore maturity is reached after a total of 11 molts
(Hicks and Coull 1983). At the end of the N6 stage (6th naupliar stage), the body form of the naulpii changes dramatically as the juvenile metamorphoses into a copepodid
(Dahms and Qian 2004). Dahms and Qian (2004) report the presence of paragnaths, a median ventral structure and swimming legs as being structures developed in the copepodid stages.
Copepod mating occurs in four stages; grasping of the female, courtship, copula and post copulatory mate guarding (Dürbaum 1995). Once the male detects a suitable female, reproduction begins with the clasping of the female by the male (Kern et al.
1984). The mating copepods may remain in this position for hours to a number of days
(Hicks and Coull 1983). Precopulatory mate guarding (or precocious coupling) between mature males and immature females is well documented for harpacticoid copepods, with adult male copepods have been observed mating with females as young as copepodid stage I (Hicks and Coull 1983). Palmer and Coull (1980) observed mature male specimens from the copepod species Microarthridion littorale always clasped stage V copepodid females, attaching the spermatophore only after her final moult.
Similarly, adult male copepods from the species Zausodes arenicolus were found to frequently clasp juvenile females from all copepodid stages (I-V), but not adult females
Page | 23 (Kern et al. 1984). In this case, Kern et al (1984) suggest that the clasping of younger females by mature males is the result of a reduction in the availability of mates caused by seasonal variation in copepod abundance. Precocious coupling has also been discovered to occur in the freshwater harpacticoid species Harpacticella inopinata by
Evstigneeva (1993). It was observed that H. inopinata males would clasp copepodid females from copepodid stages III, IV and V, but most commonly with the latter. This behaviour is thought to ensure a suitable mate as soon as the female matures, with copulation and fertilisation only occurring once the female matures not during the precocious coupling (Hicks and Coull 1983).
Dürbaum (1995) provides one of the first detailed accounts the complex courtship process for four Tisbe species and Paramphiascella fulvofasciata. Based on the results presented by Dürbaum (1995), courtship generally involves the male (grasping the female) to produce a strong swimming current, effectively lifting the female from the substrata. This allows him to position himself underneath her, facing opposite directions with ventral sides facing. In this position, the male strokes the ventral side of the female with his peraeopods resulting in her becoming motionless. In Tisbe holothuriae, the courtship behaviour may last for 15 minutes (Dürbaum 1995).
Sperm transfer occurs when the male deposits a spermatophore in the genital field of the mature female copepod (Dahms and Qian 2004). This is often very quick, often only taking a few seconds (Dürbaum 1995). For T. holothuridae, the neck of the spermatophore is attached to the genital field of the female with a glue like substance
(Dürbaum 1995). When deposited, the spermatozoa discharges and is stored in the seminal recipticles of the female (Dürbaum 1995; Dahms and Qian 2004) and fertilises
Page | 24 the eggs as they pass through the genital antrum to the external egg sack (Hicks and
Coull 1983). Following the attachment of the spermatophore, the male eases his grip on the female and remains inactive (Dürbaum 1995).
Following the deposition of the spermatophore, the male remains clasped to the female with his maxillipeds and second maxillae in a position that resemble the courtship behaviour (Dürbaum 1995). While the male is engaged in this position, no other male is able to clasp the female and attach another spermatophore, therefore securing paternity by preventing sperm competition (Dürbaum 1995). As the spermatophore discharges, the male begins to slide down the female’s furcal setae to which he is clasped to, finally detaching from the female by opening his maxillipeds and making a few vigorous swimming strokes (Dürbaum 1995). For the copepod species investigated by Dürbaum (1995), it was found that postcopulatory mate guarding may last up to 16 hours. It has been observed that female copepods only mate once during their lifetime. Burton (1985) found that electrophoretically-detected genetic markers clearly identified that multiple matings of a female were rare or absent in Tigriopus californicus, with only 1 out of 729 offspring scored showing the possibility of resulting from a multiple mated female. Dürbaum (1995) also show that multiple matings are rare with 3 out of 5 species investigated showing no evidence of successive matings. Only 5 females were observed to have more than 1 (up to 3) spermatophores attached to the genital field. However, it is evident that several broods can be produced from a single mating event. This was shown by Palmer and
Coull (1980) who observed female specimens of the species Microarthridion littorale producing up to three egg clutches following copulation. Female specimens of the
Page | 25 harpacticoid species Tisbe furcata were also observed to mate only once, with this mating sufficient to result in the production up to 12 fertile egg sacks (Johnson and
Olson 1948). Males were capable of more than one mating.
1.5.3 Harpacticoid copepods and ecotoxicology
Meiofauna, including copepods, are generally restricted to living in the oxic fraction of the sediment profile (Coull and Bell 1979; Kovatch et al. 1999; Hagopian-Schlekat et al.
2001). This puts these organisms at particular risk to contaminant exposure as metal and organic contaminants are most bioavailable between the sediment-water interface and the anoxic zone (Coull and Bell 1979; Kovatch et al. 1999; Hagopian-
Schlekat et al. 2001). Because of the importance of these organisms as a food source, metal exposure could potentially result in the introduction of metals into sediment based food webs or if toxicity to copepods occurs, a reduction in food quality and quantity. This may have adverse effects on the health and function of the contaminated ecosystem (Coull and Bell 1979; Kovatch et al. 1999; Hagopian-Schlekat et al. 2001). Due to their abundance, distribution and robustness, harpactioid copepods and nematodes have shown the most potential for use in the assessment environmental disturbances, whether anthropogenic or naturally occurring
(Schratzberger et al. 2000).
Harpacticoid copepods have been used in ecotoxicology for a number of years in
America, Europe and the United Kingdom (Raisuddin et al. 2007). The use of harpacticoid copepods for the purpose of sediment quality research began in the
1970’s (D'Agostino and Finney 1974; Linden et al. 1979), with Bengtsson (1978) making the first attempt at a standardised method for acute toxicity testing. Since then,
Page | 26 interest in the potential of copepods to serve as a model organism for marine ecotoxicology has resulted in a huge increase in the number of publications focusing on copepods. The suitability of these organisms for use in ecotoxicological testing is summarised by Brown et al. (2005) as being due to the small size of the organisms, fast development times, ease of culturing in the laboratory and the recognisable larval stages which permit the comparison of sensitivity of different life history stages to contaminants.
Currently, standard methods only exist for the harpacticoid species Amphiascus tenuiremis exposed to dissolved contaminants in water-only assays (ASTM 2004). This species has also been used in sediment bioassays and has been shown to be sensitive to a range of contaminants and suitable for use in laboratory testing (Chandler and
Green 1996; Hagopian-Schlekat et al. 2001). Novel techniques for assessing sediment quality have been developed using meiofauna in recent years. These behavioural test methods utilise the migration of meiofaunal communities (including harpacticoid copepod species) as they respond and move as a result of contaminant exposure
(Chandler et al. 1997; Ho et al. 2013). Further investigation into the development of new ecotoxicological testing methods which take advantage of the short life cycle of harpacticoid copepods as well as employ chronic and behavioural endpoints greatly improve the assessment of contaminated sediments.
1.6 Behavioural toxicology
Aquatic organisms are being exposed to increasing levels of contaminants as human pressures on the environment increase. Although the effects of contaminant exposure vary greatly, it is well known that environmental contaminants impair the healthy
Page | 27 function of aquatic organisms and ecosystems. Impacts on organism survival have been well studied for a range of contaminants and test species (McCahon and Pascoe
1991; USEPA 1994; ASTM 1999; Schulz and Liess 2000). Similarly, effects on growth and reproduction have also been extensively studied (Kallander et al. 1997; USEPA
2001; Boxall et al. 2002). Other chronic effects have been documented, with contaminant exposure found to influence other crucial behaviours such as an organism’s ability to effectively forage for food, avoid predation, find suitable habitat, compete for resources and tolerate a number of biotic and abiotic interactions within the surrounding environment (Atchison et al. 1996; Ho et al. 2002; Singare et al. 2012).
Contaminant-induced behavioural changes can be linked to biochemical and physiological effects in exposed organisms, which can ultimately affect population dynamics (Amiard-Triquet 2009). Therefore, any significant changes to behavioural traits of an organism that are critical for survival (e.g. foraging behaviour, social interaction, courtship and locomotion) may not only pose a threat to the survival of the organism, but also of the entire impacted population (Weber 1997; Roast et al.
2000). Therefore, any significant changes to behavioural traits of an organism that are critical for survival (e.g. foraging behaviour, social interaction and locomotion) may not only pose a threat to the survival of the organism, but also of the entire impacted population (Weber 1997; Roast et al. 2000).
Behavioural traits have been shown to be useful and sensitive toxicity test endpoints
(Lefcort et al. 2004; Lopes et al. 2004). Consequently, behavioural endpoints are gaining increasing attention for their ecological relevance and insights into potentially major environmental impacts, indicated by contaminant-induced changes to organism
Page | 28 behaviour (Amiard-Triquet 2009). Contaminant exposure has been observed to cause changes to the behaviour of aquatic organisms. This includes changes to predator- prey relationships (Riddell et al. 2005), swimming/mobility (Roast et al. 2000) and contaminant avoidance (Kravitz et al. 1999; Lopes et al. 2004).
1.6.1 Avoidance as a response to contaminated sediment
Contaminant concentrations in sediments are rarely homogenous. Spatial variation in the concentration of toxicants in surface sediments has highlighted the ability of a number of organisms to detect and actively avoid patches of contaminated sediment
(Rakocinski et al. 1997; Chariton et al. 2010). This means that as pollution levels increase in a particular location, there is the potential for the organisms inhabiting the area to migrate away from the contaminated site. Hence avoidance behaviour is a significant factor in determining the extent of exposure (Weber 1997; Lefcort et al.
2004), the magnitude of the hazard and the overall risk the sediment poses to ecosystem health.
Avoidance of contaminated sediment is a measurable response which often occurs rapidly following exposure to unfavourable conditions. Previous studies have indicated that the use of contaminant avoidance as an ecotoxicological endpoint has the advantage of being more sensitive than commonly used lethal and even sublethal endpoints (Burton et al. 1996; Lefcort et al. 2004; Lopes et al. 2004). Furthermore, behavioural endpoints such as avoidance can be used as an effective method of screening potentially toxic sediment samples, providing an indication of contaminant levels which may result in acute toxicity given sufficient exposure. In this way, sediment avoidance by benthic invertebrates can be used as an early warning sign of
Page | 29 impending toxic effects (Hellou 2011). Avoidance behaviour has been investigated for a diverse range of aquatic organisms and has proved to be a useful endpoint for sediment quality assessment (Roper et al. 1995; Kravitz et al. 1999; Exley 2000; Lefcort et al. 2004; Lopes et al. 2004).
Contaminant avoidance is likely to be a defence mechanism against the effect of poor environmental conditions, eliminating the threat to survival caused by the presence of toxicants (Lefcort et al. 2004). It is also possible that contaminant avoidance is an adaptive behaviour, acquired through previous exposure to unfavourable habitat
(Eriksson et al. 2006). Lefcort et al.(2004) observed that aquatic snails collected from contaminated environments could avoid contaminated sediment, while those from reference sites could not, inferring a potential genetic influence on avoidance behaviour. Regardless of the mechanisms triggering the avoidance response observed, it is important to note that the relocation of a population of aquatic organisms may ensure their survival at a new site. On an ecological scale, the disappearance of the species from the previous location is equivalent to the loss of the entire population
(Roast et al. 2000; Lopes et al. 2004). Consequently, large areas of sediments that are avoided due to contamination may be classified as high risk although short-term effects to many species used for toxicity tests may not be observed.
Static toxicity tests rarely reflect the dynamic and heterogeneous natural environment
(Lefcort et al. 2004) and therefore do not always adequately mimic the natural conditions in which organisms are exposed to contaminants, hence such methods may lack ecological significance (Cairns 1992; Burton et al. 2000). In addition, the long test durations often associated with chronic tests do not adequately represent realistic
Page | 30 exposure conditions in the natural environment given the ability of organisms to actively avoid contaminated sediment. A better understanding of the avoidance response displayed by mobile benthic organisms will shed light on the response of benthic invertebrates to contaminated sediments in the environment. This will allow for the development of environmentally relevant methods for assessing sediment toxicity caused by contamination in heterogeneous field settings.
1.7 Pulsed exposures to contaminants
As early as the 1980s, it was recognised that aquatic organisms may be exposed to episodic contamination events where contaminant concentrations reach toxic levels
(Abel 1980; ASTM 2003). Episodic contamination events may result from chemical spills, the application of pesticides and agrochemicals, urban and industrial runoff, aerial deposition, sewage or industrial discharge (Brent and Herricks 1998; Reinert et al. 2002). Factors such as the rate of discharge and dilution, changes to chemical form and solubility and degradation of the chemical also have an important part in determining an organisms exposure to a contaminant (Widianarko et al. 2001).
Contaminant pulses are often associated with stormwater and runoff from agricultural, urban and industrial settings. This is due to the flux of contamination that results from the runoff water washing contaminants into aquatic environments, which are then diluted over time due to natural processes such as currents and tides. Agro-chemicals in particular, have been identified as often occurring as contaminant pulses following application onto farming land. Considerable attention has been given to this group of chemicals (including chlorpyrifos, fenoxycarb and fenvalerate) as they are widely used in the agricultural industry. The effects of fluctuating concentrations of common metal
Page | 31 contaminants have also been investigated (Chandler and Green 1996; Brent and
Herricks 1998; Simpson et al. 2012)
1.7.1 Contaminant Pulses
Fluctuating discharges may result in a heterogeneous contaminant load in waters and sediments, which varies both temporally and spatially (Brent and Herricks 1998). The various contaminant exposure patterns observed from episodic discharge events can be grouped into two generalised categories which include: (a) pulse exposures – one or more isolated and brief exposure periods, and (b) fluctuating exposures – a continuous exposure to varying contaminant concentrations (Greenstein et al. 2008). For example, pesticides used in the aquaculture industry formed contaminant plumes that could be detected up to 5.5 hours after being released from fish holding nets and remained at concentrations toxic to some species of aquatic invertebrates for 3-4 hours (Ernst et al. 2001). Ernst et al. (2001) also concluded that the treatment of a single holding net with cypermethrin could result in a toxic plume that could potentially cover an area of 1 km2 and result in toxicity to sensitive aquatic species.
While static constant exposure test methods have been shown to be effective for the monitoring and regulation of continuous contaminant sources (Brent and Herricks
1998), it is rare that organisms in a natural environment will be exposed to a constant concentration of a contaminant for prolonged periods of time (Widianarko et al. 2001).
Continuous exposure conditions in the natural environment generally only occur where persistent contaminants (including metals) have accumulated in sediments
(Widianarko et al. 2001). However, the application of the technique of diffusive gradients in thin films (DGT) has revealed the presence of chemical ‘micro-niches’ in
Page | 32 surface sediments (Stockdale et al. 2009) confirming that the biogeochemistry of marine sediments is heterogeneous. Through the use of this method, Davison et al.
(2006) were able to measure metal concentrations in surface sediments at sub- millimetre resolution (100 µm). Their results identified horizontal gradients in both zinc and manganese concentrations and suggest that the sources and sinks of trace metals in sediments may be highly localised.
1.7.2 The effect of contaminant pulses on the environment
Due to the erratic nature of contaminant sources and the dynamic aquatic environment, it is rare that aquatic organisms will be exposed to a contaminant at a continuous concentration for an extended period of time. The response of aquatic organisms to fluctuating contaminant concentrations varies, with laboratory testing revealing that organism response may be dependent on the chemical contaminant, test species and the test design (van den Heuvel-Greve et al. 2007). For example, some pulsed exposure studies have reported a greater toxic response compared to results obtained from static toxicity testing (Bengtsson 1978; Brent and Herricks 1998).
Other studies show a reduction in the toxic effect of contaminants due to pulsed exposures (ASTM 2004; Brown et al. 2005). This can be extrapolated to include natural environments where chemical speciation, biological species, and the route of exposure may dictate the observed effect of contaminant pulses. Nevertheless, dissolved contaminants have been shown to trigger a toxic response in aquatic organisms, with the severity of the toxicity influenced by both the frequency and duration of the pulse
(Chandler and Green 1996; Hagopian-Schlekat et al. 2001).
Page | 33 Interest in the effect and toxicity of contaminant pulses on aquatic organisms has increased as our questions about the environmental relevance of continuous exposure assays have been raised. Some of the earlier work in this field was performed by
Handy (1994) , Morton et al. (2000), Naddy and Klaine (2001), Butcher et al. (2006) and
Diamond et al. (2006).
In general, two methods have been employed to create pulses of dissolved contaminants under laboratory conditions. As previous research on pulses has focussed on dissolved contaminants, these methods involve either rapidly removing the test media containing the contaminant and replacing it with uncontaminated media, or by using a flow through system where the contaminant is diluted or removed over time. To assess the effects of differing short-term exposure effects
Schulz and Liess (2000) compared the equivalent doses of chemicals (amount of contaminant exposure over time, µg /h). A study by Abel (2003) demonstrated that the amphipod Gammarus pulex had a higher rate of mortality despite a reduction in exposure time to Lindane. Similarly, this was also shown to be the case with chronic effects observed when Chironomus riparius was exposed to equivalent doses of cadmium (Simpson et al. 2012). Schulz and Liess (2000) observed that the caddis fly larvae of Limnephilus lunatus displayed greater effects from a one hour exposure to fenvalerate compared to a longer exposure of a lower concentration (10 hour exposure at an equivalent dose).
Post exposure effects of three common freshwater species were investigated by Brent and Herricks (1998). They found that cadmium and zinc exposures (pulsed) resulted in mortality that occurred for up to 172 hours following exposure. Exposure to zinc (at
Page | 34 typical urban runoff concentrations) for only 30 minutes resulted in 100% immobility of Ceriodaphnia dubia. Phenol exposure was also used in this study, however individuals were able to recover from the effects of the phenol once the pulse was removed.
While a review of the literature shows that aquatic organisms may be susceptible to contaminant pulses, it becomes evident that these studies focus on the effect of dissolved contaminants, primarily in the form of organic pesticides in storm water runoff. A recent study by Angel et al. (2010) shows that the benthic amphipod Melita plumulosa is susceptible to pulses of dissolved copper. It was also shown that both the duration and the frequency of the pulse is also important in determining the resulting toxicity that will occur due to the exposure.
1.8 Aims and objectives
The overall aim of this study was to develop environmentally relevant chronic toxicity test methods to build on sediment quality assessment techniques currently in use.
This was achieved through a series of studies using Australian benthic invertebrates to evaluate the suitability of new test species and bioassay endpoints for assessing sediment quality. Improving our ability to assess the impact of contaminated sediment on organisms, population dynamics and ecosystems will lead to the development of more robust sediment quality guidelines.
Chapter 3: An assessment of four harpacticoid copepod species for use in ecotoxicological testing of sediments
Page | 35 The discovery of new model species and the development of new techniques for assessing sediment quality are critical for improving our knowledge of the complex sedimentary environment. Ideally, sediment bioassays should be rapid, easy to conduct and inexpensive. It is also important that test species are sensitive to contaminants, have year-round availability from either natural populations or laboratory cultures, are ecologically significant and are widely distributed in aquatic environments. Despite the known advantages of chronic test methods, routine toxicity testing often utilise acute toxicity endpoints as these methods are often faster and more cost effective. It is therefore desirable that rapid methods for assessing chronic effects be developed. To achieve this, new model test species must be explored.
The relatively short life cycles of harpacticoid copepods makes them appropriate animals for use in tests that rapidly assess the acute, sublethal, or chronic effects of sediment contaminants. In this chapter, the suitability of four harpacticoid copepod species, Nitocra spinipes, Tisbe tenuimana, Robertgurneya hopkinsi and Halectinosoma sp. for use in ecotoxicology was investigated. Laboratory culturing procedures were determined and the influence of food type on juvenile growth and development assessed. The copepods species were ranked on a range of factors (ease of handling, culturing, rate of maturation, food type and sensitivity to dissolved copper) to identify the species most suitable for developing chronic toxicity test methods.
Chapter 4: Avoidance of contaminated sediments by three benthic estuarine species
With the ecological relevance of acute ecotoxicology methods in question, there is an emphasis on incorporating chronic toxicity tests into environmental regulation. The benefits are thought to include greater ecological relevance, protection of population
Page | 36 dynamics and improved sensitivity. Behavioural changes resulting from contaminant exposure can effectively be used as an endpoint in sediment bioassays. In particular, the avoidance of contaminated sediment by benthic invertebrates may provide an alternative to long labour intensive chronic methods. An understanding of the avoidance response of mobile benthic organisms will shed light on the response of benthic invertebrates to contaminated sediments in the environment, and allow environmentally relevant methods for assessing sediment toxicity caused by contamination to be developed.
In this chapter, the use of contaminant avoidance as an endpoint in whole sediment toxicity testing is evaluated. Three benthic estuarine species known to graze or ingest sediment particles were utilised for this assessment: an epibenthic amphipod (Melita plumulosa); a harpacticoid copepod (Nitocra spinipes); and a snail (Phallomedusa solida). The influence of varying sediment physico-chemical properties on the distribution of test organisms and the optimal exposure time to measure an avoidance response elicited by contaminated sediments were investigated for each species.
These details guided further research into optimal methods for avoidance bioassays and an assessment of each species to detect and move away from contaminated sediment.
Chapter 5: The effects of slow avoidance response to contaminated sediments on benthic invertebrates
Due to the heterogeneity of sediments, it is likely that benthic organisms will come into contact with contaminated sediment as short intermittent exposures similar to aquatic pulse exposures, which are known to elicit toxic effects in aquatic organisms.
Page | 37 However, little research has been done to determine the impact of sediment heterogeneity on the exposure of benthic organisms to sediment-bound contaminants.
Furthermore, bioassays used for sediment quality assessment typically rely on static continuous exposure of a test organism to a contaminant or contaminated sediment.
Static bioassay methods which force a continuous exposure throughout the duration of the experiment (often 10 days or more) will not suitably represent the nature of exposure of mobile benthic organisms in contaminated field locations. It was speculated that slow avoidance behaviour could result in toxicity to sensitive benthic invertebrates.
To further our understanding of exposure duration and frequency on toxicity caused by contaminated sediment, the acute and chronic toxicity associated with short,
'pulsed' exposures to contaminated sediment is assessed. The potential toxic effects resulting from short intermittent exposure to contaminated sediment were assessed for Melita plumulosa and Nitocra spinipes. The effect of exposure duration and frequency on acute and chronic toxicity endpoints was assessed following exposure to four field-contaminated sediments.
Page | 38 Chapter 2 General Methods
This chapter gives a general description of the methods and materials used to conduct the research in this thesis. Detailed descriptions of methods (including bioassay procedures) are contained in Chapters 3, 4 and 5 which give specific detail on each part of the study.
2.1 Sediment Collection
2.1.1 Site description
Sediment used for culturing amphipods and as a control in bioassay experiments was collected from Bonnet Bay, NSW Australia. Bonnet Bay is an estuarine embayment located on the Woronora River, Sydney. This sediment has been shown to be relatively homogenous in properties (Simpson et al. 2004). Typical properties and metal contaminant concentrations are shown in Table 2.1. The salinity and pH of pore water were 29‰ and 7.3, respectively (Simpson and King 2005). The low AVS concentrations, negligible dissolved sulfide, high porewater iron concentrations and redox potential of -40±50 mV indicated that the sediments were anoxic/sub-oxic but probably not undergoing significant sulfate reduction (Stumm and Morgan 1996; van
Cappellan and Gaillard 1996). The sediment was hydrous (70% water), silty (98% particles <63 µm), and contained >4% organic carbon. Acid-extractable (30 min, 1 M
HCl) metal concentrations show a high concentration of Fe present (6000 µg/g)
(Simpson et al. 2004). Concentrations of total polycyclic aromatic hydrocarbons (PAHs,
ANZECC/ARMCANZ 2000) in the sediments were <2 mg/kg (normalised to 1% organic carbon). Concentrations of other organics (e.g. pesticides, PCBs) were below analytical detection limits (Simpson et al. 2004).
Page | 39 Table 2.1 Physico-chemical properties of the Bonnet Bay sediment (modified from Simpson et al. 2004; Simpson and King 2005).
Parameters Value
Salinity (parts per thousand, ‰) 29±2
pH, redox potential (mV) 7.1-7.3, -40±50
Particle size: <63, 63-180, >180 µm (%) 96, 2.3, 1.7
Moisture (%), TOC (% DW), LOI (% DW) 70, 4.5, 11.8
AVS (µmol/g), porewater sulfide (mg/L) <0.05, <0.1
Porewater Fe, Mn (mg/L) 5-13, 0.2-1.0
Total particulate Fe, Mn (mg/kg) 24100, 71
Acid-extractable Fe, Mn (mg/kg) 6300, 49
Total particulate Cu, Pb, Zn (mg/kg) 40, 40, 210
Acid-extractable Cu, Pb, Zn (mg/kg) 30, 66, 160 DW = dry weight. Acid-extractable metals = 1 M HCl, 60 min, cold, filtered <0.45 µm. Total particulate metals = 2:1 concentrated HCl:HNO3, 24 h cold, 20 min microwave at 100 W.
Contaminated sediments were collected from estuarine field sites of unspecified locations in NSW and Tasmania or spiked in the laboratory. Analyses of physicochemical properties (OC, particle size, AVS) and metal contaminants were completed on all sediments as described by Simpson et al. (2006). Key properties of the contaminated sediments collected for this study are given in Table 2.2. The contaminated sediments obtained from field sites were found to consistently contain negligible concentrations of common organochlorine or organophosphate pesticides
(0.005-0.05 mg/kg), polychlorinated biphenyl (PCB) aroclors (<0.01-0.1), (<250 mg/kg) or polycyclic aromatic hydrocarbons (PAHs), BETX (<0.25 mg/kg benzene, toluene, ethyl benzene, xylene), and total petroleum hydrocarbons (<1 mg/kg) (Chariton et al.
2010; Simpson and Spadaro 2011).
Page | 40 Table 2.2 Sediment chemistry data for the test sediments used in this study
Test sediment Parameters a 1 2 b 3 4 5c 6 7c 8 9
Silt, % 76 91 15 79 49 27 47 1 84
AVS, µmol/g <5 <5 <5 <5 <5 <5 <5 <5 14±12
TOC, % 3.9 3.1 5.9 6.8 4.0 3.0 1.5 0.7 6.0
As, µg/g 3430 4 7 61 15 ND 18 ND 98
Cd, µg/g 83 <1 0 19 1 1.1 1 1.1 10
Cu, µg/g 1100 6 1070 108 1130 71 930 49 112
Pb, µg/g 14400 15 60 830 43 470 36 370 1730
Zn, µg/g 14500 44 260 2630 92 1230 95 1050 2800
Amphipod survival, 12±5 35±30 24±6 32±2 47±10 59±11 63±4 93±9 96±6 % ± SE a Silt = percent of particle <63 µm. AVS = acid-volatile sulfide. TOC = total organic carbon. All metal concentrations were dilute acid-extractable metals. Amphipod survival was in 10-d toxicity testing. b Sediment 2 was a 5% diesel-spiked sediment that contained 26 mg/kg total PAHs (the sum of 16 polycyclic aromatic hydrocarbons (U.S EPA 1991) and 11,500 mg/kg C10–C36 total petroleum hydrocarbons (TPHs). c Sediments were spiked with copper in the laboratory.
Page | 41 2.1.2 Sediment sampling
Oxic surface sediment (top 1-2 cm) was collected at low tide using a large acid-washed polyethylene spoon. The collected sediment was passed through a 1 mm mesh sieve in the field to remove large debris and organisms. The sieved sediment was stored in a plastic bag devoid of air, returned to the laboratory and refrigerated (4 oC) within 1 hour of collection.
2.2 Analytical Method
2.2.1 General cleaning
Pyrex glass beakers used in bioassays were washed in a laboratory grade dishwasher
(GW 3050, Gallay) on a three stage cycle; phosphorus free detergent (Gallay Clean A),
HNO3 rinse (1% v/v) and deionised water (Milli-Q) rinse. Where additional acid washing was required, glass and plastic ware was cleaned by soaking in 10% (v/v)
HNO3 for 24 hours followed by soaking and rinsing three times in deionised water.
After rinsing, all glassware and plastic ware was inverted and allowed to dry in the ambient laboratory conditions (filtered air, semi-clean room). Plastic syringes
® (Terumo ) and filters (0.45 µm, Sartorious Minisart) were rinsed with 4 % (v/v) HNO3 (2
× 10 mL) followed by Milli-Q water (2 × 10 mL) prior to sample filtration.
2.2.2 Reagents
Acids used for metal analyses were trace metal grade (Trace Pur, Merck). All other acids and chemicals were analytical reagent grade. Deionised water (18 MΩ, Millpore
Corp, USA) was used to rinse all glassware and plastic ware and in the preparation of all solutions. Seawater collected from Cronulla, NSW Australia, was filtered (0.45 µm)
Page | 42 and diluted (30 ‰ salinity) before use. Deoxygenation of seawater and Milli-Q water was achieved by bubbling with high purity nitrogen gas for ≥ 24 hours.
2.2.3 pH measurements pH measurements were made with an Activon digital pH/mV meter using a Hana HI
2031B glass body electrode with a robust spear shaped head. The instrument was calibrated using standard NIST buffers (pH 4 and pH 7) before use. The electrode was rinsed with Milli-Q water between measurements and stored in 3 M KCl solution when not in use.
2.2.4 Metal analyses by inductively coupled plasma atomic emission spectroscopy (ICP-AES)
Metal analyses were carried out using ICP-AES (CIROS, SPECTRO Analytical
Instruments, Kleve, Germany). The instrument was warmed for at least 30 minutes prior to use. A multi-element standard prepared from a commercial ICP standard
(QCD Analysts Laboratory Performance Check Solution 4, diluted to a metal concentration of 2 mg/L) and deionised water was used to calibrate the instrument.
The sensitivity of the optics was confirmed using a 2% acidified Zn solution (2 mg/L).
The multi-element standard was analysed every 10 samples to allow for drift correction when calculating metal concentrations (generally <2 % per hour). The system was flushed with deionised water for 15 seconds between samples and pre- flushed with the sample for 30 seconds prior to analysis. Matrix matched standards
(acidified deionised water or 30 ‰ seawater) were analysed with all samples to correct for signal suppression which may result from the high sodium concentration associated with seawater.
Page | 43 2.2.5 Water content
The dry weight:wet weight ratio (DW:WW) of test sediments was determined by weighing approximately 2 g of sediment (weight accurately recorded) into a pre- weighed polycarbonate vial. The vial was placed in a drying oven set at 110 ˚C overnight. After drying, the vial was tightly capped and placed in a desiccator to cool to room temperature. The DW:WW ratio was calculated gravimetrically.
2.3 Preparation of laboratory spiked sediments
2.3.1 Copper spiked sediment
All manipulations of copper spiked sediments were performed in a nitrogen gas-filled glove box to minimise oxygen intrusion into the sediments as described by Simpson et al. (2004). Test sediments were spiked with varying amounts of CuSO4·5H2O to achieve the desired concentrations (per gram dry sediment) of particulate copper phases. The CuSO4·5H2O spiking solution was prepared by dissolving Analytical
Reagent grade salt in deoxygenated seawater. The hydrolysis of water decreases pH when copper sulfate is dissolved. Therefore, the pH of copper-spike solutions were adjusted to pH 7.5 using deoxygenated NaOH (50 % NaOH w/w in deoxygenated deionised water) immediately prior to being mixed with the sediment. Low density polyethylene (LDPE) plastic bottles (Nalgene) containing the sediments (3:1 water:sediment ratio) were purged with nitrogen gas, sealed, shaken and rolled initially for 1 h, then rolled periodically (1 h, 1-2 times per week) to aid homogenisation and reaction of the spiked chemicals with the sediment. Test sediments were stored in a nitrogen glove box and allowed to equilibrate for a specified period of time (≥ 5 days). The pH was measured at least twice per week with adjustments to pH 7.5 made
Page | 44 as required (deoxygenated NaOH solution). Following pH adjustments, bottles were again purged with nitrogen, rolled for 1 h then returned to the glove box. At least two days before use of the sediments in bioassays, the sediment-water mixture was centrifuged (3000 rpm for 5 minutes) and overlying water was discarded. To re- establish the DW:WW ratio to be approximately equivalent to that before sediment spiking, clean deoxygenated seawater was added followed by complete homogenisation using a metal spatula. Sediments were stored in the glove box until use.
2.3.2 Diesel spiked sediment
The diesel spiked sediments were prepared in two stages. Diesel oil was added to sub- oxic Bonnet Bay sediment to create a 10% diesel stock sediment (dry weight). The stock sediment was placed on a bottle roller for 2-3 hours three times per week and allowed to equilibrate for two weeks. Following equilibration, an aliquot of stock sediment was combined with sub-oxic Bonnet Bay sediment to give a final concentration of 5% diesel sediment (w/w). The test sediments were homogenised by vigorous shaking of the containers followed by 2-3 h on a bottle roller at least 3 times per week during a one-month equilibration period.
All diesel-spiked sediments were maintained in glass containers with minimal headspace. The containers were sealed using parafilm prior to being tightly capped.
The capped containers were wrapped in aluminium foil and refrigerated (4 oC) to minimise photolysis and volatilisation of organic compounds. In addition, the containers were housed inside two polypropylene bags which were sealed with duct tape until required for use in sediment bioassays.
Page | 45 2.4 Collection, culturing and handling of test species
2.4.1 Amphipods
Melita plumulosa used in the tests were obtained from CSIRO (Land & Water, Lucas
Heights, NSW) laboratory-maintained cultures. The cultures of M. plumulosa were originally established with adults collected from intertidal and subtidal mud flats in the
Hawkesbury River, north of Sydney. Amphipod cultures were kept in a temperature controlled laboratory at 21 ± 2 °C in aquaria containing 1 cm deep, 1 mm sieved control sediment with clean overlying filtered seawater at a salinity of 30‰. Aquaria were covered with foil to minimise light disturbances to the amphipods with continuous gentle aeration of the overlying water. Amphipods were fed the marine microalgae Phaeodactylum tricornutum (1 x 105 cells/animal) and powdered fish food
(Sera Micron™, Sera, Heinsberg, Germany) twice weekly (0.5 mg/juvenile amphipod; 1 mg/adult amphipod). The overlying water in the aquaria was renewed weekly (by gentle siphoning) and sediment was changed every 4-6 weeks.
Amphipods were isolated from the culture populations by gentle sieving of the sediment. Adult amphipods were isolated by rinsing sediment through a 500 µm sieve immediately before the start of the test. From the isolated animals, amphipods of uniform size (4-6 mm) were chosen to be tested. For juvenile tests, gravid females were separated from the cultures 14 days prior to test commencement and placed in a separate culturing tray. Over the 14 day period, females would drop their brood of 3-6 juveniles which were then removed from the sediment by rinsing through a 180 µm sieve before being added to the test beakers. M. plumulosa life stage sizes are shown in Figure 2.4.
Page | 46 (a) (b) (c)
10 mm
Figure 2.1 Three life stages of the amphipod Melita plumulosa, (a) a full grown adult, (b) a young adult (1-2 months old) and (c) a juvenile (7-14 days old).
2.4.2 Copepods
Copepods were isolated from field collected sediments sampled from Grays Point and
Bonnet Bay estuaries, Sydney NSW Australia. The sediments were sieved in the laboratory using a 250 µm mesh to remove large particles and debris. The filtrate was observed under a light microscope (Leica MZ95, maximum magnification × 36) to identify the presence of copepods. Adult copepods were removed from the <250 µm sediment with a pasteur pipette and sorted into groups based on physical characteristics that were present (for example, the number of egg sacs, body shape or locomotive behaviour). Further monitoring of the isolated copepods was undertaken to confirm that these groups consisted of individuals from a single species. Specimens of both male and female (including gravid) individuals from each species were prepared in a 20 % ethanol solution and sent to Dr. Tomislav Karanovic, Department of
Zoological Sciences, University of Tasmania for identification. The copepods were identified as Nitocra spinipes, Tisbe tenuimana, and Robertgurneya hopkinsi.
Page | 47 Cultures of Nitocra spinipes, Tisbe tenuimana and Robertgurneya hopkinsi were established in the laboratory from the adult copepods isolated from the Bonnet Bay and Grays Point sediment. These cultures were established in clean plastic containers
(20 × 14 × 10 cm) with a 0.5 – 1 cm layer of sediment as a substrate. Initial observations showed that nematodes dominated the collected sediment, confounding copepod cultures. Therefore, sediment was sterilized (autoclaved at 121 °C for 30 min) to prevent naturally occurring meiofauna from contaminating cultures. Filtered seawater at 30 PSU was added to the cultures to a depth of approximately 8 – 10 cm.
The water was continuously aerated and renewed weekly. Cultures and bioassays were maintained in a temperature controlled laboratory (21 ± 2 °C) under ambient light conditions. Cultures were fed approximately 30 ml of a 1:1:1 mix of three algae species (Tetraselmis sp. (CS-87), Phaeodactylum tricornutum Bohlin (CS-29/4) and
Isochrysis galbana (CS-22), total concentration 1×107 cells/ml) twice weekly. New copepod cultures were initiated every 2-3 weeks by transferring 100 gravid females from the existing culture to a new culture container.
Copepods were removed from cultures by passing aliquots of surface sediment gently through a nylon mesh. Sediment was collected from the surface layer of the culture sediment with a plastic pipette and emptied into a sieve. The sieve was gently washed using clean seawater. A 180 µm mesh size was suitable for retaining adults while letting most of the sediment, debris and juveniles pass through the mesh back into the culture container. The material remaining on the sieve was transferred to a petri dish and placed under a light microscope. Adult copepods were sorted based on maturity, gender and the presence of egg sacs using a pasteur pipette.
Page | 48 (a) (b) (c)
325 µm 275 µm 275 µm
Figure 2.2 Adult specimens of the copepods species (a) Nitocra spinipes (b) Tisbe tenuimana and (c) Robertgurneya hopkinsi.
2.4.3 Estuarine Snails
Phallomedusa solida (Fam. Amphiboloidea), formerly known as Salinator solida, is one of the most common deposit feeding gastropods found to inhabit saltmarshes and mangroves in south-eastern Australia (Roach and Lim 2000). P. solida were collected at low tide from the estuarine shore of the Woronora River, Bonnet Bay. Individuals with a shell length of approximately 10 mm (measured along the ventral surface) were handpicked to ensure the collected snails were of a uniform size (Figure 2.3). It was not viable to maintain the snails in laboratory cultures, so P. solida were collected on an 'as needed' basis. Once collected, the snails were placed into a plastic container (20
× 14 × 10 cm) with a 2 cm layer of sediment collected from the estuary before being transferred back to the laboratory. Upon returning to the laboratory, filtered seawater at a salinity of 30‰ was added to the containers to give a depth of 5 - 7 cm.
Continuous aeration of the seawater was provided via aquarium tubing attached to an air pump. A lid was secured to the container to prevent the snails from escaping. The snails were fed a mixture of the marine microalgae Phaeodactylum tricornutum (1 x
105 cells/animal) and powdered fish food (Sera Micron™, Sera, Heinsberg, Germany; 1 mg/animal) immediately following the addition of the filtered seawater. Snails were stored for no longer than 5 days in the laboratory before use in sediment bioassays.
Page | 49 10 mm
Figure 2.3 The dorsal (left) and ventral (right) view of the common estuarine gastropod Phallomedusa solida.
Snails were separated from the sediment in the aquaria which they were housed in by gently washing the sediment through a 500 µm mesh sieve immediately before the start of a test. The isolated animals were then selected at random and gently placed into test containers using plastic forceps.
2.5 General data analysis
Statistical analyses were carried out using ToxCalc for Microsoft Excel (Version 5.0.23,
TidePool Scientific Software, California, 1994) or Microsoft Excel (Microsoft, Redmond,
WA). Where metal concentrations were less than the detection limit for ICP-AES, a value of half the detection limit was assigned for the calculation of averages. Data was tested for normality using the Shapiro-Wilk's test and homogeneity of variance was confirmed with an F-test or by using Bartlett's Test. These checks were conducted prior to statistical analysis of the data. Where nonparametric analysis was required,
Steel’s test was used. In avoidance experiments, the difference in the proportion of organisms on either side of the test chambers was determined by a t test, with p < 0.05 resulting in a significant difference being detected. Student's t-tests were also used to compare the response of the amphipods and copepods exposed to test sediment to
Page | 50 the response of organisms under control conditions. Significance in all statistical tests was set at p < 0.05.
Specific details of data analyses relevant to research objective are described in the
Chapters 3, 4 and 5.
Page | 51 Chapter 3 An assessment of three harpacticoid copepod species for use in ecotoxicological testing
This chapter consists of a co-authored published paper. The bibliographic details of the co- authored published paper, including all authors, are:
Ward, D. J., Perez-Landa, V., Spadaro, D. A., Simpson, S. L. and Jolley, D. F. (2011). An assessment of three harpacticoid copepod species for use in ecotoxicological testing. Archives of Environmental Contamination and Toxicology 61(3): 414-425
My contribution to the published paper involved: » Initial concept and experimental design. » Collection, analysis and interpretation of the data. » Manuscript preparation.
______
Daniel J. Ward
______
Corresponding author of published paper: Stuart L. Simpson
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Supervisor: Dianne F. Jolley
Page | 52 3.1 Introduction
Sediments act as a sink for many aquatic contaminants (Linnik and Zubenko 2000). To assess and manage contaminated sediments in coastal marine environments (such as estuaries), toxicity testing is performed. These tests identify the potential effects of the sediment-associated contaminants on benthic organisms (Simpson et al. 2005).
While a wide range of laboratory-based sediment toxicity tests are available that assess acute effects on benthic organisms (Hagopian-Schlekat et al. 2001; Adams and
Stauber 2004; Schipper et al. 2008), far fewer tests are available to assess sublethal or chronic effects (Scarlett et al. 2007; van den Heuvel-Greve et al. 2007; Mann et al.
2009).
Chronic toxicity tests utilise exposure periods which last for a longer component of an organisms life-cycle than acute tests. However, while acute toxicity tests generally have a time advantage over chronic tests, short exposure periods may not always provide an adequate assessment of effects (Finkelstein and Kern 2005; Scarlett et al.
2007). Typical chronic test endpoints include embryo development, growth, moulting and reproduction (Scarlett et al. 2007; Greenstein et al. 2008). Advantages of chronic tests may include greater ecological relevance, protection at the population level, increased sensitivity, better prediction of toxicity, and the ability to use results for modelling contaminant effects on populations dynamics (Finkelstein and Kern 2005;
Smit et al. 2006; van den Heuvel-Greve et al. 2007; Kennedy et al. 2009). For example, increased sensitivity following chronic exposures (compared to acute lethality tests) has been observed for amphipods (Castro et al. 2006), copepods (Bejarano et al. 2004;
Scarlett et al. 2007) and polychaete worms (Rice et al. 1995; Moreira et al. 2005).
Page | 53 However, chronic tests are not always more sensitive than acute methods (McGee et al. 2004; Greenstein et al. 2008), and long test durations often result in greater variability in the performance of these tests (McGee et al. 2004; Greenstein et al.
2008; Kennedy et al. 2009). The resulting increased variability will usually require greater test replication to meet quality control criteria, and this increases labour intensiveness and costs (Kennedy et al. 2009).
Ideally, sediment bioassays should be rapid, easy to conduct and inexpensive. Criteria for selecting an appropriate test species should also include sensitivity to contaminants, year-round availability from natural populations or laboratory cultures, ecological significance of the species and species distribution (ASTM 2003). Despite the many advantages of chronic tests, the disadvantage caused by the long test duration has resulted in acute toxicity methods remaining as the favoured tests for most sediment quality assessments (ASTM 2003; Simpson et al. 2005). It is therefore desirable that rapid methods for assessing chronic effects be developed.
As harpacticoid copepods are abundant in most sediment environments, are closely associated with sediments, and have relatively short life-cycles, they appear to be excellent candidate species for use in tests that rapidly assess acute, sublethal or chronic effects of sediment contaminants (Bengtsson 1978; Brown et al. 2005). For water-only testing, standard testing methods are available using the harpacticoid species, Amphiascus tenuiremis (ASTM 2004). A significant technical issue for using harpacticoid copepods in routine sediment toxicity tests is their small size, typically adults body lengths are <0.6 mm, which often makes it difficult to thoroughly and rapidly isolate the organisms from sediments. While A. tenuiremis is also used in
Page | 54 sediment toxicity tests (Chandler and Green 1996; 2001; Hagopian-Schlekat et al.
2001), the development of ecotoxicology tests that use a wider range of harpacticoid species will improve how we assess and manage contaminated sediments.
This study investigated the suitability of the harpacticoid copepod species Nitocra spinipes, Tisbe tenuimana, Robertgurneya hopkinsi and Halectinosoma sp. (isolated from sediments in south-eastern Australia) for use in ecotoxicology. Laboratory culturing procedures were developed and the influence of food type on life cycle progression (development) assessed. Time-to-death experiments were used to assess the sensitivity of each species to copper in water-only bioassays. The copepods ease of handling, culturing, rate of maturity, food selectivity and sensitivity to dissolved copper was used to rank the suitability of each species for developing toxicity test methods. Based on these experiments, the adult N. spinipes 96-h lethal effects thresholds were determined for Cd, Cu, Zn, ammonia and phenol in water-only exposures. Whole-sediment exposures were then used to assess the sensitivity of
N. spinipes reproduction to sediments collected from the field with a range of contaminant concentrations and properties.
3.2 Methods
3.2.1 Water and sediments
Seawater with salinity ranging from 30 to 34 PSU was collected from Port Hacking,
Sydney, Australia, membrane filtered (0.45 µm, MiniSart, Sartorius, Oakleigh, VIC,
Australia), and acclimated to the room temperature of 21±1 °C. Where necessary, the salinity of the filtered seawater was adjusted to the test salinity of 30 PSU using Milli-Q deionised water (18 MΩ/cm; Milli-Q Academic Water System, Sydney, Australia).
Page | 55 Long-term use of this seawater source had indicated that it did not contain any contaminants at concentrations of concern for ecotoxicology tests.
Control sediments were collected from a range of estuarine sites that had been characterised and shown to have low or negligible concentrations of metal and organic contaminants (Simpson et al. 2004). The surface layers (upper 2-4 cm) of sediments were collected using clean Teflon spatulas and press-sieved through a 1.1 mm mesh on site to remove coarse materials. The sediment was transferred into clean plastic bags with minimal headspace and stored in a cool room at 4 °C for no longer than one month. Clean sand that contained negligible contamination (Spadaro et al. 2008) was also used as a control material.
Contaminated sediments were collected from estuarine sites of unspecified locations, stored at 4°C in the dark and toxicity testing undertaken within 8 weeks. Analyses of physico-chemical properties (pH, organic carbon, particle size, acid-volatile sulfide) and metal contaminants were made on all sediments collected. Concentrations of total petroleum hydrocarbons were <250 mg/kg and polycyclic aromatic hydrocarbons were
<1 mg/kg in all sediments.
3.2.2 Copepod collection and culturing
Copepods were isolated from field collected sediments sampled from Grays Point and
Bonnet Bay, Sydney, NSW Australia. Sediments were sieved using a 250 µm mesh to remove large particles and debris. Adult copepods were isolated from the <250 µm sediment under a light microscope using a plastic Pasteur pipette. At least four morphologically distinct copepods with high abundance were easily identified from the mixtures of meiofauna isolated from field sediments, and were considered to
Page | 56 represent four species. Initial cultures of these species were started using 100-200 individuals, and copepods whose morphology could not be matched to these species groups were removed and discarded. Copepod cultures were established in clean plastic containers (20 × 14 × 10 cm) with a 0.5–1 cm layer of a mixture of silty sediment and clean sand (~50% particle size <63 µm) with weekly water changes (filtered seawater). All four species were cultured successfully, and after one month the adult copepods were sieved (50 µm nylon sieve) from each culture and used to commence new cultures. Juveniles, nauplii, copepodites and any copepods that did not match the morphology of the selected species were discarded during this process in order to purify the cultures. This process was repeated again after two months and then specimens of each cultured copepod species taken for taxonomic identification.
Nematodes became present in high numbers in many of the cultures and required monitoring and removal as they were likely to compete with copepods for food resources. The nematodes or their larvae were presumably either present in the sediments or seawater. New cultures were therefore initiated using sediment and seawater that had been sterilized by autoclaving 30 minutes at 120 °C (AA16,
Laboratory Equipment Pty Ltd, Sydney, Australia) to eliminate pre-existing meiofauna.
Filtered seawater at 30 PSU was added to the cultures to a depth of approximately 10 cm. The water was continuously aerated and renewed weekly. Cultures and bioassays were maintained in a temperature controlled laboratory (21 ± 2 °C) under ambient light conditions. The copepod cultures were initially fed approximately 30 mL of a
1:1:1 mix of three algae species (Tetraselmis sp. (CS-87), Phaeodactylum tricornutum
Bohlin (CS-29/4) and Isochrysis sp. (CS-177), total concentration 1×107 cells/mL) twice
Page | 57 weekly. New copepod cultures were initiated every 2-3 weeks by transferring 100 gravid females from the existing cultures to new culture containers.
Adult copepods were removed from cultures using a plastic Pasteur pipette to gently suck up aliquots from the surface sediment, which were passed through a 180 µm sieve, retaining adults while letting most of the sediment, debris and juveniles pass through the mesh back into the culture container. Adults were transferred from the sieve (using clean seawater) into a Petri dish, placed under a light microscope
(maximum 36× magnification) and individual copepods were sorted based on maturity, gender and the presence of egg sacs using a pasteur pipette.
3.2.3 Influence of food type and quantity on copepod culturing
To determine the influence of food type and quantity on the reproduction and development of copepods over 7 days, feeding experiments were conducted using two diets at three food concentrations. The first was a 1:1:1 mix of three planktonic algal species, Phaeodactylum tricornutum., Tetraselmis sp., Isochrysis sp. (total cell count of
1×107 cells/ml) added to the vials at 0, 1, 3 and 10×106 cells per feed. The second diet was a powered fish food (Sera® Micron) added to vials at 0, 0.15, 0.50 and 1.5 mg per feed. This fish food comprised a variety of dried marine algae, plants, fish and other protein sources to give a composition of approximately 50.2%, 8.1%, and 9.2% of protein, fat and fibre, respectively. During experiments, five gravid female copepods were placed in 5 ml polycarbonate vials containing clean filtered seawater. Food was added to each vial at time zero then again after each water change. The overlying water was exchanged every second day using a plastic Pasteur pipette to remove the overlying water in 1-2 mL portions. This water was checked under the microscope and
Page | 58 any living nauplii, copepodite or adult copepods were returned to the test vial. This ensured that the loss of biota during the test period due to handling was kept to a minimum. At the end of the 7-day test the number of nauplii, copepodite and adult copepods was recorded.
3.2.4 Toxicity test procedures
The sensitivity of the cultured copepods to dissolved copper was determined using time-to-death bioassays for adult copepods. Organisms were exposed to nominal concentrations of 0 to 800 µg/L of Cu in 0.45 μm filtered seawater prepared from
CuSO4·5H2O (BDH Laboratory Supplies). Fifteen adult copepods (excluding gravid females) were transferred to clean plastic Petri dishes (diameter of 50 mm) containing
10 mL of the exposure solution. The copepods were not fed during these bioassays.
Tests were observed twice daily by light microscope and the number of living copepods was recorded. The experiment was terminated when no surviving copepods could be found in the treatment, typically 4 to 7 d. LT50 values (time (h) to cause 50% mortality) and 95% confidence limits (95% CL) were calculated based on nominal copper concentrations.
To determine the effects of ammonia, cadmium, copper, zinc, and phenol on adult
N spinipes survival, nominal concentrations were achieved by adding the appropriate volume of seawater stock solutions of ammonia (1.4 g total NH4NO3/L), 100 g/L metal sulfate, or 10 g/L phenol (AR grade, BDH Laboratory Supplies) to test beakers. The 4- day survival of adults was determined using three replicates of 10 organisms at six nominal concentrations of each chemical. After 4 days of exposure the survivors were enumerated for each test vessel. Filtered (<0.45 µm) water samples were taken at the
Page | 59 start and completion of the tests and ammonia or dissolved metals were determined.
The calculations of LC50s (concentration that causes 50% lethality) were based on measured concentrations of copper and ammonia, but nominal concentrations were used for phenol.
3.2.5 Whole-sediment exposures using N. spinipes
The influence of sediment properties and contaminants on the reproductive output of
N. spinipes was assessed following exposure to undiluted test sediments over a 10-day period. Approximately 0.5 g of test sediment and ~9 mL of filtered seawater (30 PSU) was added to a 10 mL polycarbonate vial (10 mm diameter, 10-cm height), then each vial was incubated at 21 °C overnight to allow sediments to settle. The following day, overlying water was replaced and five gravid females (3-5 weeks old) were randomly assigned to each vial. For most tests, there were five replicates of each sediment. For several tests, six replicates were used for each sediment tested. All tests were undertaken at 21±1 °C with a 12 h light, 12 h dark cycle. Overlying water concentrations of metals and ammonia, along with physico-chemical parameters
(temperature, pH, salinity and dissolved oxygen) were measured twice during the 10- day test. At the completion of the test the number of nauplii (first juvenile life stage of the copepod) and copepodites (second juvenile life stage) in each vial was recorded by microscopy. The presence of sediment particulates made counting more difficult, however suitable recoveries were achieved by gently agitating the surface of the sediment and pipetting it into a petri dish using filtered seawater and counting these organisms. The remaining sediment from each treatment was then passed through a
20 µm sieve and the retained organisms transferred into another petri dish and counted. The organism numbers from both of these counting steps were combined. Page | 60 The surface sediments contained 80-90% of the organisms. Results were expressed as a percentage of the reproductive output (nauplii and copepodites) in the control sediment. Treatments were fed a diet of 2 × 104 cells/ml of 1:1 Isochrysis sp. and
Tetraselmis sp. (discussed below) as well as 0.3 mg Sera micron® fish food (<63 µm) per test vial twice a week.
3.2.6 Statistical Analyses
LT50 values (time (h) to cause 50% mortality) and confidence limits were calculated from logistic time–response curves for each copper concentration using a Microsoft
Excel (Redmond, WA, USA) spreadsheet (Barnes et al. 2003). Toxcalc for Microsoft
Excel (TidePool Scientific Software, Mckinleyville, CA) was used to perform the
LC50/EC50 calculations, the probit function was used when the data could be adjusted to a normal distribution. The remaining of LC50 (concentration causing 50% lethality) calculations were obtained by using the Trimmed Spearman-Karber method (Finney
1978). t-tests were used to determine if the response of the copepods in the test sediment was different to that in the control sediment. The appropriate Student’s t- test was performed on the data to determine significant differences between means of the different feeding concentrations and diet types. All statistical analyses were performed in Microsoft Excel 2003 Data Analysis Tool Pack and Solver add-in.
Variances were determined to be either equal or unequal using the two sample F-test.
Significance in all statistical tests was set at the p<0.05 level.
3.2.7 Analytical Methods
All chemicals were analytical reagent grade or equivalent analytical purity (BDH
Laboratory Supplies, Poole, England, or Univar, Ajax Finechem, Sydney, Australia).
Page | 61 Plasticware (made of polycarbonate or polyethylene) used for all tests and analyses was new or re-used following cleaning. Containers for analyses were cleaned by soaking in 10% (v/v) HNO3 (Analytical Reagent grade) for a minimum of 24 h, followed by thorough rinsing with Milli-Q water. Glass beakers and acrylic beaker-lids used for toxicity tests were cleaned in a dishwasher (Gallay Scientific Pty Ltd, Melbourne,
Australia) programmed for a phosphate-free detergent wash (Clean A, Gallay Scientific
Pty Ltd) and a dilute acid wash (1% HNO3), followed by thorough rinsing with Milli-Q water.
Measurements of pH were made using a pH probe and meter as described previously
(Simpson et al. 2004). Salinity, temperature (YSI 30, Springs, OH, USA) and dissolved oxygen measurements (MI-730 Dip-type O2 microelectrode and OM-4 oxygen meter,
Microelectrodes Inc., Bedford, NH) were made in accordance with the instrument manufacturer’s instructions. Samples for dissolved metals analyses were acidified with concentrated HNO3 (2% volume/volume, Tracepur, Merck, Darmstadt, Germany) and concentrations determined by inductively coupled plasma - atomic emission spectrometry (ICP-AES; Spectroflame EOP, Spectro Analytical Instruments, GmbH,
Kleve, Germany) calibrated with matrix-matched standards (QCD Analysts). Method blanks, method duplicates and spike recoveries were performed on at least 10% of the filtered samples. Method blanks were below the 2-10 µg/L limits of reporting, duplicates within 15% and spike-recoveries were 85-110% for all metals. Methods for measurement of sediment particle size (by wet sieving through 63 µm nylon sieves followed by gravimetry), total organic carbon (OC, Dohrmann DC-190 TOC analyzer,
Teledyne Tekmar, Mason, OH), acid-volatile sulfide (AVS) and porewater (PW)
Page | 62 extraction (centrifugation at 800 g for 5 min) have been described previously (Simpson
2001; Spadaro et al. 2008). Dissolved ammonia was analysed colorimetrically using a
Merck Spectroquant Kit (14752).
3.3 Results and Discussion
3.3.1 Copepod species and culturing
The four copepods in culture were identified as harpacticoid species Nitocra spinipes,
Tisbe tenuimana, Robertgurneya hopkinsi and Halectinosoma sp. The cultures of
Halectinosoma sp. were abandoned because of the small size of this species (304 ± 24
µm; mean ± SD, n=5) and a low degree of movement which made subsequent handling of this species difficult. The size of the harpacticoid copepods differed between species (Figure 3.1) with body lengths of adults (males and females) of 649 ± 72 µm for
Figure 3.1 Photographs of (i) Nitocra spinipes (Boeck, 1864) (~24day life cycle), (ii) Tisbe tenuimana (Giesbrecht, 1902) (~28-day life cycle), and (iii) Robertgurneya hopkinsi (Lang, 1965) (~35-day life cycle). The scale bar represents a length of 250 µm.
Page | 63 N. spinipes, 547 ± 15 µm for T. tenuimana and 548 ± 19 µm for R. hopkinsi, respectively
(average ± SD, n=5).
The copepods N. spinipes, T. tenuimana and R. hopkinsi, reproduced successfully in sediment types ranging from silty (98% <63 µm) to sandy (29% <63 µm). However, it was easier to isolate adult copepods from sediment when cultured in sediment that had been sieved to <20 µm, as the sediment could be washed through a 20 µm sieve leaving behind the copepods and juvenile copepod life stages (nauplii and copepodites). Water changes were generally performed once per week, however the cultures remained viable even when longer periods of time (up to 4 weeks) lapsed between water renewals.
3.3.2 Effect of food type on reproduction and juvenile development
Harpactiocoid copepods are known to feed on a range of food sources, including bacteria, diatoms (algae) and detritus (De Troch et al. 2006). The quality and quantity of food available to the copepods can have an impact on the growth, reproduction and mortality of these organisms. The response of copepod reproduction and development to the presence of food has previously been shown to depend on the nutritional value of the available food, the ability of copepods to select the most nutritious source/s and their nutritional requirements (Koski et al. 2006). This has been investigated previously for both pelagic (Tang and Taal 2005; Ismar et al. 2008; Saage et al. 2009) and harpacticoid species (Weiss et al. 1996; Rhodes 2003; Dahl et al. 2009).
The use of diatoms as food by copepods is known to be influenced by cell size, morphology and cellular composition as well as the morphology of the copepods
Page | 64 mandibles (Koski et al. 1998; De Troch et al. 2006). For calanoid copepods, it has been shown that dietary diversity is important in promoting zooplankton production and ensuring a nutritionally complete diet (Kleppel 1993; Anderson and Pond 2000).
Similarly, Wyckmans et al. (2007) demonstrated that offering a diverse diet to three species of harpacticoid copepod resulted in the copepods feeding on a wider range of algae diatoms with a reduction in the consumption of any one species. Wyckmans et al. (2007) also reported that the preference of diatom type was species specific. This makes sense in terms of the ‘optimal foraging’ theory which suggests that consumers will selectively feed on food resources that will maximise energy intake (Hughes 1980).
Also in accordance with the optimal foraging theory is the observation that the grazing rate of harpacticoid copepods increases in response to an increase in food availability both in the laboratory (De Troch et al. 2007) and in the field (Montagna et al. 1995).
Based on these previous studies, we chose to trial a multi-species algal diet to culture
N. spinipes, R. hopkinsi and T. tenuimana in the laboratory. Three algal species fed to the copepods were selected due to the small cell size and the availability of these species in our laboratory. While no estimates were made of the grazing rate, or the species selectivity of the copepods, it was found that the copepods had a positive response to the addition of food which was measured as an increase in offspring production (Figure 3.2). Furthermore, increasing the biomass of the added food stock resulted in an increase in the reproductive output for the three copepod species tested in this study.
Harpacticoid copepods are often detritus feeders (Norsker and Støttrup 1994) and are therefore opportunistic feeders, suggesting that the use of artificial foods may also be
Page | 65 suitable to maintain laboratory cultures. Rhodes (2003) found that the use of a formulated food, which was prepared from a mixture of vitamins, juice and brewer’s yeast was able to sustain cultures of the harpacticoid copepod Nitocra lacustris without compromising the density or growth rate of the cultures when compared to cultures fed live algae. Therefore we also chose to trial a commercially available powdered fish food (Sera® Micron) as a source of nutrition for the copepods. For
N. spinipes and R. hopkinsi, the use of a small amount of powdered fish food was shown to result in an increase in offspring production, which was not significantly different from the highest algal feed treatments. Higher reproductive output was observed for all three copepod species when fed the mixed algal diet in comparison to the powdered fish food (Figure 3.2). The number of nauplii that hatched during the 7- day period generally increased with the addition of higher concentrations of algae, up to 10×106 cells/feed for N. spinipes (Figure 3.2a), and to 3×106 cells/feed for T. tenuimana (Figure 3.2b) and R. hopkinsi (Figure 3.2c). A similar trend was observed for the development of juveniles (nauplii into copepodites) for the species N. spinipes and
T. tenuimana in which there were no copepodites present in the control treatments
(no food) for either species, however their numbers increased with the addition of food up to the high and medium treatments, respectively (Figure 3.2a and b). For R. hopkinsi, no copepodites were observed throughout the experiments (Figure 3.2c and f), possibly due to the longer life-cycle of this species (approximately 35 days nauplii (F0)-nauplii (F1) determined for R. hopkinsi compared to approximately 28 and
24 days determined for T. tenuimana and N. spinipes, respectively). When fed powdered fish food, the reproductive output varied between the three species.
Page | 66 Nauplii Copepodid
(a) (d) 250
200
150
100
50 Number of Individuals 0 (b) (e)
250
200
150
100
50 Number of Individuals 0 (c) (f) 200
150
100
50 Number of Individuals 0 Ctrl Low Med High Ctrl Low Med High
Algae Sera Micron
Figure 3.2 The reproductive output of N. spinipes, T. tenuimana and R. hopkinsi when fed a tri-algal diet (a-c, respectively; at concentrations of 0, 1, 3 and 10 ×106 cells per feed for control, low, medium and high treatments, respectively) and powered fish food (d-f, respectively; at concentrations of 0, 0.15, 0.5 and 1.5 mg per feed for control, low, medium and high treatments, respectively). The number of nauplii and copepodids were counted following 7 days of feeding at the respective diet and rate (mean ± standard error, n=4).
Page | 67 For N. spinipes, the addition of low and medium concentrations of powdered fish food resulted in a significant increase in the number of nauplii hatching, however the number of copepodites was greater in the low concentration treatment (p<0.01)
(Figure 3.2d).
The addition of the low concentration of fish food to T. tenuimana resulted in a decrease in the number of nauplii, but compared to the controls, there was a significant increase in the number of copepodites. For the high fish food treatment, the number of nauplii and copepodites significantly decreased (Figure 3.2e). The number of nauplii produced by R. hopkinsi increased when provided the lowest concentration of fish food, however there was a significant decrease in reproductive output when a greater concentration of fish food was provided. As with the algae treatment, no copepodites were observed during the R. hopkinsi experiment.
At the highest concentration of powdered fish food there was 100% mortality for adult copepods (all species), and consequently no juvenile life stages were observed
(Figure 3.2d-f). Dissolved ammonia concentrations remained near or below 9 mg total ammonia per litre (pH 7.9-8.1) during the tests and were unlikely to have been the cause of the low reproduction at higher powdered fish-food treatments. Previous studies have indicated that harpacticoid copepods are tolerant of low ammonia
+ concentrations. For example, the 96-h LC50 values for total ammonia (NH3 and NH4 ) were reported to range from 14.6 to 18.2 mg/L for adults of five harpacticoid copepod species exposed to ammonia (Di Marzio et al. 2009). However, Linden et al. (1979) reported a much lower 96-h LC50 of 4.5 mg total NH3/L for N. spinipes. LC50 experiments conducted in this study for adult N. spinipes, indicated that the lowest
Page | 68 observed effect concentration (LOEC) for dissolved ammonia was >20 mg total ammonia/L. The considerable decrease in the dissolved oxygen concentration at the higher feeding rates of powdered fish food (Figure 3.3) indicates that hypoxia could have caused the offspring reduction for the medium and high powdered fish food treatments.
For N. spinipes, increased amounts of algae as food stimulated offspring production and the rate of nauplii maturity, resulting in a positive relationship (Figure 3.2). When powdered fish food was used, a significant increase (p<0.01) in the production of nauplii was observed in the low and medium feed treatments (which were not significantly different from each other (p>0.01)) when compared to controls. As for algae, the low fish food treatment increased the rate of nauplii maturity, however the medium treatment resulted in a reduction in the rate of maturity where a significantly lower number of copepodites were produced compared to the lower fish food feeding regime (p<0.01). No organisms survived the seven days at the highest powdered fish food concentration. The response of N. spinipes to the two diets was compared to
120 10 120 10 Total ammonia (mg/L) Total ammonia (mg/L) 100 100 8 8 . . 80 80 6 6 60 60 4 4 DO (% sat) DO (% sat) 40 40 2 20 2 20
0 0 0 0 Control Low Med High Control Low Med High Algae Fish Food
Figure 3.3 Dissolved oxygen (DO, % saturation, ) and ammonia (mg/L, ■) concentrations measured in the test vials of the feeding experiment for N. spinipes.
Page | 69 determine which feeding regime triggered the best response in terms of reproductive output. It is clear from Figure 3.2a that the addition of 1×107 cells/feed (high algae treatment) produced the greatest number of nauplii and copepodites, which was not significantly different from the result obtained from the addition of 0.15 mg of Sera®
Micron per feed (low treatment) for both nauplii and copepodites (p>0.01). While both of these diets induced a similar increase in reproductive output of both
N. spinipes and R. hopkinsi, the tri-algal diet was considered to be superior as it did not result in a reduction in dissolved oxygen and is less likely to cause a build up of ammonia in culture containers (Figure 3.3). However, it remains untested whether an algae concentration greater than 1×107 cells/feed could have produced better results.
Tisbe tenuimana responded best to the medium algae treatment, resulting in a higher number of both nauplii and copepodites than the other algae treatments (including the control), and a significantly higher (p<0.01) number of nauplii than any of the fish food treatments. In addition, there was a significant (p<0.01) increase in the number of nauplii that matured to copepodites. While feeding T. tenuimana the low fish food treatment resulted in the highest number of copepodites of any treatment, the decrease in nauplii hatching made this feeding regime less favourable compared to the reproductive response induced by other feeding conditions. Based on the data presented in Figure 3.2b, it is clear that the greatest reproductive output was achieved by feeding 3×106 cells/feed.
For Robertgurneya hopkinsi, the treatments fed algae exhibited significantly greater nauplii numbers than treatments with no added food (p<0.01). Increasing the amount of algae did not increase the reproduction rate, instead it only resulted in a build up of
Page | 70 excess algae in the experimental vials. This may be attributed to the smaller size of
R. hopkinsi (compared to N. spinipes), or a lower need for food uptake as even the low algae treatment was sufficient to sustain a high reproductive output for this species.
In contrast, the addition of a powdered fish food caused an increase in the number of nauplii hatching at the low treatment, but for the medium and high fish-food treatments the number of nauplii produced decreased significantly. Once again, this was attributed to reduced dissolved oxygen. When the number of nauplii hatching in the low fish food treatment was compared to the algae treatments it became evident that offspring production was significantly higher in the algae treatments making the algal diet preferable for this species to promote reproductive output.
Our results suggest that the use of a tri-algal diet of Tetraselmis sp., Isochrysis sp. and
P. tricornutum: (a) is suitable to maintain laboratory cultures of N. spinipes, R. hopkinsi and T. tenuimana; (b) promotes reproduction among adults of the copepod species tested; and (c) is less likely to result in negative effects due to overfeeding, which was only observed in T. tenuimana. The results also suggest that the substitution of a living algal diet with powdered fish food is sufficient to maintain the health and reproduction of harpacticoid copepods if fed at concentrations that do not result in a decrease in dissolved oxygen.
Recent studies have identified deleterious effects resulting from the ingestion of some algae species by copepods (Lacoste et al. 2001; Dahl et al. 2009). This has been linked to the presence of aldehydes in these organisms which inhibit egg hatching rates and recruitment in copepods that graze on them (Miralto et al. 1999; Ianora et al. 2004).
During our study we became aware that Dahl et al. (2009) had observed lower
Page | 71 reproduction and juvenile survival in N. spinipes when fed P. tricornutum when compared to feeding with the algae Rodomonas salina. These effects could potentially be caused by the presence of aldehydes within the P. tricornutum cells, although this species has not been shown to possess these chemicals. The results obtained from our study did not indicate that using P. tricornutum within the tri-algal food mixture adversely affected the reproductive output of N. spinipes, T. tenuimana or R. hopkinsi, shown by the comparable offspring production and development observed between the algal and powdered fish food diets (Figure 3.2). However, because of potential deleterious effects from this algae species, it was omitted from the algal mix fed during subsequent feeding of cultures and in sediment toxicity experiments.
3.3.3 Sensitivity to dissolved copper
The effect of dissolved copper on survival was assessed by exposing N. spinipes to 400,
600 and 800 µg Cu/L, T. tenuimana to 50 and 200 µg Cu/L, and R. hopkinsi to 200, 400,
600 and 800 µg Cu/L for 48 h. In general, it was observed that an increase in the concentration of dissolved copper caused an increase in mortality for all three copepod species used in this study (Figure 3.4).
The time required to cause 50% lethality (LT50) at the given dissolved copper concentration was calculated to allow the sensitivity of the three species to be compared. T. tenuimana was the most sensitive of the three species, with an LT50 value of 24 (22-27) h at 50 µg Cu/L. For N. spinipes the LT50 values for 200, and 400 µg
Cu/L exposures were 114 (100-131) h and 36 (32-40) h, respectively. For R. hopkinsi the LT50 values for 200, 400, 600 and 800 µg Cu/L exposures were 119 (71-201) h, 25
(18-33) h, 10 (6-14) h and 7 (4-10) h, respectively.
Page | 72
a) b) c) 15 15 15
12 12 12
9 9 9
6 6 6 3 3 3 Number of Copepods Copepods of Number
Number of Copepods 0 0 0 0 24 48 72 0 24 48 0 24 48 72 96 TimeTime (h)(h) TimeTime (h)(h) TimeTime (h)(h)
Figure 3.4 The survival of a) N. spinipes, b) T. tenuimana and c) R. hopkinsi (mean ± SE, n= 4) when exposed to dissolved copper for ≥ 48 h. Nominal copper concentrations were 0 (○), 50 (■), 200 ( ), 400 (●), 600 (□) and 800 (▲) µg/L.
Page | 73 Past studies have indicated that dissolved copper is ineffective as a toxicant to
N. spinipes (Barnes and Stanbury 1948; Bengtsson 1978). Barnes and Stanbury (1948) found that a 24-h exposure to dissolved copper at a concentration of 260 µg/L caused
11.3% mortality in test organisms and a ten-fold increase in the copper concentration
(to 2600 µg Cu/L) only caused 21% mortality. Bengtsson (1978) determined a 96-h LC50 of 1800 µg Cu/L. Both of these results greatly differ from the results obtained in the present study and may indicate that the N. spinipes isolated in NSW, Australia, is more sensitive to copper than the Bengtsson’s isolates from the Gulf of Bothnia, Sweden.
To compare the sensitivity of these copepods to other test species, estimates were made of LC50 values. However, the data were not adequate for calculating an LC50 value for T. tenuimana, and the best point estimates calculated for the other species were for different exposure periods: 72-h LC50 value of 323 µg Cu/L for N. spinipes; and
2 96-h LC50 value of 238 µg Cu/L for R. hopkinsi. There was a strong correlation (R =
0.94) between the 72-h survival of the species R. hopkinsi and N. spinipes, which indicated that their sensitivities to dissolved copper over a 72-h exposure period were very similar.
3.3.4 Selection of copepod species for routine toxicity tests
To permit an informed decision to be made about the most suitable copepod species for use in toxicity tests, a rank was generated which incorporated five parameters based on the results of our study in addition to general laboratory issues. Handling and culturing are important parameters when considering the suitability of a test species that is to be incorporated into routine laboratory test methods. This rank was influenced by the ease of handling, which considered catching and isolating individual
Page | 74 copepods, which was dependant on the size and swimming speed of the species, and the ability to transfer the species between cultures and vials. Some species were more prone to being trapped on the surface of the water (due to surface tension) which increased the difficulty of handling the organisms. A rank was also provided for the ease of maintaining the copepod species in high density cultures under laboratory conditions, and the resilience of the population which could be estimated by the length of time cultures could be sustained before fouling occurred (due to the appearance and growth of nematodes, to which T. tenuimana were more susceptible).
The rate of species maturity was ranked based on observations in the laboratory where the life cycle of the species was determined, with species being ranked in order of shortest to longest life cycle. A shorter life cycle was considered more desirable as it would lead to a faster renewal of individuals in culture and is a desirable trait for test species of chronic life-cycle based toxicity testing. The species response to added nutrition (by observing reproduction and maturity rates) in the form of both algae and powdered fish food, and their sensitivity to copper was also ranked.
Using rankings of 1 (most favourable), 2 or 3 (least favourable), the ranking for handling, culturing, rate of maturity, food selectivity, and copper sensitivity, respectively, were 2, 1, 1, 1, and 3 for N. spinipes, 1, 3, 2, 2, 3, and 1 for T. tenuimana, and 3, 2, 3, 2, and 2 for R. hopkinsi. With a mean ranking of 1.6, N. spinipes was considered to be the most suitable species for developing toxicity tests. This species was robust, easily cultured and was reasonably sensitive to dissolved copper. While the sensitivity of N. spinipes to dissolved copper was similar to that of R. hopkinsi
(mean ranking = 2.4), the slower rate of maturity of R. hopkinsi made it less desirable as a test species for chronic life-cycle based tests. Despite being the most sensitive
Page | 75 species to dissolved copper, T. tenuimana (mean ranking = 2.0) was difficult to maintain in culture and did not respond well to the use of the powdered fish food. The high sensitivity of T. tenuimana to dissolved copper deserves further investigation, as it is possible that there are a large number of other harpacticoid copepods that may also be very sensitive to contaminants, but not amenable to use in whole-sediment toxicity tests. N. spinipes was selected as the species most suitable for future use in routine toxicity testing.
3.3.5 Sensitivity of N. spinipes to dissolved contaminants
A mean 96-h LC50 value for dissolved copper of 350±100 µg Cu/L (n=3) was determined for adult N. spinipes. The sensitivity of other copepods to copper has been determined for a range of exposure concentrations, e.g. 48-h LC50 of 256 µg/L for Tisbe battagliai
(Diz et al. 2009), 72-h LC50 of 450 µg/L for Tigropus japonicus (Kwok et al. 2008) and
96-h LC50 of 150 µg/L for Tigriopus brevicornis (Barka et al. 2001). Amphipods are commonly used for whole-sediment toxicity tests, and for comparison Ampelisca abdita has a 48-h LC50 of 30 µg Cu/L (Ho et al. 1999), while M. plumulosa a 96-h LC50 that decreases from 470 µg Cu/L for 30-d old adults to 120 µg Cu/L the 5-d old juveniles (Spadaro et al. 2008). In general, N. spinipes exhibited sensitivity to copper in a range similar to these species.
For Cd and Zn, no effects to survival were observed for concentrations up to 500 µg/L.
N. spinipes was also not very sensitive to dissolved ammonia, with a 96-h LC50 value of
300 mg total ammonia/L (pH 8). As ammonia occurs naturally in sediment pore waters, when assessing the toxicity of sediment contamination it is often useful to use species that are not highly sensitive to ammonia. For phenol, the 96-h LC50 for N.
Page | 76 spinipes was 37 mg/L. For comparison, 24-h LC50s for phenol have been reported of
1.8 mg/L for the harpacticoid T. battagliai (Smith et al. 1994) and 32 mg/L for the calanoid copepod Acartia clause (Buttino 1994).
3.3.6 Use of N. spinipes for assessing sediment toxicity
N. spinipes are iteroparous and females can produce multiple broods in the absence of a male from stored reserves of sperm. When single culture-collected gravid females were repeatedly separated from their released brood to new micro-well plates, observations made two days after separation showed they were again gravid and more nauplii had been released. The length of time this process could be repeated was not established, but it continued for more than two weeks without the females encountering a male. Females typically became gravid when food was abundant, and remained gravid for up to 48 h when they drop their egg sack and the nauplii hatch.
The development from nauplii to copepodites was found to occur over 7-9 days and copepodites were observed from approximately day 9 to 20 of development before developing to mature copepods. After approximately 25 days, gravid females began to be observed again. Utilising the ability of N. spinipes to produce multiple broods over a short period of time, the toxicity of sediments to N. spinipes was assessed by exposing gravid females to sediments for 10 days. Although both nauplii hatching and some development to copepodites occurred during this period, the endpoint used was the total reproductive output from gravid females. The significance of the iteroparous behaviour of N. spinipes during the 10 days was not expected to influence the interpretation of the results for different sediments. Along with bacteria and algae present in the test sediments, additional food in the form of 2 × 104 cells/ml of 1:1
Isochrysis sp. and Tetraselmis sp., as well as 0.3 mg Sera micron® fish food was Page | 77 provided on days 2, 5 and 8. As dicussed earlier, P. tricornutum was not used due to the potential deleterious effects from this algae species (Dahl et al. 2009). It was expected that the powdered fish food would complement the nutritional value of the algae.
The properties of the sediments tested and the reproductive output of N. spinipes is shown in Table 3.1. The four control sediments had properties ranging from silty to sandy and moderate to low TOC (Table 3.1), and the survival of gravid females was consistently >80% and reproductive output of 35±7 (28-42) (mean ± SE, range) offspring/gravid. The contaminated sediments caused significant reductions in the reproductive output (Table 3.1).
A combination of metals was likely to be the cause of the observed toxicity, with just 6 offspring/gravid (18% control) for a sediment that contained >2000 mg Zn/kg. For the contaminated sediment C, it was believed that the high concentration of ammonia in the pore waters (17 mg total ammonia/L, pH 8) was most likely responsible for the toxicity. For the contaminated sediment D, it was believed that the 15 mg/kg cadmium concentration also contributed to the observed toxicity.
3.4 Conclusion
Copepod cultures were successfully established in clean plastic containers with a
0.5–1 cm layer of a mixture of silty sediment and clean sand (~50% particle size
<63 µm) overlaid with filtered seawater (30 PSU) to a depth of approximately 10 cm.
The water was continuously aerated and renewed weekly. Cultures and bioassays were successfully maintained in a temperature controlled laboratory (21 ± 2 °C) under
Page | 78 Table 3.1 Sediment properties and effects to reproductive output of N. spinipes caused by exposure to contaminated sediments.
Controls Contaminated Parameter I II III IV A B C D E
Size <63 µm (%) 98 49 16 0.1 37 59 8 37 25
AVS (µmol/g) 5 2 40 0.1 0.4 <0.1 0.5 0.4 <1
Organic carbon (%) 4.5 2.4 1.4 0.1 2.2 4.5 0.62 2.2 3.0
Fe ( %) 2.4 1.2 0.58 0.06 0.96 3.7 0.16 0.96 2.2
Mn (mg/kg) 71 37 33 2.6 29 579 97 30 82
Cu (mg/kg) 25 13 64 1.0 10 835 35.2 37 126
Pb (mg/kg) 60 31.2 5 2.3 163 44.5 9.7 177 255
Zn (mg/kg) 216 109 120 1.2 437 401 145 431 2120
Total NH3 (mg/L) 4.5 1.0 5 0.1 1 2 17 1 1
Reproduction, % control 120±17 119±6 92±20 106±6 59±8 51±3 44±4 33±7 18±11 (mean±SE) Cd concentrations were <1 mg/kg in all sediments, except D, which had 15 mg Cd/kg. Nickel concentraions ranged from 1 to 12 mg/kg in controls and from 2 to 40 mg/kg in contaminated sediments. Fines particles <63 µm, AVS acid volatile sulfide, TOC total organic carbon.
Page | 79 ambient light conditions. The optimal feeding comprised of approximately 30 ml of a
1:1 mix of two algae species (Tetraselmis sp. (CS-87), and Isochrysis sp. (CS-177), total concentration 1×107 cells/ml) as well as 5 mg Sera micron® fish food (<63 µm) per culture twice a week, however, the amounts of each algae could be varied if build up of one species was occurring (Tetraselmis sp. is green and Isochrysis sp. is brown in colour). New copepod cultures were initiated every 2-3 weeks by transferring 100 gravid females from the existing cultures to new culture containers.
Based on the ease of handling and culturing, rate of maturity, food selectivity and sensitivity to copper, N. spinipes was determined to be the most suitable species for developing sediment toxicity tests. The measurement of total reproductive output of
N. spinipes over 10-days was found to be a useful endpoint for assessing the effects of sediment contamination. The test is relatively rapid, easy to perform using minimum sediment volumes, the endpoint relatively easy to measure and appeared to be as sensitive to sediment contaminants as other whole-sediment toxicity methods (McGee et al. 2004; Greenstein et al. 2008; Kennedy et al. 2009; Mann et al. 2009). It is likely that other test endpoints may also be available using N. spinipes, including survival, gravidity, reproduction and development of nauplii to copepodites.
Page | 80 Chapter 4 Avoidance of contaminated sediments by an amphipod (Melita plumulosa), harpacticoid copepod (Nitocra spinipes) and snail (Phallomedusa solida)
This chapter consists of a co-authored published paper. The bibliographic details of the co- authored published paper, including all authors, are:
Ward, D. J., Simpson, S. L. and Jolley, D. F. (2013). Avoidance of contaminated sediments by an amphipod (Melita plumulosa), harpacticoid copepod (Nitocra spinipes) and snail (Phallomedusa solida). Environmental Toxicology and Chemistry 32(3): 644-652
My contribution to the published paper involved: » Initial concept and experimental design. » Collection, analysis and interpretation of the data. » Manuscript preparation.
______
Daniel J. Ward
______
Corresponding author of published paper: Stuart L. Simpson
______
Supervisor: Dianne F. Jolley
Page | 81 4.1 Introduction
Environmental contaminants are known to have adverse effects on aquatic organisms.
They impact upon survival (Cairns 1992) and affect crucial behaviours such as the ability to effectively forage for food, avoid predation, find suitable habitat, compete for resources and tolerate a number of biotic and abiotic interactions within the surrounding environment (Atchison et al. 1996). Contaminant-induced behavioural changes can be linked to biochemical and physiological effects at the organism level, which can have flow-on effects at the population level (Amiard-Triquet 2009).
Therefore, significant changes to crucial behavioural traits of an organism (e.g. foraging behaviour, social interaction and locomotion) may pose a threat to the survival of a population (Weber 1997; Roast et al. 2000). Behavioural traits can be exploited as toxicity test endpoints and, in some cases, have been demonstrated to be sensitive to contaminant exposure (Lefcort et al. 2004; Lopes et al. 2004). Consequently, behavioural endpoints are increasingly being considered to have high ecological relevance, as contaminant-induced changes to an organisms behaviour offers insight into potentially major environmental impacts (Amiard-Triquet 2009). Previous studies have reported contaminant induced changes to a number of key behaviours, including predator-prey relationships (Riddell et al. 2005), swimming/mobility (Roast et al. 2000) and contaminant avoidance (Kravitz et al. 1999; Lopes et al. 2004).
The concentration of toxicants in surface sediments will often vary spatially and a number of organisms have demonstrated the ability to detect and actively avoid areas of contamination (Rakocinski et al. 1997; Chariton et al. 2010). In the natural environment, individual organisms, and eventually populations, will migrate away from
Page | 82 a contaminated site before the exposure results in the uptake of a harmful dose
(sublethal or lethal). Hence avoidance behaviour is a significant factor in determining the extent of exposure (Weber 1997; Lefcort et al. 2004), the magnitude of the hazard and the overall risk the sediment poses to ecosystem health. Avoidance is often a rapid and easily measurable response, with some suggesting that it is more sensitive than commonly used lethal and sublethal endpoints (Lefcort et al. 2004; Lopes et al. 2004).
Behavioural endpoints, such as avoidance, also have the advantage of providing an early warning to potential contamination which may result in lethal impacts (Hellou
2011). Avoidance behaviour has been investigated for a diverse range of aquatic organisms including fish (Exley 2000), amphipods (Kravitz et al. 1999), cladocerans
(Lopes et al. 2004), aquatic snails (Lefcort et al. 2004) and bivalves (Roper et al. 1995).
These studies have all reported the usefulness of contaminant avoidance as a behavioural ecotoxicological endpoint.
Contaminant avoidance is likely to be a defence mechanism against the effect of poor environmental conditions, eliminating the threat to survival caused by the presence of toxicants (Lefcort et al. 2004). Contaminant avoidance may also be an adaptive behaviour (Eriksson et al. 2006). Lefcort et al. (2004) observed that aquatic snails collected from contaminated environments could avoid contaminated sediment, while those from reference sites could not, inferring a potential genetic influence on avoidance behaviour. Regardless of the mechanisms triggering the avoidance response, it is important to note that the relocation of a population of aquatic organisms may ensure their survival at a new site. On an ecological scale, the disappearance of the species from the previous location is equivalent to the loss of the
Page | 83 entire population (Roast et al. 2000; Lopes et al. 2004). Consequently, large areas of sediments that are avoided due to contamination may be classified as high risk although short-term effects to many species used for toxicity tests may not be observed.
Static toxicity tests rarely reflect the dynamic and heterogeneous natural environment
(Lefcort et al. 2004) and therefore do not always adequately mimic an organisms exposure to contaminants in the environment, suggesting that these methods lack ecological significance (Cairns 1992; Burton et al. 2000). Despite this, many researchers and regulatory guidelines rely on standard methods which confine test organisms to a contaminant forcing a continuous static exposure (ASTM 2003; Simpson et al. 2005).
The relevance of these common ecotoxicology methods is widely debated in the literature, with chronic test methods often preferred over acute toxicity testing (Cairns
1992; Lopes et al. 2004). Chronic toxicity tests are considered to offer greater ecological relevance, protection at the population level, increased sensitivity, better prediction of toxicity, and the ability to model population effects (Kennedy et al. 2009).
However, chronic toxicity tests generally use long exposure periods that often exceed
20 days (U.S. Environmental Protection Authority 2001; Egeler et al. 2010). Considering the ability of mobile aquatic organisms to avoid contamination, the long test durations often associated with chronic tests do not adequately represent realistic exposure conditions in the natural environment. Understanding the avoidance response of mobile benthic organisms will shed light on the response of benthic invertebrates to contaminated sediments in the environment, and allow the development of
Page | 84 environmentally relevant methods for assessing sediment toxicity caused by contamination in heterogeneous field settings.
The present study investigated the suitability of contaminant avoidance as an endpoint in whole sediment toxicity testing using three species of benthic estuarine organisms: an epibenthic amphipod (Melita plumulosa); a harpacticoid copepod (Nitocra spinipes); and a snail (Phallomedusa solida). All of these species are known to graze or ingest sediment particles, making them suitable candidates for sediment toxicity testing
(Roach and Lim 2000; Spadaro et al. 2008; Perez-Landa and Simpson 2011). For each species the influence of varying physico-chemical properties of the sediment on the distribution of test organisms and the optimal exposure time to measure an avoidance response elicited by contaminated sediments were initially determined. This provided optimal methods for avoidance bioassays for assessing the ability of each species to detect and move away from contaminated sediment.
4.1.1 Materials and Methods
4.1.2 Test media
Clean seawater was collected from Port Hacking, Sydney, Australia, membrane filtered
(0.45 µm), and acclimated to the room temperature of 21±1°C. Where necessary, the salinity of the filtered seawater was adjusted to the test salinity of 30 PSU using deionised water (18 MΩ/cm; Milli-Q Academic Water System).
Relatively clean silty sediments were collected from the Bonnet Bay estuary, Port
Hacking, Sydney, Australia. These sediments had low or negligible concentrations of metal and organic contaminants and have been demonstrated to not cause toxicity effects to the organisms used in the study (Perez-Landa and Simpson 2011; Simpson
Page | 85 and Spadaro 2011). The surface layers (upper 2–4 cm) of sediments were collected using clean Teflon spatulas and press-sieved through a 1.1-mm mesh on-site to remove coarse materials. The sediment was transferred into clean plastic bags with minimal headspace and stored in a cool room at 4oC for no longer than 1 month. Clean sand that contained negligible contamination (Spadaro et al. 2008) was also used as a control material in sediment avoidance tests. Sediments with 50% and 10% silt were created by mixing the control sediment with clean sand.
The contaminated sediments were chosen to provide exposure to a range of contaminants. Contaminated sediments were collected from estuarine field sites in
New South Wales and Tasmania, Australia. Collected sediments were stored at 4oC in the dark, and toxicity testing undertaken within 8 weeks (Chariton et al. 2010; Simpson and Spadaro 2011). Analyses of physicochemical properties (pH, organic carbon [OC], particle size, acid-volatile sulfide) and metal contaminants were made on all sediments. Concentrations of total petroleum hydrocarbons were <250 mg/kg and polycyclic aromatic hydrocarbons were <1 mg/kg in all sediments (Simpson et al. 2006;
Chariton et al. 2010; Simpson and Spadaro 2011).
Copper-spiked sediments were prepared using the clean silty sediment (collected from the Bonnet Bay reference site) and equilibrated as described previously (Simpson et al. 2004). The preparation, manipulation and equilibration of the copper-spiked sediments were made within a nitrogen gas-filled glove box at room temperature. The sediments were thoroughly homogenised with a plastic spoon and a bottle roller for 2 h at least two times per week. The sediments were adjusted to pH 7.5 with 1 M NaOH one day after copper-spiking and maintained at this pH by using small additions of
Page | 86 NaOH as required during the one-month equilibration period. Changes in pH, redox potential, and dissolved metals in the pore water were monitored during the equilibration period.
The diesel spiked sediments were prepared in two stages. Diesel oil was added to sub- oxic Bonnet Bay sediment to create a 10% diesel stock sediment. Following a two week equilibration period, an aliquot of stock sediment was combined with sub-oxic
Bonnet Bay sediment to give a final concentration of 5% diesel sediment. The diesel- spiked sediments were maintained in glass containers with minimal headspace. Lids were securely fastened to avoid losses through volatilisation. The containers were housed inside two polypropylene bags which were sealed with duct tape. Both the stock and test sediments were homogenised by vigorous shaking of the containers followed by 2-3 h on a bottle roller at least 3 times per week during the one-month equilibration period.
4.1.3 Test species
All three species used in the present study were epi-benthic deposit feeders found in intertidal estuarine environments of south eastern Australia. Melita plumulosa (Fam.
Melitidae) is commonly found in estuarine tidal mudflats ranging from silty to sandy sediments in freshwater, estuarine and marine environments throughout south- eastern Australia (King et al. 2006a). Adult specimens of M. plumulosa typically range from 8-10 mm in length. The harpacticoid copepod species Nitocra spinipes (Fam.
Ameiridae) is known to adapt to a wide range of environmental conditions (including salinity and temperature) and as such has a world-wide distribution (Ward et al. 2011).
Mature copepods of this species are approximately 400 µm long. Phallomedusa solida
Page | 87 (Fam. Amphiboloidea), formerly known as Salinator solida, is one of the most common deposit feeding gastropods found to inhabit saltmarshes and mangroves in south- eastern Australia (Roach and Lim 2000). Snails of a uniform size (approximately 10 mm along the ventral surface) were collected for avoidance tests.
M. plumulosa and N. spinipes were obtained from laboratory cultures, maintained as per King et al. (2006a) and Ward et al. (2011), respectively. Specimens of P. solida were collected from natural populations occurring in the intertidal mangroves at
Bonnet Bay, Sydney, Australia.
4.2 Amphipod toxicity testing
Glass beakers and acrylic beaker-lids used for toxicity tests were cleaned in a dishwasher (Gallay Scientific Pty Ltd) programmed for a phosphate-free detergent wash, a dilute acid wash (1% HNO3), followed by thorough rinsing with Milli-Q water.
The 10-d whole-sediment toxicity tests with M. plumulosa were conducted to assess the effects of continuous exposure to the sediments and were performed in accordance with standard protocols (Spadaro et al. 2008). In brief, tests were performed at 21 ± 1˚C in a constant environmental chamber (Labec Refrigerated
Cycling Incubator) on a 12-h light/ 12-h dark cycle (light intensity = 3.5 μmol photons/s/m2). Dissolved oxygen (>85%), pH (7.5-8.2), salinity (30 ± 1 PSU) and temperature (21 ± 1°C) were monitored and maintained according to Spadaro et al.
(2008). For whole-sediment toxicity tests, three replicate 250 mL beakers containing
20 g of test sediment, 200 mL seawater and 15 adult M. plumulosa were used per treatment. Every two to three days, 80% of the overlying water was replaced with clean seawater. In all tests, water samples were collected before and after each water
Page | 88 change for dissolved metal copper analysis. No food was added during the 10-d bioassays.
At the termination of the tests, the contents of each beaker were gently sieved through a 180 µm stainless steel mesh sieve and the contents transferred to large amphipod counting trays. Live amphipods were identified by movement and counted.
The sieved sediment was transferred to 120 mL polycarbonate vials with 100 mL of seawater, fixed with 4 mL of 10% v/v neutral phosphate buffered formalin and stained with 5 mL of Rose Bengal solution (0.1 g Rose Bengal sodium salt (Sigma) per 100 mL
Milli-Q). The sediments were left for 72 h to enable any surviving amphipods missed in the initial count to take up the stain and be counted. Tests were considered acceptable if the physico-chemical parameters in beakers remained within the limits of pH 7.7-8.2, 20-22 °C, dissolved oxygen >80% saturation, and salinity 28-32 PSU throughout the test, and if survival of amphipods was on average ≥80% in the controls.
4.2.1 Contaminant avoidance experimental design
Avoidance assays were conducted in test chambers specifically designed and constructed for each test species (Figure 4.1). Each design utilised a plastic vessel divided into two chambers by a permanent barrier fixed across the bottom of the container. The top of this barrier was at the sediment-water interface and allowed the contiguous placement of two sediments in the same container with minimal mixing of the sediment samples. Removable barriers were constructed from polyethylene terephthalate plastic film and were used to divide the test chambers from the base of the sediment to the surface of the overlying water. This barrier was used during the initiation and termination of the experiments to prevent the migration of test
Page | 89 a) i) ii)
Chamber 1 Chamber 1
Chamber 2 Chamber 2
b) i) ii)
Chamber 1 Chamber 1
Chamber 2 Chamber 2
Figure 4.1 Schematic diagrams of the experimental vessels used in a) M. plumulosa and P. solida, and b) N. spinipes avoidance experiments. The temporary barrier is shown i) in place (blue cross-hatching) and ii) removed from the avoidance vessels. organisms between the test chambers. The non-toxic nature of all materials and adhesives used was assessed before being used to construct the testing apparatus.
Test chambers were set up with sediment and filtered seawater and equilibrated for
24 h before the initiation of the experiment. The overlying water was replaced immediately before the test organisms were added.
For M. plumulosa and P. solida, test chambers were constructed from 3 L polyethylene containers (20 × 15 cm wide, 10 cm high) (Figure 4.1a). A plastic barrier approximately
1 cm high was adhered to the base across the middle of the container with non-toxic aquarium silicone (Sellys Glass Silicone). Homogenised sediments were placed on the relevant sides of the test chamber and clean filtered seawater was added to achieve a
Page | 90 depth of ~8 cm for M. plumulosa and ~5 cm for P. solida. For P. solida experiments, gel
(Lucas’ Pawpaw Ointment) was applied to the chamber walls above the water line to prevent the snails from escaping during testing (this showed no toxic effect in control studies).
The test chambers used for the N. spinipes were constructed from 1 × 1 cm wide, 5 cm high polyethylene cells usually used for ultraviolet-visible spectrophotometer
(Figure 4.1b). An opening (0.5 cm wide × 2.5 cm tall) was cut into one side of each cell.
Two cells were placed together with the opening facing each other and secured in place with adhesive tape. A sediment slurry was created (10 g sediment in 2 ml seawater) and added drop-wise to the cells to achieve a ~0.5 cm deep layer of sediment overlain by ~3.5 cm of filtered seawater. Teflon spatulas were cut down and used to partition the two cells of the test chamber.
4.2.2 Contaminant avoidance assays
M. plumulosa and N. spinipes avoidance experiments were initiated by seeding one side of the sediment chamber with 30 test organisms. In avoidance assays where contaminated sediment was utilised, M. plumulosa and N. spinipes were always seeded into the chamber containing contaminated sediment. The removable barrier that divided the overlying water and prevented the movement of organisms between treatments was removed after 10 minutes and the organisms were permitted to move around the test chamber as desired. Because the snails had a lower mobility than the amphipods and copepods the avoidance tests with P. solida were initiated by placing
20 snails in a line across the middle of the test chamber directly above the permanent barrier (no removable barrier was necessary).
Page | 91 All avoidance experiments were conducted under ambient conditions in a temperature controlled laboratory maintained at 21 ± 1˚C. For M. plumulosa and P. solida tests, the overlying water was gently aerated to maintain dissolved oxygen concentrations >85%.
Overlying water in N. spinipes experiments was not aerated as the bubbling disturbed the sediment. Previous experiments with this species indicated that aeration of the overlying water was not necessary over short time periods (<7 days) to maintain sufficient levels of dissolved oxygen (Chapter 3; Ward et al. 2011). Sub-samples of the overlying water were collected and analysed for dissolved methods (as described above) every 2 to 3 days during the test and immediately prior to test termination. No food was added to the test chambers for the duration of the test. Tests were terminated by inserting the temporary barriers into the test vessels. The sediment on both sides of the barrier was gently sieved and the number of animals in the seeded and non-seeded chambers was recorded. For quality control purposes, an average organism recovery of ≥ 80% was required for a test to be considered acceptable.
4.2.3 General chemistry
New plastic ware was used for all chemical analyses. All chemicals used were analytical reagent grade or equivalent analytical purity. Measurements of pH, salinity, temperature and dissolved oxygen were made in accordance with the instrument manufacturers’ instructions. Deoxygenated waters were prepared by bubbling solutions with high purity oxygen-free nitrogen gas for >8 h to give dissolved oxygen concentrations <0.1 mg/L. The measurement of pH, particle size distribution (wet sieving and gravimetry), organic carbon (high temperature TOC analyser), and particulate metals (2:1 concentrated HCl:HNO3, heated), were made as described previously (Spadaro et al. 2008; Strom et al. 2011). Overlying water samples were Page | 92 rapidly filtered through acid-washed 0.45 µm membrane filters (Minisart, Sartorius) immediately following collection and acidified to 2% HNO3 (v/v) with concentrated
HNO3 (Tracepure, Merck). Acid-volatile sulfide (AVS) and simultaneously extracted metals (SEM) were analysed according to Simpson (2001). Dissolved metal concentrations in water samples and digested sediments were determined by inductively coupled plasma-atomic emission spectrometry (ICP-AES, Varian 730-ES,
Varian Australia) calibrated with matrix-matched standards as described in Angel et al.
(2010). Analyses of filter and digest blanks, replicates for 20% of samples, analyte sample-spikes and the certified reference material (PACS-2, National Research Council
Canada) were made as part of the quality assurance and recoveries were within 85-
110% of expected values. The limits of reporting for the various methods were less than one tenth of the lowest reported values. All sediment related concentrations have been reported on a dry mass basis.
4.2.4 Data Analyses
All statistical analyses were carried out using Microsoft Excel (2007) data analysis tool pack. The distribution of organisms in avoidance bioassays were reported as a percent of recovered organisms to account for small differences in the recovery rate of organisms between replicates. The difference in the proportion of organisms on either side of the test chambers was determined by a t-test, with p<0.05 resulting in a significant difference being detected. In all cases, avoidance was significant when there was >10% difference in the mean distribution of organisms (between the contaminated and non-contaminated sides of the test vessel) and a lack of overlap of standard errors. Where a difference in the mean distribution of test organisms was
<10% (equivalent to a difference of <3 animals), avoidance was considered not to have Page | 93 to occurred. This variation was assumed to reflect natural variability in bioassay experiments and t-tests were not performed.
4.3 Results and Discussion
4.3.1 Grain size and avoidance behaviour in uncontaminated sediment
Sediments can vary greatly in physical and chemical properties, and these variations can influence the nature of the resident ecosystem structure (Rakocinski et al. 1997;
Chariton et al. 2010). Before the avoidance of M. plumulosa, N. spinipes and P. solida was determined for contaminated sediments, it was first determined whether varying sediment grain-size and organic carbon content influenced the behaviour of organisms when exposed to uncontaminated sediments (Table 4.1).
Table 4.1 Properties of uncontaminated sediments.
A B C D Parameter >90% silt 50% silta 10 % silta 100% sand
Size <63 µm, % 98 49.1 9.9 <0.1
AVS, µmol/g 4.5 2.3 0.5 <0.01
Total Organic carbon, % 4.7 2.4 0.6 <0.1
Fe, % 2.4 1.2 0.3 0.06
Mn, µg/g 71 37 9.8 3
Cu, µg/g 13 7 2.2 1
Pb, µg/g 40 21 5.8 2
Zn, µg/g 210 110 24.3 3
Total PAHs, mg/kg 5 3.5 2.3 <2
Porewater NH3, mg/L <5 3 1.4 <1 a Calculated based on the mixing ratio of >90% silt and 100% sand sediments.
Page | 94 In vessels with uncontaminated silty sediment in both chambers, each of the test species dispersed to create a relatively even distribution of the test organisms on both sides of the test vessel over 48 h (Figure 4.1). This confirmed that M. plumulosa, N. spinipes and P. solida could move freely around the test chambers and had a preference to spread out rather than remain in a high density group. This behavioural characteristic has been shown for several estuarine species including the amphipod
Eohaustorius estuarius (Kravitz et al. 1999) and Rhepoxynuis spp. (Oakden 1984). Since this is considered normal behaviour, the even distribution of the test organisms also suggests that ambient laboratory conditions were not influencing organism dispersal.
When exposed to two different types of uncontaminated sediments M. plumulosa,
N. spinipes and P. solida dispersed throughout the vessel as they had when contained in vessels with the same sediment on both sides (Figure 4.2). For all three species, there was no significant difference in the distribution of the test organisms between the silt reference sediment and the silt/sand sediment mixtures. For M. plumulosa, a minor preference for 50% silt over the reference sediment occurred (Figure 4.2a).
However, a 9.4% difference in the distribution of these organisms between the sediments was considered to be the result of natural variability and not represent an avoidance response.
Page | 95
>> 90%90% silt silt 50%50% silt silt 10%10% silt silt a) A) bB)) C)c) 60 60 60
40 40 40
20 20 20 Snail distribution (%) distribution Snail Copepod Copepod (%) distribution (%) distribution Amphipod 0 0 0 > 90% silt 50% silt 10% silt > 90% silt 50% silt 10% silt > 90% silt 50% silt 10% silt
Figure 4.2 The distribution of (a) Melita plumulosa , (b) Nitocra spinipes and (c) Phallomedusa solida when exposed for 48 h to sediments of differing silt composition (>90%-, 50%- and 10% silt, respectively, mean ± SE, n = 3; n = 4 for N. spinipes). The filled bars represent control sedimentand the patterned bars represent
Page | 96 The sediment mixtures tested in these experiments encompassed a range of sediment characteristics including particle, AVS, TOC and metal concentrations (Table 4.1). In past studies, estuarine snails (Ilyanassa obsolete) exposed to coarse sand containing no organic carbon showed a preference for silty sediment. This preference was attributed to the availability of food (as organic matter) in the silty sediment (Marklevitz et al.
2008a; 2008b). Meadows (1964) reported that the amphipod Corophium volutator also had a preference for burrowing in finer grained sediments compared to sandy sediments. The contaminated sediments used in the present study (Table 4.2), all contained ≥10% silt (except Sediment 8) and had TOC and AVS levels similar to those tested in the grain size avoidance study (Table 4.1). Therefore the properties of the test and reference sediments were not expected to influence the avoidance behaviour of M. plumulosa, N. spinipes and P. solida in subsequent experiments.
4.3.2 Time to response: dispersal and avoidance of uncontaminated sediment
The time required to observe an avoidance response of M. plumulosa, N. spinipes and
P. solida exposed to contaminated sediment was tested for time periods ranging from
3 to 240 h. Under control conditions, M. plumulosa distribution was variable for shorter test durations (36±21 to 64±21% at 6 h for seeded and non-seeded sides, respectively). However, for longer durations (24 to 240 h) an even distribution of organisms occurred (Figure 4.3a). This indicated that M. plumulosa were actively dispersing and occupying available space in the test vessel within 6 to 24 h, but a 48 h exposure period induced an avoidance response with less variability.
Page | 97 Table 4.2 Sediment chemistry data for the test sediments used in this study
Test sediment Parameters a 1 2 b 3 4 5c 6 7c 8 9
Silt, % 76 91 15 79 49 27 47 1 84
AVS, µmol/g <5 <5 <5 <5 <5 <5 <5 <5 14±12
TOC, % 3.9 3.1 5.9 6.8 4.0 3.0 1.5 0.7 6.0
As, µg/g 3430 4 7 61 15 ND 18 ND 98
Cd, µg/g 83 <1 0 19 1 1.1 1 1.1 10
Cu, µg/g 1100 6 1070 108 1130 71 930 49 112
Pb, µg/g 14400 15 60 830 43 470 36 370 1730
Zn, µg/g 14500 44 260 2630 92 1230 95 1050 2800
Amphipod 12±5 35±30% 24±6 32±2 47±10 59±11 63±4 93±9 96±6 survival, % ± SE a Silt = percent of particle <63 µm. AVS = acid-volatile sulfide. TOC = total organic carbon. All metal concentrations were dilute acid-extractable metals. Amphipod survival was in 10-d toxicity testing. b Sediment 2 was a 5% diesel-spiked sediment that contained 26 mg/kg total PAHs (the sum of 16 polycyclic aromatic hydrocarbons (U.S EPA 1991) and 11,500 mg/kg C10–C36 total petroleum hydrocarbons (TPHs). c Sediments were spiked with copper in the laboratory.
The distribution of N. spinipes was also considerably variable within the first 24 h
Figure 4.3B), with shorter experiment durations having a greater proportion of organisms remaining in the seeded side of the test vessel (85±4 and 61±8 % for 1 and 3 h exposures, respectively). Exposure times of 24 h were sufficient for the copepods to disperse across both sides of the test vessel, and were used for the remaining avoidance studies with N. spinipes.
The snail, P. solida, dispersed evenly throughout the test vessel after 6 h (48±4% on the seeded side at 6 h) (Figure 4.3C). Minimal movement was observed in the snails at
Page | 98 A) 100
80
60
40
20 Amphipod distribution (%) distribution Amphipod Amphipoddistribution (%) 0 6 h 24 h 48 h 240 h 6 h 24 h 48 h 240 h Control Sediment Sediment 5 B) 100
80
60
40
20 Copepod distributionCopepod (%) 0 Copepod Copepod (%) distribution 1 h 3 h 6 h 24 h 1 h 3 h 24 h 3 h 6 h 24 h Control Sediment Sediment 3 Sediment 1 C) 100
80
60
40
20 Snail distribution (%) Snail distribution (%) distribution Snail
0 6 h 16 h 24 h 48 h 16 h 24 h 48 h 6 h 24 h 48 h Control Sediment Sediment 7 Sediment 6
Figure 4.3 The observed avoidance response of organisms (mean ± SE, n = 3; n = 4 for N. spinipes) when exposed to contaminated sediments: a) Melita plumulosa to Sediment 5; b) Nitocra spinipes to Sediment 3 and Sediment 1; and c) Phallomedusa solida to Sediment 7 and Sediment 6. The bar on the left of each pair represents the seeded side. The filled bars represent control sediments and the striped bars represent contaminated sediment.
Page | 99 shorter time periods, therefore shorter test durations were not included in the remainder of the study. In contrast to the amphipod and copepod, the distribution of
P. solida between the two chambers of the test vessel did not change significantly as the duration of the experiments increased.
The difference in time required to achieve an even distribution of the test organisms was expected to be influenced by both the mobility of the species and the starting position at the beginning of the assay. While P. solida was much less mobile than
M. plumulosa and N. spinipes (personal observation), the shorter dispersal time for
P. solida is most likely the result of placing them along the centre barrier dividing the two test chambers. As the copepods and amphipods were distributed throughout the seeded chamber, the organisms potentially had a greater distance to travel than the snails before sensory systems could trigger an avoidance response to the less desirable areas of contaminated sediment.
Zylstra (1971) found that dermal sensory cells in the freshwater pulmonate snail
Lymnaea stagnalis were prevalent on the surface of the tentacles as well as around the lips, the front edge of the foot and the mantle edge. The concentration of sensory cells on the tentacles, mouth and margins of the lips reached densities up to 5,000 cells/mm2, resulting in a greater number of peripheral sensory neurons than in the central nervous system (Wyeth and Croll 2011). This highlights the sensitivity of aquatic snails to their surrounding chemical environment and that chemoreception is an important factor controlling behaviour of these organisms. Aquatic pulmonate snails have been shown to be extremely sensitive to chemicals cues in the environment. For example, Physella columbiana has demonstrated the ability to avoid
Page | 100 soil solutions with a mixture of zinc and cadmium at concentrations <10 µg/L (Lefcort et al. 2004). Therefore, we speculate that the chemoreceptors of P. solida are more sensitive than those of N. spinipes and M. plumulosa and as such enhance the ability of this species to ‘sense’ the surrounding environment and move away from unfavourable conditions.
Organism density within a population may be a factor contributing to the speed at which the organisms disperse, for example, Rosenberg et al. (1997) found that a higher density of brittle star Amphiura filiformis resulted in a significant increase in the rate of dispersal. Estuarine snails have also been shown to exhibit behavioural changes in response to high organism density, including an increase in the incidence of floating, emigration, climbing behaviour as well as a reduction in feeding and crawling rates
(Levinton 1979; Lopez-Figueroa and Niell 1987). If aquatic organisms prefer low spatial densities, then physically larger species should react to overcrowding and more actively disperse within available space. This behaviour was previously observed for two species of estuarine snail which displayed an increase in dispersal among larger
(mature) specimens when resources were limited (Byers 2000). As P. solida was the largest of the test species used in the present study, it is reasonable to expect that population density related effects may occur and facilitate the dispersal rate. If considering only the dispersal of the test organisms under control conditions, it is possible that a density-dependant effect may encourage the spreading out of organisms in the test vessel and this may be more significant for P. solida. We did not undertake tests to determine whether this may be the case. However, based on the results from the contaminant avoidance experiments, we found that under the same
Page | 101 density constraints, all three test species showed a preference to migrate away from the contaminated sediment and occupy the clean sediment despite the resulting higher density of organisms.
4.3.3 Avoidance of contaminated sediment
The ability of the three species to avoid contaminated sediments was tested using the sediments shown in Table 4.2. The percent survival in 10-d lethality tests using the amphipod M. plumulosa is also shown. These were field collected sediments, except for the diesel-spiked Sediment 2, and copper-spiked Sediments 5 and 7.
To determine whether the behaviour of the test organisms was altered by the presence of contaminants, contaminated sediments were placed on one side of the test containers, and uncontaminated reference sediment on the other. If the organisms no longer dispersed throughout the test vessel, but remained at a higher density on the clean reference sediment, then this would be sufficient evidence that the chemical stimuli (in this case mostly metal contaminants) altered the behaviour of the test organisms (Figure 4.3). The time to achieve uniform distribution of M. plumulosa, N. spinipes and P. solida in uncontaminated sediments under control conditions was 48, 24 and 6 h, respectively, and changes to these times would also be used to provide information on avoidance behaviour.
M. plumulosa were placed in the test vessels with the reference and copper spiked sediment (Sediment 5). After just 6 h exposure the distribution of M. plumulosa was significantly lower in the contaminated seeded side (Sediment 5) compared to the uncontaminated non-seeded side of the test chamber, with a distribution of 24 ± 4 to
76 ± 4 %, respectively (Figure 4.3A). Exposure durations of 24, 48 and 240 h showed a
Page | 102 similar distribution to the 6 h result, with at least 85% of amphipods residing in the uncontaminated sediment, despite the extended time period. However, there was significantly more M. plumulosa avoiding Sediment 5 following a 24 h exposure (95 ± 2
%) than after a 48 or 240 h exposure (88 ± 5 % and 85 ± 4 %, respectively). Overall, M. plumulosa maintained their preference to inhabit ‘clean’ sediments over longer exposure times, even though they existed at a higher population density (number of organisms per given area) than observed under control conditions.
While M. plumulosa clearly avoided Sediment 5 after a 6-h exposure period, the variability of the amphipod distribution at 6 and 24 h in uncontaminated (control) treatments made interpretation of the contaminated sediment results difficult (Figure
4.3A). A longer exposure period of 48-h was chosen for future avoidance experiments as it allowed sufficient time for the organisms in the control sediments to disperse and interact with the sediment before terminating the experiment.
The copepod N. spinipes had the fastest avoidance time of the three test species.
When exposed to contaminated sediments 1 and 3 (Table 4.2) > 87% of copepods occupied the non-contaminated chamber of the test vessel after only 3 h of exposure
(Figure 4.3B). This is a highly significant proportion of the copepod population migrating away from the contaminated sediment, showing a clear preference for the non-contaminated sediment. For Sediment 3, 1 h exposure resulted in 87±4% of copepods moving to the uncontaminated sediments in the test chamber. For both sediments, there was no significant difference between the number of ‘avoiding’ organisms at the termination of the experiment for all exposure times, hence the increased exposure duration did not elicit a greater avoidance effect.
Page | 103 When exposed to Sediment 7, P. solida had an avoidance time of ≥ 16 h with a significant number of snails (61 ± 2 %) avoiding the contaminated sediment (Figure
4.3C). This avoidance increased over time with 48 h resulting in 78 ± 10 % of P. solida migrating to the uncontaminated sediment. When exposed to Sediment 6, avoidance occurred after a shorter time than for Sediment 7, with 85 ± 6 % of snails migrating away from the contaminated sediment after 6 h.
The avoidance response time of P. solida was similar to the 6 to 24 h avoidance time of the estuarine snail Ilyanassa obsoleta when exposed to clean and contaminated harbour sediments (Marklevitz et al. 2008a; 2008b). P. solida maintained their preference for the control sediment despite extending the duration of exposure to as much as 48 h. Overall, the results show that in the presence of an unfavourable chemical stimulus, all three species migrate away from the source and remain there for extended periods of time even if this results in relatively high densities of organisms.
4.3.4 Validation of sediment avoidance of M. plumulosa
M. plumulosa is commonly used in Australia for sediment quality assessment (Simpson et al. 2005; Strom et al. 2011). For nine contaminated sediments which caused varying degrees of toxicity to M. plumulosa in 10-d lethality tests (Table 4.2), the avoidance response after 48-h was determined (Figure 4.3).
Six of the contaminated sediments resulted in a significant avoidance response in M. plumulosa within 48 h. The greatest avoidance occurred for Sediments 1, 2 and 3 which resulted in 97%, 94% and 93% of M. plumulosa moving to the uncontaminated sediment, respectively. These sediments were highly toxic, with Sediment 1 and
Page | 104 Sediment 3 resulting in >75% 10-d lethality (Table 4.2). While it was not possible to accurately determine the 10-d lethality of the diesel-spiked Sediment 2 (perhaps due to the volatility of the diesel changing during the exposure), the sediment was strongly avoided. The exposure of M. plumulosa to sediments that were moderately toxic
(Sediments 5, 6 and 7) did not always result in a significant avoidance response. More than 75% of amphipods avoided Sediment 5 and Sediment 7, however avoidance was not significant for Sediment 6 despite resulting in >40% lethality in 10-d toxicity tests.
Sediment 5 and Sediment 7 both contained similar concentrations of copper but differed significantly in grain size distribution (Table 4.2). When M. plumulosa were exposed to sediments that did not elicit acute toxicity (Sediments 8 and 9), avoidance was not observed.
In all exposure chambers the overlying water was continuously exchanging between the two sides and dissolved metal concentrations in the overlying waters were relatively low (e.g. <30 µg Cu/L for tests with Sediments 5 and 7) and not significantly different for samples taken from the different sides of the test chamber (data not shown). As it is known that M. plumulosa is sensitive to sediment-bound copper (Gale et al. 2006; King et al. 2006a), the avoidance of these sediments was believed to be due to difference in particulate copper or the flux of copper at the sediment-water interface, rather than differences in dissolved copper in the overlying water.
Sediments 5 and 7 were copper-spiked sediments, and although they had been equilibrated and porewater copper concentrations were <5 µg/L (unpublished results), the copper in these sediments was likely to be more bioavailable than if present in field-collected sediments with similar copper concentrations (Strom et al. 2011). The
Page | 105 greater avoidance of Sediments 5 and 7 compared to Sediment 4 that had high lead and zinc concentrations and greater toxicity indicates that M. plumulosa may be more sensitive to copper than other metal contaminants. However, although dilute acid- extractable metal concentrations (Table 4.2) are expected to predict the bioavailability of metals better than total metal concentrations (Simpson and Spadaro 2011), the fact that Sediments 5 and 7 were copper-spiked may have resulted in this copper being easier for M. plumulosa to sense than the lead and zinc in the field-collected
Sediment 4.
Despite the high lead and zinc concentrations in Sediment 9, this sediment was not avoided and was also not toxic to M. plumulosa. This sediment had a relatively high concentration AVS (14 ± 12 µmol/g), and the lead and zinc are predicted to be present in sulfide phases of low bioavailability (U.S. Environmental Protection Authority 2005).
Interestingly, M. plumulosa appeared to show a slight attraction to this sediment with
61% of amphipods remaining on the seeded portion of the test vessel, effectively choosing to inhabit the contaminated sediment. However, there was no statistical significance between the distribution of amphipods on the seeded and unseeded chambers at the end of the test.
The experiments with uncontaminated sediments show that sediment grain size did not influence avoidance response for M. plumulosa, and this also appeared to be the case for the contaminated sediments. Marklevitz et al. (2008b) found that avoidance of contaminated harbour sediments by a mollusc was a balance between attraction to food and aversion to contamination. Given the high concentration of organic carbon present in Sediment 4 and Sediment 9 (6.8 and 6.0 %, respectively, which exceeds the
Page | 106 reference sediment), it is possible that the attraction to the added nutrition of the test sediments (presumable being provided by the addition organic content of the sediment) may have affected the observed avoidance of these sediments. Sediment 3 also had a high organic carbon content (5.9%), however the observation that it was both avoided and toxic (> 75% mortality) indicates that if nutrition in the form of TOC influences avoidance behaviour, then the quality of the TOC may also be important.
The major contaminant in Sediment 3 was copper and when considering the results for
Sediments 5 and 7, may indicate that sediment copper is more readily detected and avoided by the amphipod.
4.3.5 Influence of hazard magnitude on avoidance behaviour
The increasing hazard posed by contaminated sediments is thought to elicit a faster or greater response from exposed organisms (De Lange et al. 2006; Marklevitz et al.
2008b). For M. plumulosa, the results of the present study indicate a similar trend, as the degree of avoidance was generally greater for organisms exposed to sediment which elicited greater toxicity (Table 4.2; Figure 4.3).
This indicates that the hazard imposed by the contaminated sediment (i.e. toxicity) influences the number of amphipods that avoid the ‘unfavourable’ sediment over a
48-h period. The avoidance response of the copepods to Sediments 1 and 3 was very similar to the amphipod response (Figure 4.3), but over shorter exposure times of 3-h
(87 ± 4 % and 92 ± 5 %, respectively) and 24-h (83 ± 5 % and 82 ± 4 %, respectively).
Given the similar toxicity of these two sediment samples and the similarity between the avoidance response of the copepods and amphipods, the avoidance behaviour of
N. spinipes may also increase as toxicity increases. However, overall an insufficient
Page | 107
Figure 4.4 The observed avoidance response of Melita plumulosa following 48-h exposure to field contaminated and laboratory spiked sediments (mean ± SE, n = 3). * denotes a significant avoidance of contaminated sediment. 10-d whole sediment toxicity decreases from left to right. range of contaminated sediments were tested to determine if this was true for N. spinipes and P. solida.
Comparing the distribution of M. plumulosa and N. spinipes exposed to control conditions at each time point and the corresponding contaminant avoidance provides strong evidence that the presence of sediment-bound contaminants influences the observed avoidance response (Figure 4.3). In test chambers containing only control sediment, approximately 24 h was required for a significant percentage of the organisms to migrate away from the ‘seeded’ portion of the container compared to <6 h when contaminants were present. This was despite possible density-related pressures that may be placed on the organisms due to seeding on only one side of the
Page | 108 test container. It is particularly evident for N. spinipes where after 1-h only 15 ± 4 % of the copepods migrated away from the ‘seeded’ control treatment, compared to 89 ±
4 % of copepods avoiding Sediment 3 (to which they were ‘seeded’ onto). While the response was not as rapid or clear for M. plumulosa, the distribution after 6 h was highly variable for the controls, but 76 ± 4 % of amphipods had avoided the contaminated Sediment 5. The enhanced activity of M. plumulosa and N. spinipes was clearly avoidance behaviour and toxicity would have occurred if the species remained in the sediments.
The snail P. solida was slower moving, but more rapid avoidance was observed for
Sediment 6 (≤6 h) than for Sediment 7 (~ 16 h) (Figure 4.3). While Sediments 6 and 7 caused similar acute toxicity (Table 4.2), accompanying studies showed that Sediment
6 caused much greater chronic toxicity to M. plumulosa (unpublished data). In addition, the degree of avoidance exhibited by P. solida exposed to Sediment 6 for 6,
24 or 48-h was not significantly different (<7% difference in the number of avoiding organisms). This suggests that density-related pressures are not influencing the behaviour of P. solida, however tests were not undertaken to determine whether this was the case here. After 48 h the snail preferentially maintained a higher density in clean sediments rather than inhabiting the contaminated Sediment 6 and 7.
Previous studies have demonstrated behavioural changes in benthic organisms resulting from the presence of environmental contaminants. The bivalve
Macomona liliana was shown to crawl or drift away from sediments spiked with copper and zinc (Roper and Hickey 1994; Roper et al. 1995). They concluded that a behavioural endpoint for the bivalve was more sensitive than both burial/morbidity
Page | 109 and mortality endpoints (Roper and Hickey 1994). Physella columbiana, an aquatic pulmonate snail, demonstrated the ability to detect cadmium and zinc at concentrations of 9 µg Cd/L, 56 µg Zn/L and below 10 µg/L for mixtures (Lefcort et al.
2004). Oakden et al. (1984) report the avoidance of metal and sewerage-enriched sediments by phoxocephalid amphipods (Rhepoxynius spp.), while Kravitz et al. (1999) found that Eohaustorius estuarius could avoid sediments with moderate to heavy PAH contamination.
4.4 Conclusion
The present study demonstrated the ability of M. plumulosa, N. spinipes, and P. solida to respond to chemical cues in the environment and choose a habitat that provides the best opportunity for survival by avoiding contaminated sediment. The avoidance response of the test organisms indicates that static 10-d toxicity methods are likely to overestimate toxicity for species which would avoid contamination in heterogeneous field settings. The sensitivity of the avoidance response of each species to sediment- bound contaminants indicates that they may be suitable for developing as rapid screening methods to assess sediment quality. The study provided strong evidence that the avoidance response of M. plumulosa was related to the toxicity associated with the sediment. For this species the 48-h avoidance endpoint correctly identified as a potential hazard six of seven sediment samples that caused significant acute toxicity during 10-d exposures.
Page | 110 Chapter 5 Slow avoidance response to contaminated sediments elicits sublethal toxicity to benthic invertebrates
This chapter consists of a co-authored published paper. The bibliographic details of the co- authored published paper, including all authors, are:
Ward, D. J., Simpson, S. L. and Jolley, D. F. (2013). Slow avoidance response to contaminated sediments elicits sublethal toxicity to benthic invertebrates. Environmental Science and Technology 61(3): 414-425
My contribution to the published paper involved: » Initial concept and experimental design. » Collection, analysis and interpretation of the data. » Manuscript preparation.
______
Daniel J. Ward
______
Corresponding author of published paper: Stuart L. Simpson
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Supervisor: Dianne F. Jolley
Page | 111 5.1 Introduction
A major portion of the dissolved contaminants that enter aquatic systems bind with dissolved organic matter, inorganic colloids or suspended particulates and eventually deposit in bottom sediments (Linnik and Zubenko 2000). The distribution of contaminant concentrations in surface sediments is often heterogeneous (Harbison
1986; Lacerda et al. 1988). Variations in sediment contaminant concentrations may occur over large geographic regions (i.e. across large bays and estuaries) (Rakocinski et al. 1997; Chariton et al. 2010), or within much smaller and localised patches (<1 cm), often referred to as micro-niches (Stockdale et al. 2009). In addition to total concentrations, the speciation of contaminants may be even more heterogeneous due to bioturbation, resulting in mixing of redox layers that were created in a stratified manner, and deposition of organic matter (faeces/defunct remains) (Stockdale et al.
2009).
The heterogeneity of contaminated sedimentary environments influences the contaminant exposure for mobile benthic organisms. As invertebrates move throughout the environment (for foraging, mate searching, migration etc.), they may move in and out of sediments with high and low levels of contamination, thus resulting in intermittent exposures to contaminants. However, it is well recognised that many benthic organisms can detect and actively avoid areas of contamination (Lefcort et al.
2004; Lopes et al. 2004; Marklevitz et al. 2008a). Avoidance behaviour exhibited by mobile aquatic organisms is considered a significant factor determining the extent of exposure of an organism (Little et al. 1993; Weber 1997; Lefcort et al. 2004) and therefore the magnitude of the hazard and the overall risk the sediment poses to
Page | 112 ecosystem health. Thus a benthic organism’s exposure to contaminants may vary temporally and spatially due to the erratic nature of contaminant inputs.
Pulses of dissolved contaminants (both metal and organic) have the potential to cause toxic effects to aquatic organisms despite having short intermittent exposure times
(van der Hoeven and Gerritsen 1997; Brent and Herricks 1998), however, if the mode of action of the toxicant is reversible, recovery of the organisms between exposures to contaminant pulses may be possible (Kallander et al. 1997; Naddy and Klaine 2001).
While there has been significant research into the effect of dissolved contaminant pulses on aquatic life, little has been done to determine the impact of sediment heterogeneity on organism exposure to sediment-bound contaminants. Results of past studies have demonstrated the ability of benthic organisms to avoid contaminated sediment which suggests that exposure to contaminated sediment in the environment may not be continuous (Chapter 4; Kravitz et al. 1999). Instead, it is more likely that organisms will come into contact with contaminated sediment as short intermittent exposures similar to aquatic pulse exposures. Chapter 4 demonstrated that the rate of avoidance of contaminated sediments by an amphipod, harpacticoid copepod and a snail differed significantly (Ward et al. 2013). These species started avoiding contaminated sediments within 1 to 6 h, but sometimes as long as 48 h was required for a significant avoidance response to be observed. It was speculated that slow avoidance behaviour could result in toxicity to sensitive benthic invertebrates.
Bioassays used for sediment quality assessment typically rely on static continuous exposure of a test organism to a contaminant or contaminated sediment. Static bioassay methods which force a continuous exposure throughout the duration of the
Page | 113 experiment (often 10 days) will not suitably represent the nature of exposure that mobile benthic organisms have to the same contaminants in field locations. To further our understanding of the influence of exposure duration and frequency on toxicity caused by exposure to contaminated sediment, the acute and chronic toxicity associated with short, 'pulsed' exposures to contaminated sediment was assessed. This study investigated the potential toxic effects to two benthic invertebrates, an epibenthic amphipod (Melita plumulosa) and a harpacticoid copepod (Nitocra spinipes), associated with short intermittent exposures to contaminated sediment.
The effect of exposure duration and frequency on acute and chronic toxicity endpoints was assessed following exposure to four field-contaminated sediments.
5.2 Materials and Methods
5.2.1 General Chemistry.
New plastic ware was used for all chemical analyses. All chemicals were analytical reagent grade or equivalent analytical purity. Measurements of pH, salinity, temperature and dissolved oxygen were made in accordance with the instrument manufacturers’ instructions. Analyses of sediments included particle size (by wet sieving and gravimetry), organic carbon (high temperature TOC analyser), and particulate metals (2:1 concentrated HCl:HNO3, heated) (Spadaro et al. 2008; Strom et al. 2011). Overlying water samples were rapidly filtered through acid-washed 0.45 µm membrane filters (Minisart, Sartorius) immediately following collection and acidified to
2% HNO3 (v/v) with concentrated HNO3 (Tracepure, Merck). Acid-volatile sulfide (AVS) and simultaneously extracted metals (SEM) were analysed according to Simpson
(2001). Dissolved metal concentrations in water samples and digested sediments were
Page | 114 determined by inductively coupled plasma-atomic emission spectrometry (ICP-AES,
Varian 730-ES, Varian Australia) calibrated with matrix-matched standards (Angel et al.
2010). Analyses of filter and digest blanks, replicates for 20% of samples, analyte sample-spikes and the certified reference material (PACS-2, National Research Council
Canada, Ottawa, ON, Canada) were made as part of the quality assurance, and recoveries were within 85-110% of expected values. The limits of reporting for the various methods were less than one tenth of the lowest reported values. All sediment related concentrations are reported on a dry mass basis.
5.2.2 Test media.
Clean seawater was collected from Port Hacking, Sydney, Australia, membrane filtered
(0.45 µm), and acclimated to the room temperature of 21±1°C. Where necessary, the salinity of the filtered seawater was adjusted to the test salinity of 30 PSU using deionised water (18 MΩ/cm; Milli-Q Academic Water System).
Relatively clean silty sediments were collected as described previously from the
Bonnet Bay estuary, Port Hacking, Sydney, Australia shown to have low or negligible concentrations of metal and organic contaminants (Simpson and Spadaro 2011). The sediment was stored at 4oC for no longer than 1 month before use. Contaminated sediments were collected from estuarine field sites of unspecified locations, stored at
4oC in the dark, and toxicity testing undertaken within 8 weeks (Chariton et al. 2010;
Simpson and Spadaro 2011). Sediments from these locations had been used in earlier studies of sediment avoidance behaviour for these species, and in that study it was demonstrated that the differences in sediment properties (e.g. particle size, organic carbon (OC), AVS) would not result in avoidance behaviour (Chapter 4, Ward et al.
Page | 115 2013). Analyses of physicochemical properties (pH, OC, particle size, AVS) and metal contaminants were made on all sediments collected (Simpson et al. 2006). Previous analyses of sediments from these sites have consistently found that they contain negligible concentrations of common organochlorine or organophosphate pesticides
(0.005-0.05 mg/kg), polychlorinated biphenyl (PCB) aroclors (<0.01-0.1), (<250 mg/kg) or polycyclic aromatic hydrocarbons (PAHs), BETX (<0.25 mg/kg benzene, toluene, ethyl benzene, xylene), and total petroleum hydrocarbons (<1 mg/kg) (Chariton et al.
2010; Simpson and Spadaro 2011).
5.2.3 Test species.
Both species used in this study were epibenthic invertebrates found in intertidal estuarine environments of south eastern Australia. Melita plumulosa (Zeidler) is commonly found in estuarine tidal mudflats ranging from silty to sandy sediments in freshwater, estuarine and marine environments throughout south-eastern Australia
(Hyne et al. 2005). Adult specimens of M. plumulosa typically range from 8-10 mm in length. The harpacticoid copepod species Nitocra spinipes (Boeck) is known to adapt to a wide range of environmental conditions (including salinity and temperature) and as such has a world-wide distribution (Chapter 3, Ward et al. 2011). Mature copepods of this species are approximately 400 µm long. M. plumulosa and N. spinipes were obtained from laboratory cultures, maintained as described previously (Simpson and
Spadaro 2011; Chapter 3, Ward et al. 2011).
5.2.4 Toxicity test procedures.
Glass beakers and acrylic beaker-lids used for toxicity tests were cleaned in a dishwasher (Gallay Scientific Pty Ltd) programmed for a phosphate-free detergent
Page | 116 wash (Clean A, Gallay Scientific Pty Ltd), a dilute acid wash (1% HNO3), followed by thorough rinsing with Milli-Q water. New plastic ware was used for each copepod test performed.
The acute 10-day lethality tests with M. plumulosa were conducted as described previously (Spadaro et al. 2008; Strom et al. 2011). In brief, tests were performed at
21 ± 1 °C in a constant environmental chamber (Labec Refrigerated Cycling Incubator) on a 12-h light/ 12-h dark cycle (light intensity = 3.5 μmol photons/s/m2). Dissolved oxygen (>85%), pH (7.5-8.2), salinity (30 ± 1 PSU) and temperature (21 ± 1 °C) were monitored and maintained. The method was modified slightly to include three replicate 250 mL beakers containing 30 g of test sediment, 200 mL seawater and 15 adult M. plumulosa per treatment. This amount of sediment created a depth of 1-2 cm within the beaker and has been previously demonstrated to provide plenty of substrate and nutrition (>3% OC) for this shallow-burrowing species during 10-day acute lethality and sublethal tests (Spadaro et al. 2008; Simpson and Spadaro 2011;
Campana et al. 2012).
Every two to three days, 80% of the overlying water was replaced with clean seawater.
In all tests, water samples were collected before and after each water change, and at the end of tests for dissolved metal analysis. Although the source of the metals in the overlying waters was from the pore waters, previous studies have found that the overlying water concentrations provide a suitable level of information for interpreting the dissolved metal exposure for this epibenthic species. No food was added throughout the duration of the tests (Spadaro et al. 2008). At the termination of the tests, live organisms were sieved from the sediments and identified by movement and
Page | 117 the remaining sediment was fixed with neutral phosphate buffered formalin and stained with Rose Bengal solution. After 72 h, any surviving amphipods (now stained) that were missed initially were added to the count. Tests were considered acceptable if the physico-chemical parameters in beakers remained within the limits of pH 7.7-8.2, at 20-22 °C, dissolved oxygen >80% saturation, and salinity 28-32 PSU throughout the test, and if survival of amphipods was on average ≥80% in the controls.
5.2.5 Sublethal bioassays with pulsed contaminant exposure.
Sublethal toxicity tests that assess reproductive output of M. plumulosa and N. spinipes were modified from previously published methods for the purpose of achieving pulsed exposures to contaminated sediments (Mann et al. 2009; Perez-
Landa and Simpson 2011). The pulsed exposure durations were chosen based on avoidance times determined previously (Chapter 4, Ward et al. 2013). A visual presentation of the various pulse exposures are shown in Figure 5.1. Each pulsed exposure period was separated by a period of at least 48 h where the test organisms were in contact with uncontaminated control sediment.
Female M. plumulosa were isolated at least 7 days prior to commencing chronic assays and placed into trays containing uncontaminated sediment. This allowed any pre- existing embryos to develop and ensure that non-gravid females were used at the commencement of the test. An excess of males and females were placed into the sediment exposure chambers that had been specifically designed and constructed to allow for the quick and easy transfer of the amphipods, minimising disturbance to the organisms while allowing them to be in contact with the test sediment.
Page | 118 a)
3 × 24 h pulses 3 × 48 h pulses Continuous exposure Exposure
0 48 96 144 192 240 0 48 96 144 192 240 0 48 96 144 192 240 Time (h) Time (h) Time (h) 96% survival 14% survival 12% survival 1 × 48 h pulse 2 × 48 h pulses Exposure
0 48 96 144 192 240 0 48 96 144 192 240 Time (h) Time (h) 100% survival 75% survival b)
1 × 24 h pulses 1 × 48 h pulse Continuous exposure Exposure
0 48 96 144 192 240 0 48 96 144 192 240 0 48 96 144 192 240 Time (h) Time (h) Time (h)
> 60% survival 28% survival 0% survival
Figure 5.1 Visual representation of exposure scenarios used for a) amphipod and b) copepod bioassays. Acute toxicity (survival) is also shown for organisms exposed to Sediment 1 - green, yellow and red represent non-toxic, moderately toxic and toxic, respectively. See Appendix III for a summary of acute and chronic toxicity results.
Page | 119 The chambers were made from polycarbonate containers (diameter = 7 cm; height = 10 cm) with the base removed and a large hole (diameter = 6 cm) cut into the lid. A 250 µm mesh was secured over the base and lid of the container so that when pressed down onto the surface of the test sediment, the surface layer of the sediment could push through the mesh and overlying water could exchange with water outside of the chambers. The test vessels and sediments were placed into glass tanks containing filtered seawater, an aquarium pump and an air stone for the duration of the tests (see Appendix I for additional details). The chambers were carefully rinsed with clean seawater to remove all sediment particles when being transferred between sediments. Controls were handled in the same way as the treatments. Test chambers were set up with sediment and filtered seawater and equilibrated for 24 h before the initiation of the experiment. The overlying water was replaced immediately before the test organisms were added.
Separate exposures to contaminated sediments were undertaken for male and female amphipods to provide information on the effect of gender. Following a 24- or 48-h exposure period, 7 male and 5 female amphipods were randomly selected and transferred to beakers containing uncontaminated sediment and the test continued under the same conditions as used for the acute toxicity tests. The sublethal assays were terminated after 10-days. Juveniles present were sieved from the sediment and counted. Females were inspected under a light microscope to count the number of eggs/embryos being carried in the brood pouch. In the case that amphipods were still amplexed when the test was terminated, the pair were placed in a beaker containing clean seawater and returned to the incubator for a further 48-h before the number of
Page | 120 embryos were counted on the female. The test endpoint was ‘total offspring’ which included the juveniles and the embryos produced from the mating of the amphipods exposed to the test sediments.
For the experiment on the copepod N. spinipes, gravid females were selected from the cultures and placed into a series of 10 ml polycarbonate tubes (maximum of 20 females for tube) containing test sediment. The females were exposed to the test sediment for the desired period and then gently removed from the tubes (using a pasture pipette) and pooled in a Petri dish containing clean seawater. Five females were randomly selected from the pool and placed into each of a series of 10 mL polycarbonate tubes (5 replicates per treatment) containing uncontaminated sediment for 10 days. A small amount of food (150 µl vial from a stock of 1 × 104 cells/ml of both
Isochrysis sp. and Tetraselmis sp. and 0.3 mg Sera Micron fish food sieved to <63 µm) was added to each vial at the beginning of the exposure and when copepods were transferred to vials with uncontaminated sediment. In addition, a small volume of food was added to the vials following water changes every 2 to 3 days during the assay. This species is iteroparous, meaning that females are capable of producing multiple broods of offspring from reserves of sperm stored after mating, giving this species the ability to produce multiple broods in a short period of time (Bengtsson 1978). This characteristic makes N. spinipes a suitable species to assess sediment toxicity using a reproductive output as a chronic endpoint (Chapter 3, Ward et al. 2011). The test endpoint was ‘total offspring’ which included all nauplii and copepodites.
Page | 121 5.2.6 Amphipod gender-exposure tests.
Trays containing uncontaminated sediment and either solely male or female M. plumulosa were set up in the laboratory at least 7 days prior to commencing these tests. This was considered sufficient time for fertilised or gravid females to drop young and become non-gravid. Individuals of these males and non-gravid females were randomly selected and separately exposed to contaminated sediment for 48-h. By separating and exposing male and female M. plumulosa to contaminated sediment, the mating of exposed and unexposed organisms over a 10-d period could be controlled. Following the 48-h exposure period, 7 males and 5 females were placed into a 250 ml beaker containing uncontaminated sediment for 10 days. Four reproduction scenarios were investigated as follows: a) unexposed males × unexposed females, b) exposed males × unexposed females, c) unexposed males × exposed females, and d) exposed males × exposed females. Microscopy was used to determine the number of offspring produced per female during the last 10 days of the bioassay.
5.2.7 Data Analyses.
Descriptive statistics were generated using the Microsoft Excel (2007) data analysis tool pack. Survival and reproductive output of test organisms in bioassays are reported as a percentage relative to controls. All statistical analyses were performed using
Toxcalc for Microsoft Excel (TidePool Scientific Software). Survival and reproductive output data, expressed as percent control, was tested for normality (using the Shapiro-
Wilk’s test) and homogeneity of variance (Bartlett’s test) prior to analysis.
The t-tests were performed to assess a significant reduction in survival and reproduction of amphipods and copepods exposed to contaminated sediments
Page | 122 compared to controls. Dunnet’s test was then used if assumptions of normality and homogeneity of variances were met. Where nonparametric analysis was required,
Steel’s test was used. For toxicity tests on single sediments, t-tests were used to determine significant differences in the response of the amphipods and copepods exposed to test sediment compared to that in the control sediment. Significance in all statistical tests was set at the p < 0.05.
5.3 Results and Discussion
The properties of the contaminated sediments are shown in Table 5.1. The AVS concentrations were <5 µmol/g in all sediments and were consistent with them being predominantly oxic/sub-oxic in nature. While the metal concentrations in these sediments were considered very high, both test species demonstrated sensitivity to metal contaminants, with sublethal effects occurring when dilute acid-extractable metal concentrations exceed lower guideline values (Simpson and Spadaro 2011).
Dissolved metal concentrations in the overlying water of M. plumulosa bioassays remained well below reported 10-d and 96-h LC50 values for both adult and juvenile amphipod survival (King et al. 2006b). Concentrations of dissolved Cu, As, Cd and Pb measured in the overlying water did not exceed 40 µg/L in any bioassay, and dissolved zinc reached a maximum of 280 µg/L. Because of the small volume of seawater used in the copepod bioassays, it was not possible to subsample the overlying water during the assays to analyse dissolved metal concentrations. It is assumed that dissolved metal concentrations were similar to those measured in the amphipod tests and this species has been found to have a similar sensitivity to metals as the amphipod (Perez-
Landa and Simpson 2011; Simpson and Spadaro 2011).
Page | 123 Table 5.1 Properties of control and contaminated sediments used in sediment pulse experiments
Test sediment Parameters a Control 1 2 3 4
Silt, % 98 76 15 79 27
AVS, µmol/g 4.5 <5 <5 <5 <5
TOC, % 4.7 3.9 5.9 6.8 3.0
As, µg/g 4 3430 7 61 ND
Cd, µg/g 0 83 0 19 1.1
Cu, µg/g 13 1100 1070 108 71
Pb, µg/g 40 14400 60 830 470
Zn, µg/g 210 14500 260 2630 1230
Static 10-d amphipod survival 100±2 12±5 24±6 32±2 59±11
48-h exposure amphipod survival 100±3 100±5 93±2 96±2 ND
a Silt = percent of particle <63 µm. AVS = acid-volatile sulfide. TOC = total organic carbon. All metal concentrations are 1-M HCl extractable metals. Amphipod survival expressed as % Control ± SE. ND = not determined.
5.3.2 10-day lethality from continuous or pulsed exposures to contaminated sediment
For 10-day lethality tests with continuous exposures, all contaminated sediments were highly toxic to M. plumulosa, with survival of 12±5 %, 24±6 %, 32±2 % and 59±11 %
(relative to controls) for Sediments 1, 2, 3 and 4, respectively (Table 5.1; also see
Appendix II). In Chapter 4, it was demonstrated that in some cases M. plumulosa has the ability to avoid contaminated sediment within 6-h of the initial exposure and all
Page | 124 toxic sediments were avoided within 48 h. Therefore, a 48-h exposure period was used to represent a ‘worst case’ scenario for exposure to contaminated sediments. In all experiments, the survival in control treatments was >95%. Control amphipods were transferred from clean sediment to clean sediment to mimic the handling stress that may occur when creating the pulsed exposures.
When exposed to Sediments 1, 2 and 3 for 48-h during the 10-day assay (remaining time in clean sediments), the survival was 100±5 %, 93±2 % and 96±2 %, respectively, and not significantly different from controls. Thus, as M. plumulosa would avoid these contaminated sediments within 48 h, acute lethality would not be expected to occur if cleaner sediments are located suitably close to where this species can move. It can also be concluded that latency effects did not result from the 48-h exposure, as surviving amphipods were still alive after eight days in the clean sediment.
To investigate the effect of increased exposure frequency and duration on toxicity, the survival of M. plumulosa was determined for various exposure scenarios (Figure 5.1 and Figure 5.2). This series of experiments used Sediment 1 as it resulted in the greatest toxic effect in the continuous 10-d exposure (Table 5.1). Significantly lower survival (75±8 %) was observed following two 48-h exposure periods (total exposure time of 96 h) with a period of 48-h of no-exposure between exposure pulses
(Figure 5.2). For three 48-h exposure pulses (total exposure time of 144 h during the
10-day test) much greater mortality occurred (14±5 % survival) and was comparable to the toxicity observed for continuous exposure to Sediment 1.
To consider the influence of exposure duration on toxicity to M. plumulosa, the exposure period was reduced to 24-h. Survival of M. plumulosa exposed to Sediment 1
Page | 125
Figure 5.2 The effect of increasing periods of exposure to contaminated Sediment 1 on survival and reproduction of M. plumulosa. The shaded region represents duration of exposure required to elicit contaminant avoidance (from Chapter 4, Ward et al. 2013). The error bars represent standard error (n=4). for three 24-h exposure pulses did not decrease significantly in comparison to controls
(96±5 % survival; total exposure time of 72-h, with each exposure pulse separated by a
72-h period of exposure to uncontaminated sediment). This confirmed that the duration of exposure is a significant factor affecting the toxic response of organisms that come into contact with contaminated sediment.
In the case of N. spinipes, significant toxicity occurred following a 48-h exposure to
Sediment 1 with survival reduced to 28±2 % (relative to controls). When the exposure time was reduced to a 24-h duration, survival remained significantly lower than the control, however remained higher than the 48-h treatment (24-h survival >60% relative to controls). This is a clear indication of the significance of exposure duration
Page | 126 in the expression of acute toxicity caused through the exposure to contaminated sediment. However, this species was demonstrated to have a fast avoidance response time, being able to avoid contaminated sediment within 1-6 h of exposure (Chapter 4;
Ward et al. 2013). Again, this would imply that acute toxicity could be avoided if clean uncontaminated habitat quickly be found.
The lethality tests conducted in the present study suggest that acute lethality from exposure to contaminated sediment may not occur in mobile benthic invertebrates if there is a possibility to escape the contaminant and inhabit uncontaminated sediment.
The lethality observed in the contaminated sediments was clearly dependent on the extent of exposure an organism had to the contaminated sediment (Figure 5.2). The extent of the exposure varies considerably between the traditional continuous exposure commonly used for acute and chronic toxicity methods, which may not reflect the true nature of exposure for mobile benthic invertebrates (Greenstein et al.
2008; Kennedy et al. 2009; Simpson and Spadaro 2011). Under most scenarios, where non-toxic sediments exist within a suitable distance, M. plumulosa and N. spinipes would likely avoid contaminated sediments and lethality to adults could potentially be avoided. However, the research has also confirmed that the exposure duration is a major factor contributing to the toxic effects elicited by sediment contamination and that the species may not adequately recover during periods in clean sediment between multiple exposures to contaminated sediment.
Page | 127 5.3.3 Sublethal toxicity from pulsed exposure to contaminated sediment
For the amphipod, M. plumulosa, and copepod, N. spinipes, effects to reproduction are known to occur in sediments that do not effect survival of adults (Mann et al. 2010;
Perez-Landa and Simpson 2011; Simpson and Spadaro 2011). Both of these organisms have shown the ability to avoid contaminated sediment within as little as 6 hours of exposure (Chapter 4; Ward et al. 2013). Based on the results presented above, the effective avoidance of contaminated sediment by M. plumulosa and N. spinipes may alleviate acute toxicity, however it is speculated that sublethal effects may occur as a result of short exposures to contaminants.
Sublethal effects were assessed when M. plumulosa was exposed to contaminated
Sediments 1 and 4 which caused 88% and 41% lethality during 10-day continuous exposure, respectively (Table 5.1, Figure 5.3). A 48-h exposure pulse to Sediment 4 resulted in a 46% decrease in reproductive output from 6.7±0.8 to 3.6±1.2 juveniles per female in control and contaminated treatments, respectively. The 48-h exposure to Sediment 1 resulted in the reproductive output decreasing from 6.7±0.8 to 2.4±0.6 juveniles per female in control and contaminated treatments, respectively, (a decrease of 65%). Increasing the time of exposure to Sediment 1 increased the magnitude of the toxicity exhibited by M. plumulosa, with reproductive output being 85±15%,
66±3% and 34±8% (of control) for 8-, 16- and 48- h exposures, respectively
(Figure 5.1 – the shaded area represents the time period where avoidance of
Page | 128 120 10-d survival Total offspring 100
80
60
% control % 40
20
0 Control Sed. 1 Sed. 4
Figure 5.3 The effect of short (48-h) exposure to contaminated sediment on the survival and reproductive output of M. plumulosa exposed to uncontaminated sediment (control) and field- collected contaminated sediments (Sediments 1 and 4; mean ±SE, n=4). contaminated sediment by M. plumulosa is likely to occur (Chapter 4; Ward et al.
2013)). Despite not being acutely toxic, sublethal effects to M. plumulosa were significant following the 16- and 48- h exposures (p<0.05).
For the benthic copepod N. spinipes, exposure to Sediment 1 and 2 for 24- and 48-h pulses, respectively, significantly lowered reproductive output after 10 days
(Figure 5.4). Significant lethality occurred when gravid copepod females were exposed to Sediment 1 for 48-h (27 ± 2% survival, mean ± SE, n = 4) and the exposure time was reduced to 24 h for subsequent tests. Following a single 24-h exposure to Sediment 1, the reproductive output was 16% (reduced from 10±0.8 to 2±0.4 juveniles per female in control and Sediment 1 exposed treatments, respectively; Figure 5.4). Nauplii and copepodite numbers declined by 82% and 95%, respectively. Following a single 48-h
Page | 129 90 24-h exposure to Sed. 1 80 48-h exposure to Sed. 2 70 60 50 40 30 20 10 Reproductive output (% control) (% output Reproductive 0 Nauplii Copepodites Total Offspring
Figure 5.4 The effect of short exposure to contaminated sediment on the reproductive output of N. spinipes (mean±SE ) exposed to Sediment 1 for 24-h (n=4) and Sediment 2 for 48-h (n=5). Total offspring was counted after a 10-d period in uncontaminated sediment following prior exposure to contaminated sediment. exposure to Sediment 2, the reproductive output was 63% and nauplii and copepodite numbers declined by 31% and 49%, respectively.
5.3.4 Gender of M. plumulosa influencing reproduction
For M. plumulosa, effects on reproductive output were observed following a single 16- h pulse exposure to contaminated field sediment. While the identification of this sublethal effect is ecologically important in implicating significant ecotoxicological consequences at the population level, the gender responsible for limiting the reproductive output of the population has not been identified.
The influence of the amphipod gender being exposed to contaminated sediment prior to mating was investigated. It was found that fecundity was significantly lower in treatments where females had been exposed to Sediment 1 for 48-h prior to mating.
Page | 130 When unexposed males were mated with exposed females, reproductive output over
10-d was 34% of controls. In the scenario where exposed males were mated with exposed females, fecundity was 63% of controls (Figure 5.5). Conversely, when exposed males were mated with unexposed females there was a slight increase in offspring production (12%) although this was not significantly different from control treatments. These findings are consistent with field population studies which found that female M. plumulosa collected from polluted sites along the east-coast of
Australia were less fecund than those obtained from uncontaminated sites (Chung et al. 2008). It is important to note that there is no statistically significant difference
(p>0.05) in the reproductive output observed between scenarios A and B or between C and D (Figure 5.5).
Interestingly, reproductive output was slightly increased in scenarios where male amphipods were exposed to contaminated sediment compared to the corresponding scenario with similar females. McCurdy et al. (2000) found that parasitised male C. volutator were more likely to mate immediately after being introduced to a receptive female when compared to unparasitised males. The slight increase in offspring production where only males were exposed may be the result of similar factors that increase the male drive to reproduce. McCurdy et al. (2000) also observed that pairings of parasitised males resulted in higher initial brood sizes. In a similar study, it was reported that females showing a greater degree of infection by parasites are less likely to invest in reproduction (McCurdy et al. 1999). It has also been found that mating in the isopod Lirceus fontinalis is ultimately controlled by the female of this species (Sparkes et al. 2000). As for amphipods, pair-formation is controlled by the
Page | 131 female as they must position themselves correctly to allow amplexing to occur
(Borowsky 1991). In light of this, it is concluded that contaminant exposure in female
M. plumulosa results in a physiological or behavioural change that is responsible for reducing its reproductive output. This could be due to the resorption/abortion of eggs and/or embryos (McCurdy et al. 1999), or non-cooperative behaviour which limits the ability of male M. plumulosa to amplex with receptive females (Borowsky 1991).
These results are also reflected in field studies of this species, with adult female M. plumulosa living in contaminated habitats found to be generally smaller and less fecund (Chung et al. 2008).
120
100
80
60
40
20 Juveniles per female (% (% control) female per Juveniles
0 Unexposed females Exposed males Unexposed males Exposed males Unexposed males Unexposed females Exposed females Exposed females
A B C D Scenario
Figure 5.5 Reproductive success observed for four gender-exposure scenarios generated by exposing male and female M. plumulosa to contaminated Sediment 1 (mean ±SE, n=4). A reduction in reproductive success was found to occur in treatments where females had been exposed to contaminated sediment prior to mating
Page | 132 5.4 Implications
Lethality associated with contaminated sediment resulting from 10-d static toxicity test methods was alleviated when the organism exposure was limited to short and intermittent exposures (Appendix III). This result suggests that the toxic effect resulting from contaminant exposure is determined by the frequency and duration of exposure to contaminants. M. plumulosa and N. spinipes are known to avoid contaminated sediment within 6 to 24-h and 1 to 6-h, respectively (Chapter 4; Ward et al. 2013). These findings indicate that acute toxicity may not occur for these species if suitable uncontaminated habitat can be found within 48-h of exposure to contaminated sediment.
The present study also confirmed that while latent effects of short pulses may not be observed in terms of lethal effects, sublethal effects to reproduction may occur as a result of exposure before M. plumulosa or N. spinipes sense and avoid contaminated sediment. It is also speculated that exposure of M. plumulosa to contaminated sediment causes a physiological change in females which reduces fecundity. This has significant implications for the design of sediment bioassays and the use of acute toxicity tests for sediment quality assessment purposes which may over-estimate the toxicity associated with sediment in a field location. Specifically, these results imply that traditional standard toxicity tests which employ static continuous exposure methods do not provide an assessment of the impact contaminated sediments have on aquatic ecosystems at the population level.
Page | 133 Chapter 6 Conclusions
This study assessed the response of marine benthic fauna to continuous and pulsed exposures of contaminated sediment. This was achieved by i) identifying suitable test species, ii) understanding the avoidance response exhibited by the chosen test species and iii) assessing the effect of continuous and pulsed exposure to contaminated sediment on mortality and reproduction. The high sensitivity of chronic and behavioural endpoints highlighted the importance of considering the environmental relevance of the methods used to assess sediment toxicity.
There are numerous published studies detailing methods for assessing the toxicity of contaminated sediments to benthic invertebrates (Mann et al. 2010; Perez-Landa and
Simpson 2011; Simpson and Spadaro 2011; Ho et al. 2013; Molisani et al. 2013). While the majority of studies utilise an acute toxicity endpoint (e.g., survival) in static bioassays, in recent years there has been increasing interest and emphasis on utilising more environmentally relevant test endpoints. This has led to the development of a range of methods which allow the assessment of chronic toxicity (e.g. reproduction, growth, behaviour) following exposure to contaminated sediments.
The importance of obtaining effects data for species that have differing life cycles, behaviours (e.g. burrowing and feeding) and exposure pathways is well recognised.
This is reflected in the development of legislative guidelines, where sediment quality guideline development typically requires biological effects data for at least three species per trophic level from three different levels. Consequently, there is an inherent need for the ongoing development of routine toxicity test methods which utilise a variety of aquatic organisms.
Page | 134 Due to the dynamic nature of contaminant inputs, the distribution of contaminants is seldom homogeneous in aquatic systems. Often, a high concentration of contaminants occurs at the site of the contamination source, resulting in a plume in the water column that gradually dilutes as it mixes within the body of water. This dilution causes a concentration gradient which varies over time as well as distance from the contamination source. The effect of temporal and spatial variations in dissolved organic contaminant concentrations has received increasing attention in the literature.
However, dissolved organic contaminants such as pesticides, insecticides and fertilizers remain the focus of contaminant pulse research. Sediment-bound contaminant concentrations are generally considered to be more stable and show less temporal and spatial heterogeneity compared to dissolved contaminants present in the water column. However, studies documenting the distribution of contaminants in surface sediments indicate that this is not the case. Instead, both organic and inorganic contaminant concentrations in surface sediments vary across large bays and estuaries.
Bioturbation and redox chemistry can also influence the partitioning of contaminants between sediment particles and sediment porewater, influencing the vertical distribution of contaminants below the sediment-water interface.
The recent development and application of diffusive gradients in thin films (DGT) has shown that variation in contaminant concentrations in sediment environments exists within an area of 1 cm2 (termed micro-niches), making them particularly heterogeneous. The heterogeneous nature of environmental contaminants makes it likely that mobile benthic organisms will repeatedly come into contact with patches of contaminated sediment while moving throughout the environment to forage for food,
Page | 135 search for mates, or through natural migration. This intermittent organism exposure to contaminated sediment and any potential effects have not been adequately addressed in the literature. This intermittent exposure is further complicated by the avoidance behaviour exhibited by many benthic invertebrate species. The ability of some species to detect and avoid contaminated sediment will influence the time of exposure the organisms have to contaminated sediment. Hence, the slower the avoidance response, the longer the duration of exposure and the more likely that toxic effects will result.
In this study, harpacticoid copepod species were isolated from local environments, identified and evaluated in terms of their suitability for developing rapid sublethal bioassays and latter project components (Chapter 3). While harpacticoid copepods have been used for assessing sediment toxicity in both North America and Europe, and standardised methods are now available for one North American species, the species had not been used for assessments of sediment quality in Australia. Furthermore, no studies had been reported on the ability of copepods to sense and avoid contaminated sediments. Four copepod species obtained from local estuarine sediments (Woronora
River and Port Hacking Estuary, Sydney, Australia) were studied to determine which species had the greatest suitability as a test species with the potential for the development of chronic toxicity test procedures.
Nitocra spinipes was determined to have the greatest potential to be a model test species for ecotoxicology when compared to Robertgurneya hopkinsi, Halectinesoma sp. and Tisbe tenuimana. N. spinipes was easy to culture and handle in the laboratory and have a sensitivity to contaminants that is comparable to the amphipod Melita plumulosa, which is routinely used for sediment ecotoxicology. Therefore, N. spinipes
Page | 136 is considered to be an appropriate candidate species for use in ecotoxicologal assessments of sediment from tempertate coastal environments. This copepod species was also demonstrated to be suitable for rapid assessment of sediment-metal induced effects on reproduction, and this endpoint was utilised in the avoidance components of this study. In addition, the methods presented in this study have commercial benefits as the test is rapid (10-d chronic test with reproduction endpoint), uses only small volumes of sample and has an easily identifiable endpoint, making this bioassay easy to use and less labour intensive than other chronic toxicity methods proposed for various aquatic organisms.
This research also investigated the ability of an amphipod, copepod and snail
(M. plumulosa, N. spinipes and P. solida) to respond to chemical cues and avoid contaminated marine sediments (Chapter 4). This study clearly demonstrated that benthic invertebrates are sensitive to and can detect/respond to chemical cues present in their surrounding environment. Each of the three test species demonstrated the ability to migrate away from contaminated sediment within 48 h of exposure to inhabit cleaner more suitable sediment. Validation of these results was undertaken with M. plumulosa, and suggested that the observed response was indicative of the toxicity associated with the specific sediment, with a greater number of organisms avoiding sediment eliciting higher toxicity. This study demonstrated that sediment avoidance is an effective and rapid screening method with the potential to identify contaminated sediments which may threaten the healthy functioning of an ecosystem.
The implications of contaminant avoidance on the toxic effects elicited to M. plumulosa and N. spinipes were also explored (Chapter 5). Experiments were designed
Page | 137 to assess both the rate at which these organisms migrated across heterogeneously contaminated sediments, and used pulsed contaminated-sediment exposure bioassays to assess the likelihood of effects due to slow avoidance responses. For sediments that caused significant acute lethality during 10-day exposures, short exposures caused insignificant acute lethality but resulted in sublethal effects to the reproduction of M. plumulosa. Similarly, for sediment that was acutely toxic for a 10-day exposure, a 48-h exposure resulted in a significant decrease in the 10-day reproductive output of N. spinipes. Overall, this study demonstrated that benthic organisms such as M. plumulosa, N. spinipes and P. solida have the ability to detect and avoid contaminated sediment. However, the significant sublethal effects observed following 16- and 48-h exposures to contaminated sediment for M. plumulosa and N. spinipes, respectively, must be considered in relation to their avoidance behaviour. The results of this study suggest that short exposures to contaminated sediment may not result in acute toxicity to benthic organisms. This is significant if organisms are able to avoid contaminated sediment and find more suitable sediment to inhabit in a suitable period of time, which has been demonstrated by this research and in the literature for a number of aquatic organisms. For M. plumulosa and N. spinipes it has been shown that avoidance of contaminated sediment can occur within 6-24 h and 1-6 h, respectively.
In the case of M. plumulosa, acute toxicity would not be likely to occur in the natural environment if M. plumulosa is able to move away from an area of contamination and inhabit clean sediment.
While the results of Chapter 4 suggested that avoidance of contaminated sediment (as exhibited by mobile benthic organisms) can reduce or eliminate acute toxicity
Page | 138 associated with contaminated sediment, it was unclear whether sublethal effects could occur due to the intermittent exposures. Thus the influence of repeated exposure to a contaminant, even at short intervals, on sublethal effects to reproduction of M. plumulosa was investigated (Chapter 5). A significant reduction in the reproductive output of aquatic organisms may result in profound population level effects, which may threaten the survival of a species in a particular ecosystem. The loss of a species from an ecosystem (whether due to avoidance of uninhabitable sediment or through mortality due to contaminant exposure) could potentially lead to significant ecological impacts, especially if the species is a keystone predator, has an important link in the food chain or a role in nutrient cycling.
The findings presented in this thesis have implications for current approach for sediment quality assessments used in ecotoxicology. Clearly traditional static toxicity tests may over-estimate the acute toxicity associated with a sediment sample. This occurs due to the confinement of the test species to the contaminated test sediment used in the bioassay forcing a continuous exposure to the contaminants present. As many benthic invertebrates commonly used as test species can detect and avoid contaminants, a lengthy continuous exposure is not likely to occur under real environmental conditions, which raises questions about the environmental relevance of such testing methods. However, it is noted that preventing organisms from avoiding unfavourable sediment in bioassays has merit as a precautionary assessment of maximum possible effects resulting from exposures, which is the aim of most assessment frameworks.
Page | 139 In an attempt to move towards more environmentally relevant testing, the last decade has seen increasing importance placed on the use of chronic toxicity endpoints. As chronic endpoints are considered to be more sensitive than acute endpoints, they are a better indicator of environmental contamination by identifying potential environmental risk at lower concentrations than contaminant exposures that result in loss of species. The increased sensitivity of chronic endpoints was observed in this current study, when decreased reproduction occurred after short exposures to contaminated sediment, despite no observable impact the survival of M. plumulosa or
N. spinipes occurring.
The significance of this finding becomes clear when this result is considered in relation to contaminant avoidance. For M.plumulosa, N. spinipes and P. solida, a rapid avoidance response was observed (generally ≤ 8 h), which under optimal conditions
(i.e. unimpeded access to clean habitable sediment) would ameliorate acute toxicity and reduce chronic impacts resulting from the contamination.
Contaminant avoidance is highlighted as a sensitive endpoint. Therefore, avoidance bioassays provide a rapid method of screening sediments to identify potential harm which may result from longer exposures. However, it is important to put these results into context with real-world environmental exposures. Given the dynamic nature of the environment, it is unlikely that organisms will be exposed to a single exposure of contamination. It is more likely that mobile benthic species will be repeatedly exposed to patches of contamination as they move across the sediment surface, with the exposure varying with the severity of contamination, contaminants present, and avoidance response of the species. Our research suggests that organisms do not
Page | 140 recover between exposures (observed for M. plumulosa) which leads to lethality as the total accumulated exposure time increases (Chapter 5). So if avoidance is slow, or is not possible due to other confounding environmental factors, it is likely that toxic effects will result (Figure 6.1). For example, the high sensitivity of reproductive output was demonstrated in this study with a significant reduction in offspring of M. plumulosa following a single 16-h exposure. It was proposed that this reduction in
Figure 6.1 The cumulative effect of short exposures to contaminated sediment causes a significant reduction in survival of M. plumulosa. Similarly, as the total exposure time increases, offspring production decreases. Significant chronic effects occur following shorter exposure times than those that cause significant mortality. This indicates contaminant avoidance may prevent acute toxicity from occurring but may result in chronic effects if clean habitat cannot be found within a sufficient period of time.
Page | 141 reproductive output was the result of a physiological change in the female amphipods following exposure to contaminants.
The research findings presented in this thesis push the boundaries of sediment ecotoxicology and traditional bioassay methods. Current standard methods used for regulatory purposes utilise static toxicity testing with a heavy reliance on acute toxicity endpoints. The conditions of these exposure methods only represent scenarios where there is sufficient contamination with no available refuge. In this study, we have shown how these methods fail to consider the dynamic nature of aquatic sediments which leads to an over estimation of toxicity, requiring costly and time consuming input from stakeholders and regulators alike.
This research highlights the need for a more holistic approach to sediment quality assessment, with an emphasis placed on the environmental relevance of the methods used for conducting ecotoxicological studies. In order to develop more robust guidelines which both protect the environment and accurately assess the harm of contaminant inputs, it is crucial that standard methods are introduced which assess the behaviour and potential chronic toxicity caused by exposure of model test species to sediments from contaminated sites. Integration of current guidelines with new techniques which assess a wide range of impacts to ecosystems and the organisms that inhabit them, will ensure that the sediment guidelines put in place are progressive, provide environmental protection based on cutting edge ecotoxicology research and allow for a greater understanding of the potential harm caused by environmental contaminants.
Page | 142 6.2 Further research
With an ever growing list of chemicals being manufactured, their widespread use makes it almost inevitable that contamination of waterways and ultimately aquatic sediments will occur. It would be a near impossible task to assess the infinite number of ‘exposure scenarios’ that exist as a result of sediment heterogeneity for one of these contaminants let alone all contaminants and chemical species. However, this study has shed new light on the way in which benthic organisms react to sediment contamination, and then in turn, the consequences associated with the exposure resulting from the change in their behaviour. As the approach of ‘pulsed’ sediment exposures is a novel concept, there is little research available from similar studies conducted on different organisms.
Evidence presented in this thesis indicates that Melita plumulosa did not recover in the
24 to 48 h period between exposure to the contaminated test sediment. This is also reflected in the results presented in the gender-exposure experiments which clearly showed exposed females as less fecund than unexposed females. Similarly, the results of the pulsed exposure experiments suggest that for M. plumulosa, sediment toxicity increases as the accumulated exposure time increases. This relationship supports the inference that M. plumulosa do not recover from contaminant exposure. It is noted that specifically testing the ability of these organisms to detoxify contaminants and the resulting recovery was beyond the scope of this project. In saying this, it would be of interest to allow longer periods of time between exposures to identify the ability of organisms to recover from the effects of contamination. A more detailed study investigating the potential recovery of benthic invertebrates following exposure to
Page | 143 contaminated sediment would help indicate the time required for organisms to detoxify contaminants and return to a ‘healthy’ state. By extending the time between exposure periods, amphipods may be able to recover from any toxic effects, meaning that subsequent exposures may not result in toxicity. Further research to identify possible correlations exist between the toxicity caused by different scenarios which have the same total exposure time (e.g. 4 × 24 h exposures/2 × 48 h exposures/1 × 48h and 2 × 24 h exposures) would also help identify the effectiveness of detoxification mechanisms of benthic organisms.
In reference to sediment avoidance, experiments conducted in this study gave the test organisms a choice of ‘clean’ control sediment and the contaminated test sediment.
The results of these tests clearly showed that avoidance was a measurable response with the capability to identify toxic sediments without resulting in significant mortality.
In completing this study, three key factors have been identified for further research which will further our understanding of contaminant avoidance by benthic invertebrates and the factors influencing the migration of test organisms:
Food availability and nutrition – All test sediments used in this study had comparable organic carbon content to that of the control sediment. This suggests that these sediments offered similar nutritional value to the test organisms. By increasing the nutritional value of the contaminated sediment, the trade-off between avoidance of contaminated sediment and food availability could be investigated.
Shelter – Benthic organisms often seek out shelter in the environment natural environment. In the artificial environment of the bioassay chamber, there are far fewer places available for the organisms to seek shelter. It is unclear whether or not
Page | 144 the provision of shelter on the contaminated sediment would entice the organisms to leave the open area offered by control sediment for the protection offered by the contaminated sediment.
Varying combinations of contaminated sediment – In this study, organisms were given the choice to inhabit either ‘clean’ reference sediment or a field collected contaminated sediment. However, the response of organisms in multiple choice experiments using a combination of contaminated sediments was not investigated.
Would benthic organisms avoid the most contaminated or toxic sediment if given the choice of two contaminated sediments? Would all test species migrate to the reference sediment and avoid all contaminated sediments, or would they avoid only the most toxic samples and be distributed across the less toxic sediments available? A more detailed study could be designed to determine if aquatic organisms have the ability to seek out the ‘cleanest’ habitat which poses the lowest threat of contaminant exposure and consequent toxic effects.
This research is intended to provide a starting point for the development of new approaches to sediment quality assessment. The findings reported in this thesis highlight several issues relating to the environmental relevance of current methods of sediment quality assessment. These results provide a foundation which can be built on to provide a greater understanding of the impact of sediment contaminants and the role of behavioural ecotoxicology in risk assessments.
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Page | 174 Appendix I
Further details of the sediment pulse experimental design
Test vessels for M. plumulosa exposure experiments were constructed from 250 mL polycarbonate containers with a screw lid (diameter = 7 cm; height = 10 cm). The base of the container was removed and a hole was cut into the top of the lid (approximately
6 cm diameter). Nylon mesh (250 µm mesh size) was stuck to the lid and the base of the container with aquarium safe silicone (Selleys®). The silicone was allowed to cure for 4 days (as per manufacturer recommendations) and containers were rinsed thoroughly with deionised water before use. Toxicity of all materials used in the construction of the vessels was checked prior to use in bioassays. Sediment was placed in a disk with a diameter slightly larger than the base of the 250 mL polycarbonate container. The test vessels were pressed down onto the surface of the sediment discs so that the test organisms inside the vessels were in contact with the sediment surface through the mesh. Test vessels were housed in glass tanks measuring approximately
40 × 40 × 40 cm3 filled with approximately 30 L of filtered seawater (Figure A1). An aquarium pump and air stone was added to the tanks to provide aeration and circulation of the water around the vessels, facilitating the transfer of water into and out of the vessels during tests. Only one type of sediment was contained within the tanks at a time and tanks were thoroughly washed with 10% HNO3 and deionised water prior to reuse. This method allowed for the quick and easy transfer of organisms between sediments to create the 'pulsed' exposure scenarios whilst minimising stress caused from handling M. plumulosa.
Page | 175 Mesh lid
Filter for Holding aeration vessel
Sediment
Figure A1 Experimental setup of pulsed sediment exposures showing the placement of the holding vessels on contaminated sediment and the aquaria in which the vessels were housed.
Page | 176 Appendix II
Acute lethality resulting from the intermittent exposure of M. plumulosa to contaminated sediment
100
80
60
40 Survival (% control) 20
0 Control 3x 24h 1x 48 h 2x 48h 3x 48h Continuous Exposure duration and frequency
Figure A2 The effect of duration and frequency of contaminated sediment exposure on the acute toxicity of M. plumulosa exposed to Sediment 1.
Page | 177 Appendix III
Summary of sediment pulse scenario tests
Table A1 A summary of acute and chronic toxicity (relative to controls ± standard error) observed for M. plumulosa following pulsed exposure to contaminated sediment.
Sample Endpoint Duration of Exposure Result Organism
Sediment 1 Acute Lethality 48-h 100±5% M. plumulosa Acute Lethality 72-h (3×24-h exposures) 96±5% M. plumulosa Acute Lethality 96-h (2×48-h exposure) 75±8% M. plumulosa Acute Lethality 144-h (3×48-h exposures) 14±5% M. plumulosa Acute Lethality 240-h 12±5% M. plumulosa Reproductive Toxicity 8-h 85±15% M. plumulosa Reproductive Toxicity 16-h 66±3% M. plumulosa Reproductive Toxicity 48-h 34±8% M. plumulosa Reproductive Toxicity 48-h (pre-exposure of females) 34±7% M. plumulosa Reproductive Toxicity 48-h (pre-exposure of males) 112±10% M. plumulosa Sediment 2 Acute Lethality 48-h 93±2% M. plumulosa Acute Lethality 240-h 24±6% M. plumulosa Sediment 3 Acute Lethality 48-h 96±2% M. plumulosa Acute Lethality 240-h 32±2% M. plumulosa Sediment 4 Acute Lethality 240-h 59±11% M. plumulosa Reproductive Toxicity 48-h 54±18% M. plumulosa
Table A2 A summary of acute and chronic toxicity (relative to controls ± standard error) observed for N. spinipes following pulsed exposure to contaminated sediment.
Sample Endpoint Duration of Exposure Result Organism
Sediment 1 Acute Lethality 24-h > 60% N. spinipes Acute Lethality 48-h 28±2% N. spinipes Reproductive Toxicity 24-h 16±4% N. spinipes Sediment 2 Reproductive Toxicity 48-h 63±9% N. spinipes
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