marine pollution and echinoderms: a biomarker study ...

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04 University of Plymouth Research Theses 01 Research Theses Main Collection

2009

MARINE POLLUTION AND

ECHINODERMS: A BIOMARKER

STUDY INTEGRATING DIFFERENT

LEVELS OF BIOLOGICAL

ORGANISATION

Canty, Martin Neil

http://hdl.handle.net/10026.1/1936

University of Plymouth

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MARINE POLLUTION AND ECHINODERMS: A BIOMARKER STUDY INTEGRATING DIFFERENT

L E V E L S OF BIOLOGICAL ORGANISATION

by

MARTIN NEIL CANTY

A thesis submit ted to the Universi ty of P lymouth in partial fulf i lment for the degree of

DOCTOR OF PHILOSOPHY

School of Biomedical a n d Biological Sciences University of P lymouth

December 2 0 0 9

MARINE POLLUTION AND ECHINODERMS: A BIOMARKER STUDY INTEGRATING DIFFERENT L E V E L S OF BIOLOGICAL ORGANISATION

Martin Neil Canty

Abstract

There is growing concern that the invertebrate lest organisms commonly

employed in the field of aquatic ecoloxicology may not be sufficient to

accurately screen for the possible deleterious effects of contaminants

discharged into the marine environment. The use of echinoderms has been

proposed to redress this problem, due to their ecological importance and their

evolutionary closeness to the chordates. But to date, there is a paucity of data

in the published literature which has utilised the adult stages of echinoderms in

laboratory based toxicology studies. The present studies aimed to fill this lack of

information.

A suite of biomarkers which operated at different levels of biological

organisation (sub-cellular, cellular and individual level) were identified for use

with different echinoderm species (the common sea star, Asterias rubens, the

purple sea urchin, Paracentrotus lividus and the common brittle star, Ophiothrix

fragilis). These biomarkers used were micronucleus induction, the Comet

assay, the modified Comet assay, phagocytosis, neutral red retention,

clearance rate and righting time.

Concurrent exposures showed that echinoderms were more sensitive to model

contaminants than a commonly used sentinel ecotoxicological test species,

namely the blue mussel, MytHus edulis. These contaminants included: the

reference toxicants hydrogen peroxide (Chapter 2) and methyl

methanesulphonate (Chapter 3); a pharmaceutical, cyclophosphamide (Chapter

3); a polycyclic aromatic hydrocarbon. benzo(a)pyrene (Chapter 4) and a heavy

metal, cadmium (Chapter 5).

The results for all the exposures showed that the biomarkers which operated at

the lower levels of biological organisation (i.e. at the subcellular level - namely

the micronucleus and Comet assays) were the most sensitive at detecting the

deleterious effects of the contaminants. But, interestingly, some strong

correlations were found between these sub-cellular consequences and those

that operated at higher levels of biological organisation (for example, between

righting time and both micronucleus induction and Comet assay in Asterias

rubens following cyclophosphamide exposure). Theses correlations suggest

that biomarkers which operate at the whole organism level (namely righting time

and clearance rate) may serve as rapid and accurate indicators of possible

damage induced by xenobiotics in echinoderms and bivalve molluscs.

CONTENTS

CHAPTER 1 1

Introduction 2

1.1 Pollution in aquatic ecosystems 2 1.2 Biomarkers and their use in aquatic toxicology 3 1.3 Specific types of marine pollutants 21 1.4 Reference toxicants 25 1.5 Use of echinoderms in toxicology and ecotoxicology 27 1.6 Aims and objectives 37

CHAPTER 2 41

Materials and methods 42

2.1 Test organism collection and maintenance 42 2.2 Haemolymph and coelomic fluid collection 47 2.3 Cell membrane stability (neutral red retention assay) 52 2.4 Protein determination 55 2.5 Immunocompetence (phagocytosis assay) 56 2.6 Micronuclei induction 59 2.7 Cellular viability (Eosin Y staining) 62 2.8 The Comet assay 63 2.9 Clearance rate 68 2.10 Righting behaviour 72 2.11 Determination of cadmium levels in tissues of organisms using

analytical methods 75

CHAPTER 3 79

Echinoderms {Asterias rubens) are more sensitive to know genotoxins than mahne molluscs {Mytilus edulis).

Genotoxic responses can be linked to cytotoxic and behavioural (physiological) consequences.

3.1 Introduction 80 3.2 Materials and methods 86 3.3 Results 92 3.4 Discussion 105

CHAPTER 4 117

A model PAH (benzo(a)pyrene can cause behavioural (physiological), cytotoxic and genotoxic consequences in three different echinoderm

IV

species {Asterias rubens, Paraceritrotus lividus and Ophiothrix fragilis) and a bivalve mollusc (MytHus edulis).

These consequences at different levels of biological organisation will be reversible.


CHAPTER 5 147

A heavy metal (cadmium) can cause behavioural (physiological) and cytotoxic consequences in echinoderms (Asterias rubens) and molluscs {Mytilus edulis), and reversible immunotoxic, genotoxic and oxidative stress consequences In both species.


CHAPTER 6 193

General discussion 194

LIST OF FIGURES

Fig. 1.1 At lower levels of biological organisation, biologists have an increased

mechanistic understanding the effects of pollutants, despite the low ecological

relevance at these levels. Conversely, at the more ecologically relevant levels,

there is a reduced understanding the effects of pollutants (from Hinton et al.,

2005) 4

Fig. 1.2 The progression of the health status of an organism as pollutant

exposure increases; h, the departure from normal homeostatic responses; c,

the limit of the individual's compensatory responses; r, the limit beyond which

repair mechanisms are ineffective (from Walker et al., 1996) 7

Fig. 1.3 The methodology followed for the Comet assay and the modified

Comet assay. The enzyme incubation step is only used in the modified Comet

assay to quantify oxidised DNA bases; this step is omitted from the Comet

assay 17

Fig. 1.4 Micronuclei (Mn) formation in a dividing nucleated cell (adapted from

Fenech, 1993) 20

Fig. 1.5 The structure of the genotoxin cyclophosphamide 22

Fig. 1.6 The structures of eight of the sixteen priority PAHs as termed by the

USEPA (adapted from ATSDR, 1995) 23

Fig. 1.7 The structure of the genotoxin methyl methanesulphonate 26

Fig. 1.8 The structure of hydrogen peroxide 27

Fig. 1.9 Phylogenetic tree showing the closeness of echinoderms to the

chordates (vertebrates). The organisms along the top have specific genome

projects currently initiated or completed. (From Sodergren et al., 2006) 28

Fig. 1.10 Outline of the thesis, detailing the topics covered in each chapter... .40 vi

Fig. 2.1 The test species used in these studies (a) Asterias rubens

(b) Paracentrotus lividus (c) Ophiothrix fragilis (d) Mytilus edulis. Scale bars =

10 mm 44

Fig. 2.2 An outline of the British Isles and Eire with animal collection sites

highlighted (white circles denote collection areas at each specific site), (a) The

area in and around Plymouth Sound (Devon, UK) where sea stars and brittle

stars were collected, (b) The area near Bantry, (Co. Cork, Eire) where sea

urchins were collected, (c) Port Quin (Cornwall, UK) where mussels were

collected. (Satellite images from http://maps.google.co.uk) 45

Fig. 2.3 Husbandry stock tanks used to house Asterias rubens. Air-driven

filtration units and Mytilus edulis as food source are visible 46

Fig. 2.4 Haemolymph extraction from the posterior adductor muscle of Mytilus

edulis. (a) and (b) show syringe and needle positioned in the posterior adductor

muscle, (c) A syringe containing haemolymph and (d) is a mussel with both

valves opened, showing the position of the posterior adductor muscle (red

arrows) 48

Fig. 2.5 Giemsa stained haemocytes from Mytilus edulis haemolymph. Scale

bar =15 pm 49

Fig. 2.6 Coelomocyte extraction from the 'inter-arm' region of Asterias rubens..

50

Fig. 2.7 Giemsa stained cells from Asterias rubens coelomic fluid. Phagocytic

amoebocytes (pa) are the most common cell types, followed by spherule cells

(s). Scale bar = 10 pm 51

Fig. 2.8 Flow chart illustrating the experimental design for the in vitro validation

of the NRR assay in Asterias rubens, Paracentrotus lividus and Mytilus edulis.

54

VM

Fig. 2.9 Neutral red retention as expressed by cell membrane stability (optical

density at 550 nm mg'^ protein) following in vitro exposure to H 2 O 2 for Asterias

rubens, Paracentrotus lividus and Mytilus edulis. Concentrations are expressed

in pM. The values are means (n = 5) ± SEM. * indicates a significant difference

from controls at, or above, the 95 % confidence level 55


of the phagocytosis assay in Asterias rubens and Mytilus edulis 58

Fig. 2.11 Phagocytotic activity expressed as zymosan particles phagocytosed

(x 10^ mg protein'^) following in vitro exposure to H 2 O 2 for Asterias rubens and

Mytilus edulis. Concentrations are expressed in pM. The values are means

(n = 4)± SEM. * indicates a significant difference from controls at, or above, the

95 % confidence level 59

Fig. 2.12 The appearance of some typical Mn. (a) A cell with one Mn present,

(b) Two Mn present, one close to. but not overlapping the main nuclei, (c) A cell

with two Mn which are close to one another, but not overlapping 61

Fig. 2.13 Examples of Mn present in (a) a Mytilus edulis haemocyte and in (b)

an Asterias rubens coelomocyte. The Mn are shown by the arrows. Scale

bar = 10 pm 61

Fig. 2.14 Mussel haemocytes stained with Eosin Y. The cell on the left (a) is

viable (green), the cell on the right (b) is unviable/ dead (red). Scale bar =

10 Mm 62

Fig. 2.15 Mytilus edulis haemocytes illustrating increasing levels of DNA

damage as detected by the Comet assay, with (a) representing undamaged

cells, through to (d) exhibiting the most. Scale bar = 50 pm 64


of the modified Comet assay in Asterias rubens, Paracentrotus lividus and

Mytilus edulis 66

VIII

Fig. 2.17 Percentage tail DNA following in vitro exposure to H 2 O 2 for (a)

Asterias rubens, (b) Paracentrotus lividus and (c) MytHus edulis. Concentrations

are expressed in pM. The values are means (n = 5) ± SEM. * indicates a

significant difference from controls at. or above, the 95 % confidence level 67

Fig. 2.18 Clearance rate beakers arranged across two 15-point magnetic stirring

plates (IKA ® RO 15). Mussels are visible on the base of each of the

experimental vessels and the empty 'blank' beakers are seen at the front of

each stirring plate 69

Fig. 2.19 Beckman™ Coulter Particle Size and Count Analyser (Z2) used to

count algal particles in samples collected during the clearance rate assay 70

Fig. 2.20 Experimental set-up for determining the clearance rates of brittle stars.

The algal suspension in each vessel is kept in suspension via a Pasteur pipette

driven by an air pump. The left most beaker is a system blank and brittle stars

are visible on the base of the other beakers 71

Fig. 2.21 Aquarium used to determine righting behaviour (with a partially righted

Asterias rubens) 72

Fig. 2.22 Series of images illustrating the righting response in Asterias rubens

(the sequence of images runs from (a) to (i)). Scale bar = 10 mm

74

Fig. 2.23 (a) An opened mussel showing the location of the gills (red arrows)

and the gut (black arrow), (b) Mussel with gills removed, (c) Mussel with both

the gills and gut (digestive gland) removed, (d) The body compartments on

slides prior to drying to a constant weight 75

Fig. 2.24 (a) Asterias rubens specimen prior to dissection, (b) Outer dermis

removed from central disc revealing the cardiac stomach (black arrow), (c)

Specimen with cardiac stomach removed (black arrow), (d) Tubed feet along an

ambulacral ridge, (e) Pyloric caeca visible on two dissected arms (black

arrows) 77

ix

Fig. 3.1 Flow chart illustrating the integrated experimental design adopted in the

study to evaluate the behavioural (physiological), cytotoxic and genotoxic

effects of genotoxins in (a) Asterias rubens and (b) Mytilus edulis 89

Fig. 3.2 Behavioural (physiological), cytotoxic and genotoxic effects in Asterias

rubens exposed to MMS for 3, 5 and 7 d. Concentrations are expressed as

mg L'\ The values are means ± SEM. * indicates a significant difference from

controls at, or above, the 95 % confidence level. NT, not tested; M, 100%

mortality; NR. animal failed to right following 3 x 600 s attempts 96

Fig. 3.3 Behavioural (physiological), cytotoxic and genotoxic in Mytilus edulis

exposed to MMS for 3, 5 and 7 d. Concentrations are expressed as mg L"V The

values are means +/- SEM. * indicates a significant difference from controls at,

or above, the 95 % confidence level 97

Fig. 3.4 Linear regression analysis illustrating correlations for behavioural

(physiological), cytotoxic and genotoxic effects in Mytilus edulis following 7 d

exposure to MMS. The solid line is a linear regression and the dashed lines

represent 95 % confidence limits 98


rubens exposed to CP for 3, 5 and 7 d. Concentrations are expressed as

mg L'V The values are means ± SEM. * indicates a significant difference from

controls at the 5 % level. NT, not tested 100


(physiological), cytotoxic and genotoxic effects in Asterias rubens following 7 d

exposure to CP. The solid line is a linear regression and the dashed lines


Fig. 3.7 Behavioural (physiological), cytotoxic and genotoxic effects in Mytilus

edulis exposed to CP for 3. 5 and 7 d at 15 ± 0.2 °C. Concentrations are

expressed as mg L'\ The values are means ± SEM. * indicates a significant

difference from controls at, or above, the 95 % confidence level 103



exposure to CP. The solid line is a linear regression and the dashed lines


Fig. 3.9 Metabolic pathway of cyclophosphamide in a vertebrate system

(McCarrolletal.. 2008) 108

Fig. 4.1 A benzo(a)pyrene molecule (consisting of a benzene ring attached to a

pyrene molecule) 119

Fig. 4.2 The sequence of metabolic activation of B(a)P in vertebrates. From

(Jacob. 1996) 120

Fig. 4.3 Flow chart illustrating the integrated experimental design adopted in the

studies to evaluate the behavioural (physiological), cytotoxic and genotoxic

effects of B(a)P in (a) sea stars/ urchins, (b) mussels and (c) brittle stars 126


rubens exposed to B(a)P for 3 and 7 d and following a 3 d recovery period (in

clean seawater). Concentrations are expressed as \^g L'\ The values are

means ± SEM (n = 5). * indicates a significant difference from controls at, or

above, the 95 % confidence level 132

Fig. 4.5 Behavioural (physiological), cytotoxic and genotoxic effects in

Paracentrotus lividus exposed to B(a)P for 3 and 7 d and following a 3 d

recovery period (in clean seawater). Concentrations are expressed as pg L'\

The values are means ± SEM (n = 5). * indicates a significant difference from

controls at, or above, the 95 % confidence level 133

Fig. 4.6 Behavioural (physiological), cytotoxic and genotoxic effects in Mytilus

edulis exposed to B(a)P for 3 and 7 d and following a 3 d recovery period (in

clean seawater). Concentrations are expressed as pg L'\ The values are

means ± SEM (n = 5). * indicates a significant difference from controls at. or

above, the 95 % confidence level 134 xl

Fig. 4.7 Behavioural (physiological) effects in Ophiothrix fragilis exposed to

B(a)P for 3 and 7 d and following a 3 d recovery period (in clean seawater).

Concentrations are expressed as pg L'V The values are means ± SEM (n = 5).

* indicates a significant difference from controls at. or above, the 95 %

confidence level 135


(physiological), cytotoxic and genotoxic effects in >As^er/as rubens following 7 d

exposure to B(a)P. The solid line is a linear regression and the dashed lines



(physiological), cytotoxic and genotoxic effects in Paracentrotus lividus following

7 d exposure to B(a)P. The solid line is a linear regression and the dashed lines






Fig. 4.11 Linear regression analysis illustrating correlation for behavioural

(physiological) effects in Ophiothrix fragilis following 7 d exposure to B(a)P. The

solid line is a linear regression and the dashed lines represent 95 % confidence

limits 139

Fig. 4.12 Four possible pathways for the activation of B(a)P leading to DNA

lesions in vivo In a mammalian cell system (from Canova et al., 1998) 144

Fig. 5.1 Flow chart illustrating the integrated experimental designs adopted In

the studies to evaluate the behavioural (physiological) and cytotoxic effects of

Cd (as well as body burdens) in (a) sea stars and (b) mussels 153

XII

Fig. 5.2 Flow chart illustrating the integrated experimental designs adopted in

the studies to evaluate the immunotoxic and genotoxic effects of Cd in (a) sea

stars and (b) mussels 154

Fig. 5.3 Cadmium concentrations in exposure vessels at the beginning (0 d) and

following end (5 d) of static exposure of Mytilus edulis (n = 5) and Asterias

rubens (n - 5) for 5 d to nominal concentrations of 100, 180 and 320 pg L" of

cadmium. The values are the means ± SEM 160

Fig. 5.4 Cadmium body burdens for different body components in (a) Asterias

rubens and (b) Mytilus edulis (n = 5) following 5 d exposure to nominal

concentrations of 100, 180 and 320 pg L" of cadmium. The values are the

means ±SEM 161

Fig. 5.5 Linear regression analysis illustrating correlations for amount of Cd

removed from aqueous solution in the exposure vessels over the 5 d exposure

period (pg L'^) and Cd body compartment burdens (mg kg'^) for the (a) gut, (b)

pyloric caeca and (c) tubed feet in Asterias rubens following 5 d exposure to Cd.

The solid line is a linear regression and the dashed lines represent 95 %

confidence limits 162

Fig. 5.6 Linear regression analysis illustrating correlations for amount of Cd

removed from aqueous solution in the exposure vessels over the 5 d exposure

period (pg L'^) and Cd body compartment burdens (mg kg'^) for the (a) gut and

(b) gills in Mytilus edulis following 5 d exposure to Cd. The solid line is a linear

regression and the dashed lines represent 95 % confidence limits 163

Fig. 5.7 Behavioural (physiological) effects in (a) Asterias rubens (righting time)

and (b) Mytilus edulis (clearance rate) exposed to cadmium for 5 d.

Concentrations are expressed as pg l ' \ The values are means ± SEM (n = 5).

165


(physiological) effects and cadmium body compartment burdens (mg kg"^) for

the (a) gut, (b) pyloric caeca and (c) tubed feet in Asterias rubens following 5 d xiii



Fig. 5.9 Linear regression analysis Illustrating correlations for behavioural

(physiological) effects and cadmium body compartment burdens (mg kg"^) for

the (a) gut and (b) gill in Mytilus edulis following 5 d exposure to B(a)P. The


limits 167

Fig. 5.10 Cytotoxic effects (the neutral red retention assay) in (a) Asterias

rubens and (b) Mytilus edulis exposed to cadmium for 5 d. Concentrations are

expressed as yig L"V The values are means ± SEM (n = 5) 168

Fig. 5.11 Linear regression analysis Illustrating correlations for behavioural

(physiological) and cytotoxic effects in (a) Asterias rubens and (b) Mytilus edulis

following 5 d exposure to Cd. The solid line is a linear regression and the

dashed lines represent 95 % confidence limits 169

Fig. 5.12 Linear regression analysis illustrating correlations for cytotoxic effects

and cadmium body compartment burdens (mg kg'^) for the (a) gut, (b) pyloric

caeca and (c) tubed feet in Asterias rubens following 5 d exposure to Cd. The


limits 170

Fig. 5.13 Linear regression analysis illustrating correlations for cytotoxic effects

and cadmium body compartment burdens (mg kg'^) for the (a) gut and (b) gill in

Mytilus edulis following 5 d exposure to Cd. The solid line is a linear regression

and the dashed lines represent 95 % confidence limits 171

Fig. 5.14 Immunotoxic effects in (a) Asterias rubens and (b) Mytilus edulis

exposed to cadmium for 5 d. Concentrations are expressed as pg L'V The

values are means ± SEM (n = 3). * indicates a significant difference from

controls at, or above, the 95 % confidence level 175

XIV

Fig. 5.15 Genotoxic effects in Mytilus edulis exposed to cadmium for 5 d and

following 1 d and 5 d recovery periods. Modified Comet assay conditions:

(a) enzyme buffer control; (b) Endo III enzyme and (c) PPG enzyme.

Concentrations are expressed as pg L'V The values are means ± SEM (n = 3).

* indicates a significant difference from controls at, or above, the 95 %


Fig. 5.16 Repair of single strand DNA breaks (a), oxidised purines (b) and

oxidised pyrimidines (c) by Asterias rubens coeiomocytes in vivo following 1 d

(Rdl) and 5 d (Rd5) recovery periods after a 5 d exposure to cadmium (d5).

Concentrations are expressed as pg L'\ The values are means ± SEM (n = 3). ..

177

Fig. 5.17 Linear regression analysis illustrating correlations for immunotoxic

(phagocytosis assay) and genotoxic responses (modified Comet assay, buffer

enzyme control treatment) in Mytilus edulis following (a) 5 d exposure to Cd. (b)

1 d recovery and (c) 5 d recovery. The solid line is a linear regression and the



(phagocytosis assay) and genotoxic responses (modified Comet assay, Endo III

enzyme treatment) in Mytilus edulis following (a) 5 d exposure to Cd, (b) 1 d

recovery and (c) 5 d recovery. The solid line is a linear regression and the



(phagocytosis assay) and genotoxic responses (modified Comet assay, FPG

enzyme treatment) in Mytilus edulis following (a) 5 d exposure to Cd. (b) 1 d

recovery and (c) 5 d recovery. The solid line is a linear regression and the


Fig. 5.20 Genotoxic effects in Mytilus edulis exposed to cadmium for 5 d and

following 1 d and 5 d recovery periods. Modified Comet assay conditions:

(a) enzyme buffer control; (b) Endo III enzyme and (c) FPG enzyme.

Concentrations are expressed as pg L'V The values are means ± SEM (n = 3). XV

* indicates a significant difference from controls at. or above, the 95 %


Fig. 5.21 Repair of single strand DNA breaks (a), oxidised purines (b) and

oxidised pyrimidines (c) by Mytilus edulis haemocytes in vivo following 1 d

(Rdl) and 5 d (Rd5) recovery period after a 5 d exposure to cadmium (d5).

Concentrations are expressed as M9 L 'V The values are means ± SEM (n = 3)...

182


(phagocytosis assay) and genotoxic responses (modified Comet assay buffer

enzyme control treatment) in Mytilus edulis following (a) 5 d exposure to

cadmium, (b) 1 d recovery and (c) 5 d recovery. The solid line is a linear

regression and the dashed lines represent 95 % confidence limits 183


(phagocytosis assay) and genotoxic responses (modified Comet assay Endo III

enzyme treatment) in Mytilus edulis following (a) 5 d exposure to cadmium, (b)




(phagocytosis assay) and genotoxic responses (modified Comet assay FPG

enzyme treatment) in Mytilus edulis following (a) 5 d exposure to cadmium, (b)



XVI

LIST OF T A B L E S

Table 1.1 Examples of selected biomarkers observed at different levels of

biological organisation in aquatic organisms 8

Table 1.2 Examples of biomarkers used with the adult stages of echinoderm

species belonging to the three echinoderm classes (Asleroidea. Echlnoidea and

Ophiuroidea) utilised in the present studies 30

Table 3.1 Percentage survival for Asterias rubens and Mytilus edulis following

7 d exposure to cyclophosphamide (Experiment 1) and methyl

methanesulphonate (Experiment 2) 93

Table 4.1 Estimated global input of PAHs into the marine environment (from

Clarke, 2001) 119

Table 4.2 Examples of selected in vivo ben2o(a)pyrene biomarker studies with

the adult stages of different aquatic test species 122

Table 6.1 Human disease genes and their sea urchin ortholog genes (adapted

from Sodergren et al, 2006) 198

XVII

Acknowledgements

There are countless individuals to whom I am deeply indebted to for their help

over the course of these studies. Firstly, and most importantly, I would like to

thank my director of studies, Dr Awadhesh Jha for his guidance and support

throughout this PhD. I am profoundly grateful for his words of wisdom and

encouragement from day one of these studies right through to their culmination.

I would also like to thank my other PhD supervisors. Prof Mai Jones, Prof Tom

Hutchinson and Dr Becky Brown for their support. I am also very grateful to

Debbie Petherick for her understanding and very kind words when I needed

them the most.

I am grateful to NERC and my former colleagues at Brixham Environmental

Laboratory (AstraZeneca) for their financial support which gave me the

opportunity to carry out this PhD.

Many other people have offered me support; these include members of T e a m

Ecotox' past and present - especially Dr Awantha Dissanayake, Dr Mark

Browne, Dr Marie Hannam, Dr Alan Scarlett. Dr Chris Pook, Dr Vikki Cheung

and Dr Jo Hagger. Thanks must also go to Dr Rob Clough and Dr Andy Fisher

for their tireless help with my metal analysis. Particular thanks also go to Stan

McMahon and Ben Eynon for assistance with livestock husbandry and to Chris

Booker (and other members of the University of Plymouth Diving School) for

carrying out animal collections on my behalf at very short notice.

I need to thank Lou Ramsay for going above and beyond the call of duty to help

out a friend in need. Finally, I would not have been in a position to commence

(let alone complete) this work if it had not been for the tireless love and support

offered to me by my family. Thank you so much Mum, Dad, Sis and Nige - I

hope one day to be able to repay you all for everything.

In loving memory of two people who inspired me to study all things aquatic,

Marcia Flint and Bob Hunter. In the words of two great men, Messrs. Lawrence

and Adams: "Stand still and eventually you go backwards" and "So long and

thanks for all the (star) fish". xviii

AUTHOR'S D E C L A R A T I O N

At no time during the registration for the degree of Doctor of Philosophy has the

author been registered for any other University award without prior agreement

of the Graduate Committee.

This study was financed with the aid of a CASE award studentship from the

Natural Environment Research Council in collaboration with AstraZeneca

(Brixham Environmental Laboratory, Brixham, U.K.).

Relevant scientific seminars and conferences were attended at which work was

presented and manuscripts were prepared for publication.

Publications (please refer to Appendix II for peer-reviewed manuscripts):

Martin N. Canty. Thomas H. Hutchinson, Rebecca J. Brown. Malcolm B. Jones

and Awadhesh N. Jha (2009). Linking genotoxic responses with cytotoxic and

behavioural or physiological consequences: differential sensitivity of

echlnoderms (Asterias rubens) and marine molluscs (Mytilus edulis). Aquatic

Toxicology. 94 (1): 68-76.

Martin N. Canty. Rebecca J. Brown, Thomas H. Hutchinson, Malcolm B. Jones

and Awadhesh N. Jha (2007). Determination of oxidative DNA damage and

cellular viability in marine invertebrates. Mutagenesis, 22 (6): e18.

Platform Presentations:

The differential sensitivity of echinoderms and molluscs to genotoxins using

multiple biomarkers. SETAC Europe 16th Annual Meeting, Hague, 8'^ May

2005.

xix

Poster Presentations:

Martin N. Canty. Thomas H. Hutchinson, Rebecca J. Brown, Malcolm B. Jones

and Awadhesh N. Jha (2007). Determination of oxidative DNA damage and

cellular viability in marine invertebrates. VII International Comet Workshop,

University of Ulster. Coleraine, 24'^ - 27'^ June 2007.

Word count of main body of thesis; 33,033

Signed nO^t^ Mor^A Ce^:

Date. !.t/p7/ i O .

XX

ABBREVIAT IONS AND A C R O N Y M S

ANOVA ASTM B(a)P BOA BER c. Cabs CAS Cd CHO CP CR d DMSO DNA EA ECETOC

EMS Endo III ERAS EU Fig. FPG 9 h H2O2 HEPES HNO3 ICP-MS ITS LC50 KCI LMP M MFO Min MMS Mn M MTC MTD MTs n Na2EDTA NADPH NER NMP NR

Analysis of variance American Society for Testing Materials Ben2o(a)pyrene Bicinchoninic acid Base excision repair Circa Chromosomal aberrations Chemical abstracts service Cadmium Chinese Hamster ovary cells Cyclophosphamide Clearance rate Day Dimethyl sulphoxide Deoxyribonucleic acid Environment Agency European Centre for Ecotoxicology and Toxicology of Chemicals Ethyl methanesulphonate Endonuclease III Environmental Risk Assessments European Union Figure Formamidopyrimidine DNA glycosylase Standard gravity Hour Hydrogen peroxide 4-(2-hydroxyethyl)-1 -piperazineelhanesuffonlc acid Nitric acid Inductively Coupled Plasma Mass Spectrometry intelligent testing strategies Lethal concentration (of 50% of test organisms) Potassium chloride Low melting point Mortality Mixed-function oxygenase Minute Methyl methanesulphonate Micronucleus/ micronuclei Mortality (100%) Maximum tolerated concentration Maximum tolerated dose Metallothioneins Number of test organisms per sample (concentration) Disodium ethylenediaminetetraacetic acid Nicotinamide adenine dinucleotide phosphate Nucleotide excision repair Normal melting point Neutral red

XXI

NRR Neutral red retention NT Not tested P Probability P450 Cytochrome P450 PAH Polycyclic aromatic hydrocarbon PAM Phosphoramide mustard Pb Lead PBS Phosphate buffered saline PCBs Polychiorinated biphenyls PCPs Personal care products ppb Parts per billion ppm Parts per million PRI Profileralive rate index PSU Practical salinity units R Pearson correlation coefficient rd Recovery day REACH Registration, evaluation, authorisation and restriction of

chemicals ROS Reactive oxygen species RT Righting time s Seconds SAC Special Areas of Conservation SCEs Sister chromatid exchanges SD Standard deviation SEM Standard error margin SCGE Single cell gel electrophoresis SSBs Single strand breaks SW Seawater TBT Tributyltin TPT Triphenyltin TIES Toxicity identification evaluations USEPA United States Environmental Protection Agency UV Ultraviolet V Volts v/v Volume/ volume w/v Weight/ volume WFD Water Framework Directive WSF Water-soluble fraction Zn Zinc

XXII

Chapter 1

Introduction

1. Introduction

1.1 Pollution in aquatic ecosystems

Coastal systems are experiencing ever increasing population and exploitation

pressures; nearly 4 0 % of the people in the world live within 100 km of the

coast. Within these coastal populations, 71 % live within 50 km of estuaries

(Agardy and Alder, 2005). These growing coastal populations inevitably lead to

anthropogenic inputs into the aquatic environment. Anthropogenic pollutants

can include (but. are not restricted to): Industrial chemicals (e.g. petrochemicals

and solvents); industrial products (e.g. flame retardants and lubricants);

consumer products (pharmaceuticals and personal care products); biocides

(e.g. insecticides and pesticides) and 'natural' chemicals and compounds (e.g.

heavy metals and human hormones) (Schwarzenbach et al., 2006). These

anthropogenic pollutants not only adversely affect the flora and fauna of aquatic

ecosystems, but can also ultimately impact on human health via the food chain

(Jha et al.. 2000a).

Traditionally chemical monitoring has been used to detect and monitor

pollutants in the aquatic environment. This approach however has limitations; it

is possible that certain pollutants may occur at levels below the detection limits

for chemical analysis, yet they may be present at sufficient concentrations to

deleteriously affect the health of marine organisms. As a result, biomonitoring is

carried out alongside chemical monitoring. Biomonitoring programmes utilise

'sentinel' organisms that are omnipresent in aquatic systems and are

representative of the communities they inhabit, the latter point being important

as it is not feasible to assess the health of all the different species present in a

given community (Depledge and Fossi, 1994). The health of these target

species are monitored on a frequent basis as a surrogate for the overall 'health'

of the aquatic environment. To this end, biomarkers have been devised and

utilised to determine the 'well being' of individual vertebrate and invertebrate

aquatic species at different levels of biological organisation.

1.2 Biomarkers and their use in aquatic toxicology

Biomarkers are indicators of adverse biological effects evident at the

biochemical, cellular, histological, physiological or behavioural levels of

biological organisation. Biomarkers may indicate exposure to, and the toxic

effect of a physical pollutant(s) or chemical(s) (Depledge and Fossi, 1994).

Biomarkers were originally used in the human health arena to specifically

determine, and evaluate, the genetic and cancer risk presented by a range of

environmental toxins (AlberlinI et al., 1996). Following this early work, the

'biomarker approach' has been adopted for use in the fields of biomonitoring

and toxicology.

Biomarkers can detect adverse effects of environmental contaminants at

concentrations below chemical detection limits (Rand, 1995). The majority of

toxins impact initially at lower levels of biological organisation (molecular, or

cellular levels) before they are apparent at higher levels (such as at the

physiological or behavioural levels). By implementing biomarkers in

environmental monitoring programmes, the potential impact of pollutants may

be detected in individual sentinel organisms before any deleterious

consequences are expressed at population levels (Hinton et al., 2005). These

lower levels of biological organisation are thought to have low ecological

relevance, but do provide an understanding of the mode of action of toxins at

these levels (Fig. 1.1). Traditionally, one of the problems with utilising these

lowers levels of biological organisation to detect environmental contamination is

that the ecological relevance of any individual-based responses are difficult to

extrapolate to higher levels of organisation. However, Depledge and Galloway

have made a clear case for the value of individuals in environmental monitoring

(Depledge and Galloway, 2005).

Increasing Ecological Relevance

C o m m u n i t y

P o p u l a t i o n

Ind iv idua l

Ce l lu la r

M o l e c u l a r

Increasing Mechanistic Understanding

Fig. 1.1 At lower levels of biological organisation, biologists have an increased

mechanistic understanding the effects of pollutants, despite the low ecological

relevance at these levels. Conversely, at the more ecologically relevant levels,

there is a reduced understanding the effects of pollutants (from Hinton et al.,

2005).

The biomarker concept is illustrated in Figure 1.2. Physiological state is show/n

on the horizontal axis and health status on the vertical axis. The figure shows

that at low levels of pollutant concentration the individual is healthy, as

exposure concentrations increase the organism becomes stressed. This is due

to energy expenditure on cellular defence mechanisms which are responsible

for reducing cellular levels of the xenobiotic. As pollutant concentration

continues to increase the individual becomes further stressed, but this damage

is reversible, the individual enters an irreversible stressed slate upon a further

increase concentration. Organism death will occur a short time later (Walker,

1998).

There is increasing pressure to implement the 'biomarker approach" in the

biomonitoring and protection of aquatic systems. One such example is the

application of biomarkers to improve environmental risk assessments (ERAs) of

chemicals in requirement of the European Commission's Water Framework

Directive (WFD) (2000/60/EC) (Hagger et al., 2008). The WFD is a piece of

legalisation from the European Union which aims to ensure that all EU member

countries have 'good quality' (in terms of good ecological and chemical status

(Hagger et al., 2006)) water bodies by 2015. Member countries are required to

periodically assess the 'quality' of all water bodies (including rivers, lakes,

estuaries, coastal waters and groundwater) and then classify them (from high

quality through to bad) dependent upon their ecological quality (EA, 2002).

Hagger et al (2008) showed that biomarkers could play an important role in

measuring organism health and therefore improve the quality and accuracy of

ERAs prepared in accordance with the WFD legalisation.

A second example of blomarker use in biomonitoring Is the application of

biomarkers to determine the ecological condition of Special Areas of

Conservation (SAC) at marine sites around England (Hagger et al., 2009). The

European Habitats Directive (Council Directive 92/43/EEC) designated areas

that contain rare, threatened, or endangered flora and or fauna as SACs.

Hagger et al (2009) illustrated the usefulness of the biomarker approach to

monitor the health of invertebrates present in a SAC and therefore help in

protecting these areas further.

Many different biomarkers have been devised to assess the deleterious effects

of a range of pollutants on aquatic organisms (examples of which are given in

Table 1.1). A number of the biomarkers listed in Table 1.1 are to be utilised in

these studies. The biomarkers selected operate at a wide-range of levels of

biological organisation, from the whole organism (righting time and clearance

rate), to cellular level (the neutral red retention and phagocytosis assays) and

genetic level (DNA strand breaks and micronucleus Induction).

n

X

Suessed

Reversible Irreversible

Compensation Noncompensation Homeostasis

Physiologic condition

Pollutant concentration

Fig. 1.2 The progression of the health status of an organism as pollutant exposure increases; h, the departure from normal homeostatic

responses; c, the limit of the individual's compensatory responses; r, the limit beyond which repair mechanisms are Ineffective (from

Walker eta l . , 1996).

Table. 1.1 Examples of selected biomarkers observed at different levels of biological organisation in aquatic organisms.

Level of biological organisation

Biomarkers References

Biochemical Induction of Cytochrome P450 Induction of metallothioneins Heat shock protein induction

(Danis et al., 2004a; Sole and Livingstone, 2005; Rewitz et al., 2006) (Bebiannoand Langston, 1991; Gerete ta l . , 2002; Brown eta l . , 2004) (Oweson et al., 2008)

Genetic Micronuclei induction DMA strand breaks

(Sanchez-Galan et al., 1999; Mouchet et al.. 2005; Binelli et al., 2009b) (Collins, 2004; Jha et al., 2005; Jha, 2008; Reeves et al., 2008; Frenzilli et al., 2009)

Cellular / Immunological Cellular membrane stability Phagocytosis

(Hannam eta l . , 2009a) (Fournier et al., 2000; Hannam et al., 2009b; Stabili and Pagliara, 2009)

Physiological Heart rate Tissue regeneration Osmoregulation

(Dissanayake eta l . , 2008) (Candia Carnevali et al., 2001; Sugni et al., 2008) (Felten et al., 2008)

Behavioural

Feeding rate Valve gape Burrowing behaviour Prey localisation Righting behaviour

(McWilliam and Baird, 2002b; Canty et al., 2007) (Tranet al., 2004) (Scarlett et al.. 2007) (Georgiades eta l . , 2003) (Kleitman, 1941;Speare etal . , 1996)

1.2.1 Righting time

Righting time (RT) is a behavioural (physiological) assay which has been

implemented with some echinoderm species. Righting time Is defined as the

time taken for an animal placed upside down to return to its normal, upright,

position, and it can be used as an indicator of the general stress of an

organism. An inherent righting response is an obvious survival advantage in

echinoderms, for instance asteroids, which inhabit rocky intertldal areas which

are exposed to heavy wave action would need to right themselves rapidly (Polls

and Conor, 1975). The righting behaviour of echinoderms was first studied

experimentally by Vulplan (1862) and by Romanes and Ewart (1881) (in Polls &

Conor, 1975). These initial studies were mainly interested in the stimulus that

caused the righting response In asteroids. The effects of ambient water

temperature on the righting response In asteroids, urchins and brittlestars were

studied by Kleitman (1941). It was found that the speed of the righting response

increased as the temperature rose, but the temperature was not properly

controlled during these trials. The data showed a strong relationship between

size and time, so it is probable that any effects of temperature were probably

less than the effects of size (Kleitman, 1941).

This assay was used in 1982 (Shirley and Stickle, 1982) to show the detrimental

effect of low salinity on the asteroid Leptasterias hexactis. The righting

behaviour of the brittle star Ophiophragmus filograneus was found to be

impaired following exposure to zinc (Clements et al., 1998). The righting

behaviour of Asterias rubens has been found to be suppressed following

exposure to the insecticide Ivermectin (Davies et al.. 1998). The righting

behaviour of Asterina gibbosa has been found to be impaired following 48 h

exposure to cadmium (Bowett, 2002) and exposure to ethyl methanesulphonate

(EMS) (Leaney. 2003). Australian sea stars Patiriella exigua were unable to

self-right after exposure to polycyclic aromatic hydrocarbon (PAH) contaminated

sediments for up to 32 d (Ryder et al., 2004). The righting behaviour of

echinoids has also been examined. Axiak and Saliba (1981) found that the

righting response of the urchin Paracentrotus lividus was prolonged following

topical exposure to crude oil. In cases of very slow righting times, a general low

level of tube feet activity was observed, suggesting a narcotising

(stupor-inducing) effect of the hydrocarbons (Axiak and Saliba, 1981). Despite

being a valuable indicator of stress at the whole organism level the intrinsic self-

righting behaviour of echinoderms has not been fully exploited to assess the

possible detrimental effects of environmental contaminants. This behavioural

trait could be a particularly valuable end-point to investigate following

contaminant exposure as it could be rapidly assessed and it is non-destructive

in nature.

1.2.2 Clearance rate

1.2.2.1 Clearance rate in bivalves

The clearance rates of bivalves, especially the blue mussel. MytHus edulis, have

been measured by researchers as a simple yet robust indication of the overall

health (or physiological condition) of an animal (Scarlett et al., 2005; Canty et

al., 2007; Scarlett et al., 2008). Clearance rate (CR) is sometimes used as a

component of the 'scope for growth' assessment (Widdows et al.. 1987;

Widdows et al., 1995). The CR (also termed feeding rate) is calculated by

determining the rate of loss of a known amount of algal particles from a given

volume of water over a given period of time. Polycyclic aromatic hydrocarbons

10

(PAHs) have been found to impair the CR of M. edulis in the form of

fluoranthene and benzo(a)pyrene (Eertman et al., 1995) and branched

alkylbenzenes (Scarlett et al., 2008). Chemicals used to disperse crude oil

spills, namely Corexit 9527 and Superdispersant-25, also significantly reduced

the CR of M. edulis (Scarlett et al., 2005), whereas the organophosphorus

pesticide azamethiphos had no effect (Canty et al., 2007).

1.2.2.2 Clearance rate in brittle stars

To date, no researchers have devised a means of quantifying the CR of

suspension-feeding echlnoderms, such as bhttlestars. In 1960, Roushdy and

Hansen observed the suspension-feeding behaviour of ophiuroids whilst

removing phytoplankton {Sceletonema costatum) particles from suspension in

static vessels. Ophiothrix fragilis removed c. 70 % of initial algal particles after

nine hours and Ophiopholis aculeate removed c. 60 %. The CR was the

greatest in the first hour (c. 35 %) for both species (Roushdy and Hansen,

1960). Field observations of O. fragilis have shown that this species can occur

in very dense aggregations (up to 2,000 individuals m^) and confirmed that it

does feed by using its arms to suspension-feed (Warner and Woodley, 1975;

Davoult and Gounin, 1995). Warner and Woodley described the particle capture

and handling technique utilised by O. fragilis. When suspension feeding,

particles were captured by the spines (located on the arms) and by the

mucus-coated tube feet; the feet also occasionally removed particles from the

spines. The captured 'food' particles then travelled down the arms towards the

mouth (Warner and Woodley. 1975). The suspension feeding behaviour of the

brittle star Amphiura filiformis has been examined in a laboratory flume and

found to be a function of flow velocity, with few animals extending their arms to

11

feed when the water was still (Loo et al., 1996). The orientation of the feeding

arms was also affected by the water flow rate. The suspension feeding

behaviour in the brittle star O. fragilis has also been investigated using video

recording equipment and polystyrene spheres, with which the particle-retention

efficiency of the animals was determined (Allen, 1998). Apart from these limited

studies, the potential use of measuring CR in suspension-feeding brittle stars,

as an indictor of pollutant exposure, has not been explored.

12.3 Neutral red retention

Neutral red (2-methyl-3-amino-7-dimethylamino-phenazine) is a weakly

cationic, supravital dye which readily penetrates cellular membranes by the

process of non-ionic diffusion (Babich and Borenfreund, 1993). Neutral red

(NR) has a history of use in the selective staining of low pH cellular

compartments, lysosomes and phagosomes (Rashid et a!., 1991). Neutral red

was first used to determine the cytotoxicity of six different toxicants to mouse

fibroblast cells using spectrophotometric analysis (Borenfreund and Puerner,

1985), the rationale being that healthy cells would take-up more NR dye than

stressed cells, therefore the higher the optical density measured

spectrophotometrically the 'healthier' the cells. This technique was termed the

neutral red retention (NNR) assay. The NRR assay gave an indication of overall

cellular membrane stability as the technique did not differentiate between NR

taken-up through the cell membrane and that which had entered the

lysosomes. Following this early work with NR, the technique was adapted for

use to specifically determine lysosomal membrane stability. In these cases a

light microscope was employed in the place of a spectrophotometer. The NR

dye was incubated on microscope slides along with the cells under investigation

12

and the time taken for the dye to enter the lysosomes of the cells was

determined. With this technique it was possible to quantify the amount of the

dye which entered the lysosomal compartment of the cells, thereby determining

the integrity (stability) of the lysosomal membranes (Babich and Borenfreund.

1991; Lowe and Pipe, 1994; Lowe et al., 1995; Cheung et a!., 1998; Binelli et

al.. 2009a). Damage to these membranes can have a detrimental effect on

maintenance of homeostasis (Lowe and Pipe. 1994), immunocompetence of an

organism, and may even lead to the autophagic loss of body tissue (Bayne and

Moore, 1998). The assay has since been modified to provide an indication of

overall cellular membrane stability through the use of a spectrophotometer and

a microtitre plate (Olabarrieta et al., 2001; Gomez-Mendikute et al., 2002;

Galloway et al.. 2004; Asensio et al., 2007; Dissanayake et al., 2008; Dailianis,

2009; Hannam et al., 2009b). This technique was much more efficient than the

NR 'slide technique' as multiple samples could be easily and rapidly assessed

at the same time.

The NRR assay has been employed with different aquatic species, including the

Arctic scallop. Chlamys islandica (Hannam et al., 2009a), the shore crab,

Carcinus maenas (Dissanayake et al., 2008), the blue mussel, M. edulis (Canty

et al., 2007; Scarlett et al., 2008) and the common limpet Patella vulgata

(Brown et al., 2004), but relatively few studies have employed this technique

with echinoderms. The lysosomal stability of Asterina gibbosa was reduced

following in vivo exposure to cadmium (Cd) (Bowett, 2002). Preliminary

investigations involving in vitro hydrogen peroxide (H2O2) exposures with

A. gibbosa and two other asteroid species (Asterias rubens and Marthasterias

glacialis) yielded mixed results. Of these test organisms only A. gibbosa

13

appeared to have reduced lysosomal stability at the very highest concentration

of H2O2 (100 |JM) (Leaney. 2003).

1.2.4 Phagocytosis

Many marine invertebrates (e.g. molluscs and echinoderms), in common with

vertebrates, possess phagocytic cells in their vascular systems (Andrew, 1962;

Warren, 1965). These cells are an important component of the immune system

as they are specialised in the recognition, ingestion and digestion of xenobiotics

(Sminia. 1980). The process of phagocytosis involves a number of stages

including chemotaxis, attachment, ingestion and killing of pathogens (Pipe et

al.. 1995). Measurement of phagocytic activity has been suggested as being a

generic and useful biomarker of immune function (Fournier et al.. 2000). One of

the simplest and most robust methods to determine the phagocytotic activity of

Invertebrates was devised by Pipe et al in 1995. This method determined

phagocytotic activity by measuring the uptake of zymosan particles (from yeast.

Saccharomyces cerevisiae) which had been dyed with NR dye. This assay has

been used to determine the immunotoxicity of a range of xenobiotics to

M. edulis haemocytes. The phagocytotic capability of M. edulis haemocytes

was impaired following exposure to PAHs (Grundy et a!., 1996; Hannam et al.,

2009b), whereas increased an increased Immune response was detected

following exposures to copper (0.2 ppm for 7 d) (Pipe et al., 1999) and the

organophosphorus compound azamethiphos (0.1 mg L" for 24 h) (Canty et al.,

2007).

There is only a very small body of work which has determined the phagocytotic

activity of echinoderms in response to xenobiotics. Exposure of the asteroid

14

Leptasterias polaris to tributyltin (TBT) led to a decrease in phagocytotic ability

(Bekri and Pelletier, 2004) and a similar trend was observed in A. rubens

following exposure to Cd (Coteur et al., 2005b). A five day manganese

exposure (at 15 mg L' ) caused a reduced phagocytotic response in A. rubens

(Oweson et al., 2008). The activity of lysozyme (an enzyme involved in immune

response to a bacterial threat) has been found to be reduced in the epidermal

mucus of the spiny sea star M. glacialis following an exposure to zinc (Stabili

and Pagliara, 2009).

1.2.5 The Comet assay

The Comet assay (sometimes termed single cell gel electrophoresis (SCGE)) is

the standard method for assessing DNA damage, with applications in

genotoxicity testing, human blomonitoring and molecular epidemiology,

ecogenotoxicology, as well as fundamental research In DNA damage and repair

(Collins, 2004). It is a rapid, sensitive and inexpensive method for measuring

DNA strand breaks (Lee and Steinert, 2003). The method was first developed in

1984 to detect double strand DNA breaks In cells that were induced following

exposure to X-rays (Ostling and Johanson, 1984), and was later adapted to

determine single strand breaks (Singh et al., 1988; Olive et al., 1990).

Subsequently, the assay was adapted and widely applied for use with aquatic

vertebrates and invertebrates as some of the other assays used to determine

DNA damage are unsuitable for aquatic organisms.

The assay detects DNA strand breaks by measuring the migration of DNA from

immobilized nuclear DNA (Fig. 1.3 outlines the procedure for preparing and

analysing samples in the Comet assay). Initially, the isolated cells are fixed in

15

an agarose gel on slides and then placed in a lysis solution. The DNA is then

denatured in an alkaline solution and the samples are subjected to

electrophoresis - when an electrical charge is applied across the slides. Once

the slides have been stained, the degree of migration away from the nucleus,

the so-called 'comet tail length", can be quantified using image analysis

software (the greater the migrafion, the greater the amount of DNA strands

breaks) (Yendle et al., 1997). The advantages of using this technique for

assessing DNA damage in aquafic organisms over other techniques are: DNA

damage in single cells is measured; only a small number of cells are required

(< 10,000 cells); the technique can be carried-out on virtually any eukaryotic

cell type and the assay is very sensifive (Lee and Sleinert, 2003). The slides

produced are scored under an epifluoresence microscope (e.g. Leica. DMR)

either by visually scoring, or using commercially available software, such as the

Komet 5.0 image-analysis software (Kinetic Imaging. Liverpool. UK). Although

the software provides a range of parameters (for example, Olive Tail Moment

and Tail Extent Moment), % tail DNA is considered to be the most reliable

parameter which allows inter-laboratory comparison (Kumaravel and Jha,

2006).

16

Cells embedded on agarose-coated slide

Cells lysed

Incubation with enzymes

Unwinding of DNA and electrophoresis

Scoring of slides

Microscope slide

Nucleus

Nucleoid

• EndoIII-Convertsoxidised pyrimidine bases to DNA strand breaks

• FPG-Converts oxidised purine bases to DNA strand breaks

Comet' tail

Cellular membrane

• •

• •

•

• •

• •

•

• t

f I

• •

• •

• • r 4 \

Comet head

Control cell Damaged ce l l -w i t h visible comet

Fig. 1.3 The methodology followed for the Comet assay and the modified Comet assay. The enzyme incubation step is only used in the

modified Comet assay to quantify oxidised DNA bases; this step is omitted from the Comet assay.

The Comet assay has been implemented previously with a number of different

aquatic invertebrate and vertebrate species (Jha, 2008; Frenzilli et al., 2009).

but its use with echinoderm species has been more limited. DNA strands breaks

have been quantified in a small number of echinoderm species. Studies carried

out on A. rubens in the North Sea found that animals collected at near shore

sites had lower DNA integrity than those from offshore reference sites

(Everaarts. 1995; Everaarts. 1997). The technique used was similar to the

Comet assay; it was the alkaline DNA unwinding assay. In vitro exposure of

coelomocytes to H2O2 has shown a dose-dependent effect on percentage tail

DNA in three different asteroid species: A. rubens, A. gibbosa and M. glacialis

(Leaney, 2003). Exposure to crude oil resulted in a significant

concentration-related increase of the percentage DNA in the comet tail of the

urchin species Strongylocentrotus droebachiensis (Taban et al., 2004). The

Comet assay has been successfully adapted to assess DNA strand breakage in

sea urchin eggs following exposure to UV radiation (Nahon et al., 2008). An

environmentally relevant dose of UV radiation caused significance increase in

percentage tail DNA in Paracentrotus lividus and Sphaerechinus granulans.

Apart from these limited studies; there have not been sufficient investigations to

determine the induction of DNA damage in different echinoderm species.

1.2.5.1 Modified Comet assay

The Comet assay can be modified with the addition of an enzyme digestion

stage to determine the degree of oxidative damage in the target DNA (Collins et

al., 1993; Dusinska and Collins, 1996). The enzymes used are

formamidopyrimidine DNA glycosylase (FPG) and endonuclease III (Endo III),

they convert oxidised purines and pyrimidines, respectively, into DNA single

18

strand breaks which can then be quantified to give an approximafion of

oxidative stress. Specifically, FPG recognises 8-hydroxydeoxyguanine and

open ringed pyrimidines, and it removes them to generate DNA strand breaks

(Kruszewski et al., 1998), whereas Endo III nicks DNA strand breaks at the sites

of oxidised pyrimidines (Collins et al.. 1993) (Fig. 1.3 outlines the procedure for

preparing modified Comet assay samples). This technique was first employed in

mammalian studies (Azqueta et al., 2009). but has also been used in vitro with

fish cells (Reeves et al., 2008) and in vivo with Mytilus edulis (Emmanouil et al.,

2007) . To date, there have been no studies which have investigated oxidafive

DNA damage in echinoderm species.

1.2.6 Micronuclei

Micronuclei (Mn) are caused by both chromosome breakages and spindle

apparatus dysfuncfion (chromosomes failing to correctly segregate at mitosis)

in mitotically dividing cells. This process of Mn formafion is illustrated in Figure

1.4. When viewed under a light microscope, Mn resemble smaller nuclei within

the cytoplasm of the cell, but, they appear completely separate to the main

nucleus.

Measuring Mn as an assessment of chromosome damage in mammalian

systems was first proposed in 1975 (Schmid, 1975). The assay has since been

employed with range of aquafic vertebrates and invertebrates, including fish

(Nepomuceno et al., 1997; Bombail et al., 2001; Grisolia, 2002; Guha et al..

2007; Winter et al.. 2007) and mussels (Dailianisa et al.. 2003; Kalpaxis et al.,

2004; Hagger et al.. 2005b; Jha et al., 2005; Villela et al.. 2007; Siu et al.,

2008) .

19

Cell

Nucleus

Chromosome Micronucleus (containing chromosome fragment)

Fragmentof chromosome Lagging

chromosome

Main nucleus

Micronucleus (containing chromosome)

Daughter cells with micron uclei

Fig. 1.4 Micronuclei (Mn) formation in a dividing nucleated cell (adapted from

Fenech, 1993).

The Mn assay has been applied to echinoderms, such as adults of the asteroid

A. gibbosa (Leaney, 2003) and larvae of the echinoids Strongylocentrotus

purpuratus (Hose and Puffer, 1983), Hemicentrotus pulcherrimus and

Clypeaster japonicus (Saotome, 1999; Saotome and Hayashi, 2003), but has

yet to be implemented successfully in other echinoderm species. The work of

Saotome initially lead to promising results following embryo larval exposure to

genotoxins (Saotome, 1999; Saotome and Hayashi, 2003), but the high density

of the embryos in the exposures may have impaired their development and

thereby, compromised the validity of these findings. Leaney (2003) found that a

3 d in vivo exposure of A. gibbosa to ethyl methanesulphonate (EMS) produced

a significant increase in Mn frequency. A parental exposure of 100 mg kg'^ of

benzo(a)pyrene was found to result in a significant increase in Mn abundance

in sea urchin gastrulae (Hose and Puffer, 1983).

20

1.3 Specific types of marine pollutants

Marine pollution is defined as "the introduction by man. directly or indirectly, of

substances or energy to the marine environment resulting in deleterious effects

such as: hazards to human health; hindrance of marine activities, including

fishing; impairment of the quality for the use of seawater, and reduction of

amenities" (Clark, 2001). Marine pollutants can take many forms, including

pharmaceuticals, polycyclic aromatic hydrocarbons (PAHs) and heavy metals.

1.3.1 Pharmaceuticals (i.e. cyclophosphamide)

Pharmaceutically active compounds are produced and administered in very

large volumes globally and their use, and diversity, increases annually (Bound

and Voulvoulis, 2004). As these compounds are not completely removed from

wastewaters by sewage treatment works they are found in surface waters and

even in potable (drinking) water (Jones et al., 2005).

One such pharmaceutical compound is cyclophosphamide (CP) {N.N-b'\s{2-

chloroethyl) tetra-hydro-2H-1,3,2-oxphosphorin-2-amine, 2-oxide monohydrate)

(Fig. 1.5). Cyclophosphamide is an indirect acting genotoxin, meaning that it

requires metabolic activation within the body of the test organism, which acts as

an alkylating agent. Cyclophosphamide is used frequently in cancer therapy to

inhibit the proliferation of tumour cells, although it is sometimes used in the

treatment of rheumatoid arthritis (Anderson et al., 1995).

21

CI

Fig. 1.5 The structure of the genotoxin cyclophosphamide.

Cyclophosphamide is normally administered orally (100 - 200 mg kg'^ daily) or

intravenously (600 - 1000 mg m^ every 3 - 4 weeks) for the treatment of cancer

(McCarroll et al., 2008). It has been detected in the sewage water from

hospitals (Steger-Hartmann et al.. 1996), as following its use in cancer

chemotherapy up to 20 % of the dose may leave the body in an unmetabolized

form (Steger-Hartmann et al., 1997). Little data exists for the impact of this

pharmaceufical by-product on the marine environment. The larvae of a marine

annelid Platynereis dumerilii were found to be sensitive to low levels of CP (Jha

et al., 1996). Chronic exposure to adult rats has been found to adversely affect

their spermatozoa (Qlu et al.. 1995), hence CP could have implications for

population structure and integrity.

1.3.2 Polycyclic aromatic hydrocarbons (i.e. benzo(a)pyrene)

Polycyclic aromafic hydrocarbons (PAHs) are a large group of persistent

organic pollutants of great environmental concern that occur ubiquitously in the

marine environment. They can originate from natural biological sources such

as plants, animals, bacteria and algae (Readman et al.. 2002). but It is the ever

increasing amount of anthropogenic PAH inputs that have led to elevated

22

concentrations in estuarine and marine systems. The sources of these inputs

include: vehicle exhaust fumes, run-off from roads, agricultural burning,

shipping traffic, discharges from industrial and/ or wastewater plants (ATSDR.

1995; Readman et al., 2002). Discharges from wastewater plants in North

America and Europe have been measured as being over 625 [jg L' and up to

4.4 mg L' from industrial effluent discharges (Latimer and Zheng, 2003).

Sixteen PAHs have termed 'priority PAH pollutants' by the United States

Environmental Protection Agency (USEPA) as they have been Identified as

being of particular importance due to their potential toxicity to mammals and

aquatic organisms. These PAHs were: acenapthylene, anthracene,

ben2o(a)athracene, benzo(a)pyrene, benzo(a)fluoranthene,

benzo(k)fluoarnthrene, benzo(ghi)perylene, chrysene, dibezo(ah)anthracene,

fluoranthene, indeno(1,2,3-cd)pyrene, naphthalene, phenanthrene and pyrene

(USEPA. 1987). The structures of eight of these sixteen priority PAHs are

shown in Figure 1.6.

Acenaphthylene

Fluoranthene

Acenaphthene Benzo(a)pyrene

Fluorene Pyrene

Chrysene

Phenanthrene

Fig. 1.6 The structures of eight of the sixteen priority PAHs as termed by the

USEPA (adapted from ATSDR. 1995).

23

Benzo(a)pyrene (B(a)P) (Fig.1.6) is one such priority PAH and it has also been

classed as one of 33 priority pollutants by the European Union (annex II of the

Directive 2008/105/EC). The impact of B(a)P has been investigated on a range

of aquatic organisms by many researchers, these include: flounder,

Paralichthys olivaceus (Woo el al., 2006), crab, Carcinus aestuarii (Fossi et al.,

2000), the marine annelid Platynereis dumerilii (Jha et al.. 1996), scallop,

Chlamys farreri (Pan et al., 2008), zebra mussel Dreissena polymorpha (Binelli

et al., 2008), green-lipped mussel Perna viridis (Ching et al.. 2001; Siu et al.,

2004; Fang et al., 2008; Siu et al., 2008) and Mediterranean mussel Mytilus

galloprovincialis (Venier et al.. 1997; Canova et al., 1998; Gomez-Mendikute et

al., 2002; Marigomez and Baybay-Villacorta, 2003). These studies have

encompassed a wide range of assays which operate at different levels of

biological organisation, from whole organism level (e.g. heart rate (Fossi et al.,

2000)), to cellular level (NRR (Gomez-Mendikute et al.. 2002; Fang et al.,

2008)), through to sub-cellular level (Mn induction (Venier et al., 1997; Siu et

al., 2004; Siu et al.. 2008) and at DNA level using the Comet assay (Bihari and

Fafandel. 2004; Siu et at., 2004; Siu et al., 2008)).

7.3.3 Heavy metals (i.e. cadmium)

Heavy metals pose an ecological and public health threat due to their toxicity,

persistence and their capacity to accumulate in organisms (Bolognesi et al.,

1999). The phrase 'heavy metal' commonly refers to metals that have a specific

gravity of four or five, but in terms of ecotoxicology it is often used to refer to

metals that have been shown to be of specific environmental concern

(Depledge et al.. 1994).

24

Cadmium (Cd) is one such heavy metal and although it is omnipresent in

seawater, it is of great environmental concern due to its high toxicity, its

industrial production and emissions from the combustion of fossil fuels (Erk et

al., 2005). Cadmium affects cell proliferation, differentiation, apoptosis and other

cellular activities and can cause numerous molecular lesions that could be

relevant to carcinogenesis (Filipic et al., 2006). Many different biomarkers have

been implemented with sea stars and mussels to determine the toxicity of Cd.

Increased levels of metallothioneins (MTs) have been quantified (den Besten et

al., 1989; Temara et al., 1997b; Erk et al., 2005) and reduced cellular immune

response (Coteur et al., 2005b) following Cd exposure in the sea star A.

rubens. Work with mussels that investigated the effects of Cd have

implemented a variety of endpoints, including genotoxicity (Bolognesi et al.,

1999; Emmanouil et al., 2007) and induction of MTs (Bebianno and Langston,

1991; Geret et al.. 2002; Erk etal., 2005).

1.4 Reference toxicants

In order to prove that biomarker responses used in toxicology studies are

sensitive, reliable and reproducible known reference toxicants are used as

positive controls (Jha, 2008). This approach was followed in these present

studies by using two reference agents, namely methyl methanesulphonate

(MMS) and hydrogen peroxide (H2O2).

25

1.4.1 Methyl methanesulphonate (MMS)

Methyl methanesulphonate (MMS) is an example of a direct-acfing genotoxic

compound (Fig. 1.7) which causes DNA lesions that lead to strand breaks MMS

as it acts a methylating agent (Pfuhler and Wolf, 1996).

o o

Fig. 1.7 The structure of the genotoxin methyl methanesulphonate.

Methyl methanesulphonate has been used previously as a reference genotoxin

(or positive control) in studies targefing various mammalian cells, including

Chinese hamster ovary cells (Natarajan et al., 1983) and human lymphocytes

(Andreoll et al.. 1999). This reference genotoxic chemical has also been used

with aquatic invertebrates, for example M. edulis larvae (Jha et al., 2000b;

McFadzen et al., 2000; Hagger et al.. 2005b).

1.5.1 Hydrogen peroxide

Hydrogen peroxide (H2O2) (Fig. 1.8) is often used as a model reference

compound in toxicology studies. It has previously been used specifically to

validate the Comet assay (Singh et al., 1988; Singh et al., 1991; Collins et al.,

1995; Lee et al., 1996; Collins et al., 1997; Horvathova et al.. 1998;

Mitchelmore et al., 1998; Andreoli et al., 1999; Bombail et al., 2001) and as a

reference cytotoxin (Richter-Landsberg and Vollgraf, 1998).

26

H H

Fig. 1.8 The structure of hydrogen peroxide.

1.5 Use of echinoderms in toxicology and ecotoxicology

Echinoderms are closely positioned to the chordates on the deuteroslome line

of development of the Animal Kingdom. As can be seen in Figure 1.9,

echinoderms are more closely related to the chordates than the other

invertebrate phyla commonly investigated in aquatic ecotoxicology. The position

of all animal phyla on the classical evolutionary tree has been based on

anatomical features of adult animals and on specific features of their embryonic

development. The higher metazoans are separated into two very distinct

groups, the protostomes (consisting of the arthropods, annelids and molluscs)

and the deuterostomes (the echinoderms and chordates) (Campbell and

Reece, 2002). These two groups are divided by fundamental differences that

occur at three different stages during embryonic development. These

differences in protostome and deuterostome early development occur at the

following stages: during cleavage, the fate of the blastopore and the formation

of the coelom (Moore, 2001).

27

fruit fly nematode

ro 00

Ecdysozoa

iea urchin oscidian mouse mollusc annelid

\/ Lophotrochozoa Hemichordata Echinodermata Urochordata Cephalochordata Vertebrata

I

human

Protoslomia

Chordata

Deuterostomia

Bilateria

Fig. 1.9 Phylogenetic tree showing the closeness of echinoderms to the chordates (vertebrates). The organisms along the top

have specific genome projects currently initiated or completed. (From Sodergren et al., 2006).

Modern molecular techniques have been used to create different animal trees

from those based on traditional comparative anatomy and embryology.

How/ever, the close relationship between the echinoderms and chordates

remains unchanged; it is only the position of some of the other phyla which has

altered. Work carried to investigate the similarity between the echinoid immune

system and the phylogenetic occurrence of immune mechanisms in primitive

chordates further reinforces this evolutionary closeness between echinoderms

and chordates (Smith and Davidson, 1992). Regardless of which criteria are

used to show the relationships of the phyla in the Animal Kingdom, the

echinoderms and chordates show phylogenetic 'closeness'. Due to the similar

development pattern that the two deuterostome phyla exhibit, echinoderms are

very useful in ecotoxicological studies investigating the effects of chemicals on

the development of organisms (Roepke et al., 2005). The evolutionary

closeness of the echinoderms to vertebrates makes them very useful as

toxicologlcal test organisms, meaning that they could potentially replace the

vertebrates commonly used In aquatic toxicology testing (e.g. fish). This would

be in line with the three Rs (Replacement, Refinement and Reduction)

approach to vertebrate testing, as echinoderms could take the place of the

protected vertebrate species. This could be especially Important as industrial

companies have to conform to new European Union regulations (Regulation

(EC) No 1907/2006) concerning the Registration, Evaluation, Authorisation and

restriction of Chemicals (REACH) which came into force on 1st June 2007.

These regulations mean that it is necessary to conduct toxicological tests on

thousands of chemicals with the aim of determining their potential impact on

humans and the environment.

29

Table 1.2 Examples of biomarkers used with the adult stages of echinoderm species belonging to the three echinoderm classes (Asteroidea, Echinoidea and Ophiuroidea) utilised in the present studies.

C l a s s Species Biomarkers used References

Asteroidea Asterias rubens Heat shock protein induction Cytochromes P450 induction

Metallothionein induction Comet assay Phagocytosis

Neutral red retention Righting behaviour

(Oweson et al., 2008) (den Besten et a!., 1993; Everaarts et al., 1998; Danis eta l . , 2004a) (den Besten et a!., 1989; Temara et al., 1997b) (Everaarts, 1995; Everaarts, 1997) (Coteur et al., 2002; Coteur et al., 2005a; Coteur eta l . , 2005b) (Leaney, 2003) (Davies et al., 1998)

Asterina gibbosa Micronucleus induction Comet assay Neutral red retention Righting behaviour

(Leaney, 2003) (Leaney, 2003) (Bowett, 2002; Leaney. 2003) (Bowett, 2002; Leaney, 2003)

Coscinasterias muricata Cytochrome P450 Induction Prey localisation behaviour

(Georgiades eta l . . 2006) (Georgiades eta l . . 2003)

Leptasterias polaris Phagocytosis Neutral red retention

(Bekri and Pelletier, 2004) (Bekri and Pelletier, 2004)

Marthasterias glacialis Cytochrome P450 induction Comet assay Neutral red retention

(den Besten et al., 1990) (Leaney, 2003) (Leaney, 2003)

Patiriella exigua Contaminant avoidance behaviour (Ryder et al., 2004)

Echinoidea Diadmea antillarum Locomotion and spine orientation (Bielmyer et al., 2005)

o

CO

Echinus esculentus Cytochrome P450 induction (den Besten et al., 1990) Paracentrotus lividus Heat shock protein induction (Matranga eta l . , 2000)

Reactive oxygen species production (Coteuretal . , 2001) Stronglycocentrotus Comet assay (Tabaneta l . , 2004) droebachiensis Righting behaviour (Axlak and Saliba, 1981)

Ophiuroidea Amphiura chiajei Lysosomal integrity (Newton and McKenzie, 1998b) Sub-cuticular bacterial counts (Newton and McKenzie, 1995; Newton and

McKenzie, 1998a) Tissue regeneration (Nilsson. 1999) Burrowing behaviour (Newton and McKenzie, 1998a) Contaminant avoidance behaviour (Newton and McKenzie, 1998b)

Amphiura filiformis Sub-cuticular bacterial counts (Newton and McKenzie, 1995) Tissue regeneration (Nilsson and Skold, 1996; Nilsson, 1999) Contaminant avoidance behaviour (Bowmeret al., 1986)

Microphiopholis graciHima Tissue regeneration (D'Andrea eta l . , 1996) Ophioderma brevispina Tissue regeneration (Walsh eta l . , 1986) Ophiophragmus filograneus Tissue regeneration (Clements eta l . , 1998)

Burrowing behaviour (Clements eta l . , 1998) Ophiothrix fragilis Lysosomal integrity (Newton and McKenzie, 1998b)

Sub-cuticular bacterial counts (Newton and McKenzie. 1995) Righting behaviour (Newton and McKenzie, 1998b)

Ophiothrix ophiura Sub-cuticular bacterial counts (Newton and McKenzie, 1995) Feeding behaviour (Valentincic, 1991a; Valentincic, 1991b)

1.5.1 Class Asteroidea and their use in ecotoxicology

Asteroids (sea stars) have a recent history of use in ecotoxicological studies (see Table 1.2). Elevated metallothionein (proteins which can bind to heavy metals) levels have been found in A. rubens following Cd exposure (Temara et al., 1997b). The regeneration of amputated ambulacral spines (by the process of skeletogenesis) was impaired following exposure to lead (Pb) in the same species (Temara et al., 1997a). Biomonitoring studies involving asteroids have found evidence that they can accumulate various toxins and pollutants (a useful attribute in a sentinel organism). Asterias rubens is a valuable indicator of temporal and spatial trends of Pb and Cd contamination in the field. It is also possible to differentiate between evidence of long-term (by sampling skeletal tissue) and short-term exposure (by sampling the pyloric caeca) (Temara et al., 1998). Asterias rubens was also found to have accumulated Pb and Cd, along with zinc, copper and seven different PCBs when collected from a polluted region near the Belgian coast (Danis et al., 2004b). Other chemicals which have been shown to be bioaccumulated in field-collected asteroids include: radionuclides (Warnau et al.. 1999); triphenyltin (TPT) and TBT (Shim et al.. 2005). Longer term 'evidence' of the exposure was commonly found in the calcite-rich skeletal tissue. It is thought that toxins are partitioned in this tissue to detoxify them, as once these chemicals are in the calcite skeleton they are metabolically unavailable (Temara et al., 1997a; Temara et al., 1997b; Warnau et al., 1999; Boisson et al., 2002). The use of A. rubens larvae in ecotoxicological studies has been very limited, but exposure to the naturally occurring aldehyde 2, 4-decadienal reduced the survival rate of the larvae (Caldwell et al., 2004).

32

1.5.2 Class Echinoidea and their use in ecotoxicology

Echinoids (sea urchins) can occur at very high densities on intertidal rocky sites;

up to 100 individuals per (Prince, 1995), and as a result of their grazing

behaviour, are capable of affecting the community structure of benthic

environments, especially deep-water communities (Valentine el al., 2000). Sea

urchins have a relatively long life span, for example, Paracentrotus lividus can live

for up to nine years and remains relatively sessile during this time (Crook et al.,

2000). This longevity, coupled with a sessile lifestyle, means that adult urchins

would make useful sentinel organisms in monitoring marine pollution. One such

example of this is the use of the sea urchin Psammechinus miliaris in detecting

the presence of aquacufture antibiotics, such as oxytetracycline, in the water

column in the vicinity of fish farms (Campbell et al.. 2001). The densities of certain

echinoid species can also be used as indicators of the 'health' of a coral reef

ecosystem. Following events such as El Niiio in Northern Bahia, north-eastern

Brazil, the number of opportunistic sea urchin species {Diadems anttiHarum and

Echinometra lacunter) increased at the expense of overall species diversity (Altrill

e ta l . , 2004).

The use of sea urchins in ecotoxicology has been limited mainly to the use of their

gametes and embryos in developmental and cellular research which commenced

in the mid-1800s (Monroy, 1986). Gametes are useful toxicology tools as they are

very small, meaning that they have high surface area to volume ratios which

enables rapid toxicant interaction. Also, during spawning and fertilization a rapid

series of events occur that can be used as endpoints in short exposure periods

(these events include: sperm activation and motility, egg fertilization and protein

33

synthesis) (Dinnel et al., 1988). Originally, sea urchin gametes and embryos were

used to determine the adverse effects of metals on fertilization and development

of embryos as early as the 1920s and 1930s (Dinnel et al. , 1988). This work

continued and eventually these 'simple' observations were linked to the toxic

effects of metals, and other pollutants, in the natural environment (Wilson and

Armstrong, 1961). This led to suggestions that sea urchin embryos could be used

to assess water quality in the environmenL

A large body of work on the adverse effects of toxins on the development of sea

urchin embryos has been carried out. The range of chemicals and materials which

have been found to adversely effect larval development of urchins include: arsenic

(Pagano et al., 1982); chromium (Pagano et al., 1983a); diphenyl and diphenyl

ether (Pagano et al., 1983b); effluents (Pagano et al., 1993); the hormone

thyroxine (Johnson, 1998); the metals arsenate, nickel and chromate (Garman et

al., 1997); estradiol and endocrine disrupting compounds (Roepke et al., 2005)

and copper (Rosen et al., 2005). The use of echinoids in regulatory toxicity testing

has focused mainly on sea urchins and sand dollars, especially in testing carried

out by the American Society for Testing Materials (ASTM) and the United States

Environmental Protection Agency (USEPA). The use of the larval stages of

echinoids are preferred to adults due to the relative sensitivity of these early life

stages, meaning that toxicity screens can be carried out in a short space of time.

Sea urchin sperm tests can be carried out following exposure times as short as 10

to 60 min in order to determine sediment and effluent toxicity and to carry-out

toxicity identification evaluations (TIEs). Embryo tests normally have a 48 to 96 h

34

exposure lime and toxicity is determined by quantifying the frequency of embryo

malformations (Bay el al., 1993).

The use of adult echinoids in toxicology has been somewhat limited (see Table

1.2), when compared to the widespread use of their embryo-larval stages. This is

despite the fact that adult echinoids are easy to collect, are robust to laboratory

conditions and are often ecologically important in the areas where they are located

(Pearse, 2006). The relatively large size and robustness of the adults makes them

unsuitable for rapid screening assays (where gamete and embryo tests are better

suited). But, adults could be used to carry out longer term chronic tests to evaluate

the long-term effects of chronic pollution on growth and reproduction (Bay et al.,

1993). Behavioural endpoints have been used previously to assess the health of

adult echinoids. The righting behaviour of adult sea urchins (Paracentrotus lividus)

was found to be suppressed by topical (applied directly onto the outer dermis)

exposure to crude oil (Axlak and Saliba. 1981). The role of adult echinoids in

research and regulatory ecotoxicology is likely to become much more prominent

as a result of the recent sequencing of the sea urchin genome in 2006. The

complete genome (which encodes about 23,300 genes) of the California purple

sea urchin, Strongylocentrotus purpuratus was sequenced primarily due to the

usefulness of the urchin embryo as a model for modern molecular, evolutionary

and cellular biology (Sodergren et al., 2006). To date, this sea urchin genome is

the first non-chordate deuterostome genome to be fully sequenced (Davidson,

2006).

35

7 .5.3 Class Ophiuroidea and their use in ecotoxicology

In common with asteroids and echinoids, the ophluroids (the brittle stars) have

also been the subject of toxicological studies (see Table 1.2). The brittle star

Ophiothrix fragilis has been found to accumulate various radioisotopes (Hutchins

et al., 1996). The reproductive stage of ophluroids has been found to influence the

relative toxicology of different toxins (Bowmer et al. , 1986). The infaunal ophiuroid

Amphiura filiformis was more sensitive to both copper and pentachlorophenol

when the test animals were gravid (bearing sperm or eggs), compared to

individuals which had shed their gametes (Bowmer et al., 1986). The degree of

PCB accumulation in ophluroids can be influenced by the lipid content in the

animals. In the brittle stars A. filiformis and A. chiajei, PCBs were accumulated in

the disks of the animals (a site of high lipid content), but not in the arms (where

lipid levels were negligible) (Gunnarsson and Skold, 1999).

One Important ecological aspect of the capacity of ophiurolds to accumulate toxins

is the degree to which the animals are predated upon. Predalion on Infaunal

brittle stars may occur by the consumption of the whole animal, but more

commonly by the cropping of exposed arms by fish and crustaceans. Arm loss,

and subsequent regeneration, may occur in 50 % of a given population (Skold,

1998). Accumulation of PCBs by infaunal organisms has been shown to be the

first step in their transfer to higher trophic levels. Amphiura spp. arms were found

to account for 60 % of the stomach contents of the dab, Limanda limanda, which

is frequently consumed by humans. In this manner a trophic transfer of toxicants

can occur indirectly through the brittle stars (Skold, 1998). This involvement of

ophiuroids in the trophic transfer of toxins makes them very desirable organisms

36

for selection in aquatic toxicology studies. If certain ophiuroid species can be

shown to accumulate toxins, then the possible effects of this bioaccumulation al

higher levels in the trophic ladder can be addressed.

1.6 Aims and objectives

Against the backdrop of the above information, the overall aim of this research

programme was to implement a suite of non-destructive biomarkers which

operated al different levels of biological organisation on a range of adult

echinoderm species to evaluate the effects of a range of reference pollutants

under laboratory conditions.

The hypotheses to be tested were:

1) Echinoderms are more sensitive at different levels of biological organisation

following exposure to reference toxicants than molluscs.

2) Toxic responses a l lower levels of biological organisation are linked to, or

expressed at, higher level responses.

3) Different echinoderm species show different responses following exposure

to an environmentally relevant PAH (e.g. B(a)P).

37

4) Asteroid echinoderms and bivalve molluscs have differing DNA repair

capacities (as determined by the Comet assay).

5) Immunotoxic responses are linked to DNA damage and the DNA repair

capacity of the organisms.

6) Asteroids and bivalve molluscs show differential tissue specific

accumulation patterns for heavy metals (i.e. Cd).

Specifically, the aims and objectives were:

1) To devise, optimise and validate appropriate biomarkers (at different levels

of biological organisation) in a number of echinoderm species.

2) To determine the relative sensitivity of asteroids and bivalve molluscs to

known, or reference, genotoxins (Chapter 2).

3) To investigate whether any genotoxic responses can be linked to cytotoxic

and behavioural (or physiological) consequences (Chapter 2).

4) To determine the responses of three different echinoderms (sea stars, sea

urchin and brittle stars) and molluscs (blue mussel) to an environmentally

relevant PAH (Chapter 3).

38

5) To investigate the behavioural (physiological), cytotoxic, immunotoxic.

genotoxic and oxidative stress consequences in echinoderms {Asterias

rubens) and molluscs (Mytilus edulis) to a heavy metal (Chapter 4)

6) To determine whether any immunotoxic, genotoxic and oxidative stress

effects are reversible (Chapter 4).

A schematic outline illustrating the objectives of this PhD thesis is given in Figure

1.10.

39

Chapter 1 Introduction and Aims

Chapter 2 Genera! Materials and Methods

Chapter 3 Reference genotoxin effects on

A. rubens and M. edulis

J

Chapter 4 PAH effects on A. rubens, P. lividus,

O. fragilis and M. edulis

Chapter 5 Metal effects on A. rubens and M. eduh's

r Chapter 6 Discussion and conclusions

Recommendations for future work

Fig. 1.10 Outline of the thesis, detailing the topics covered in each chapter.

40

Chapter 2

Materials and methods

The results of part of this chapter were presented as a poster a l the 7'^

International Comet Assay Workshop, Coleraine, 2007 (Canty el al., 2007.

Mutagenesis 22, e1-e21).

41

2. Materials and methods

2.1 Test organism collection and maintenance

2.7.7 Collection and maintenance of common sea star, Asterias rubens

(Echinodermata: Asteroidea)

Adult common sea stars, Asterias rubens (Fig. 2.1a) were collected from In and

around Plymouth Sound. Devon, UK (Latitude 50.20° N; Longitude 4.10° W) (Fig.

2.2a), using SCUBA divers and trawling. Only intact, f ive-armed, sea stars were

returned to the laboratory. All starfish were 66 - 100 mm In diameter. They were

held in 20 L glass aquaria equipped with air-driven filters until required (Fig. 2.3).

The holding conditions were 15 °C ± 1.5 °C, 34 ± 1.5 PSU, with a light: dark cycle

of 12h: 12h. Water changes (50 % ) were carried out twice weekly. Any animal

which appeared stressed or unhealthy was immediately transferred into separate

holding aquaria to prevent mortalities adversely affecting the water quality. Prior to

use, A. rubens were acclimatised for a minimum of two weeks and were fed

mussels ad libitum.

2.1.2 Collection and maintenance of purple sea urchin, Paracentrotus lividus

(Echinodermata: Echinoidea)

Adult purple sea urchins (Fig. 2.1b) Paracentrotus lividus were purchased from

Dunmanus Seafoods Ltd (Durrus. Bantry, Co. Cork, Eire) (Latitude 51.41° N;

Longitude 9.28°W) (Fig. 2.2b). Upon arrival any damaged urchins (spines missing,

or broken) were immediately disposed of and the healthy animals were transferred

to glass aquaria under the same conditions as the sea stars. All sea urchins were

70 - 95 mm in test diameter (including the spines). They were fed a commercially

42

available algal pellet (Telra Variety Wafers, Tetra, UK) ad libitum and held for a

minimum of two weeks prior to use.

2.1.3 Collection and maintenance of common brittle star Ophiothrix fragilis

(Echinodermata: Ophiuroidea)

Common brittle stars (Fig. 2.1c), Ophiothrix fragilis were collected from in and

around Plymouth Sound, Plymouth, Devon, UK (Latitude 50.20° N; Longitude

4.10° W) (Fig. 2.2a) by trawling. Only intact, f ive-armed, brittle stars were returned

to the laboratory. All brittle stars were 40 - 65 mm In diameter. They were held in

glass aquaria under the same conditions as the sea stars with the exception that

they were fed a suspension of Isochrysis galbana algae (Varicon Aqua Solutions

Ltd., Malvern, UK) and a commercially available algal pellet (Tetra Variety Wafers,

Tetra. UK) ad libitum.

43

Fig. 2.1 The test species used in the studies (a) Asterias rubens (b) Paracentrotus lividus (c) Ophiothrix fragilis (d) Mytilus edulis.

Scale bars = 10 mm.

ii

(b)

cn (C)

Fig. 2.2 An outline of the British Isles and Eire with animal collection sites highlighted (white circles denote collection areas at each

specific site), (a) The area in and around Plymouth Sound (Devon, UK) where sea stars and brittle stars were collected, (b) The

area near Bantry, (Co. Cork, Eire) where sea urchins were collected by Dunmanus Seafoods Ltd (Durrus, Bantry, Co. Cork, Eire),

(c) Port Quin (Cornwall, UK) where mussels were collected. (Satellite images from http://maps.google.co.uk).

/ 1 f m-:

Fig. 2.3 Husbandry stock tanks used to house Asterias rubens. Air-driven filtration

units and Mytilus edulis as food source are visible.

2.1.4 Collection and maintenance of blue mussel Mytilus edulis (Mossuica: Bivalvia)

Adult blue mussels (Fig. 2.1d). Mytilus edulis were collected by hand from Port

Quin. Cornwall. UK (Latitude 50.35° N; Longitude 4.52° W) (Fig. 2.2c). All mussels

were 35 - 45 mm in length; they were cleaned of any epibioants and then placed

into 20 L glass aquaria until required (holding conditions were the same as for the

sea stars) Prior to use. they were acclimatised for a minimum of one week and

they were fed twice weekly with a standard aquarium invertebrate food mixture

(Waterlife Invert Food Liquidfry Mahne^ . Rotl-Rlch " and dried algae).

46

2.2 Haemolymph and coelomic fluid collection

2.2.1 Haemolymph extraction from Mytilus edulis

Haemolymph samples were extracted from the posterior adductor muscle of the

mussels (Fig. 2.4). A pair of scissors was used to slightly prise open the valves

and mantle water was allowed to drain onto absorbent paper. A 1 mL syringe fitted

with a 21 gauge needle was used to extract haemolymph (average volume

0.5 mL) from the adductor muscle (Fig. 2.4). The volume of haemolymph collected

was sufficient to carry-out all the necessary cellular and sub-cellular assays;

therefore it was not diluted with physiological saline. The needle was removed

from the syringe before transferring the sample to a siliconised centrifuge tube (to

prevent damaging the cells) and held on ice prior to use. The samples were

thoroughly mixed by vortexing prior to use. Giemsa stained haemocytes from

M. edulis are shown in Figure 2.5.

47

Fig. 2.4 Haemolymph extraction from the posterior adductor muscle of t\Aytilus

edulis. (a) and (b) show syringe and needle positioned in the posterior adductor

muscle, (c) A syringe containing haemolymph and (d) is a mussel with both valves

opened, showing the position of the posterior adductor muscle (red arrows).

48

Fig. 2.5 Giemsa stained haemocytes from Mytilus edulis haemolymph.

Scale bar = 15 pm.

2 2.2 Coelomic fluid extraction

2 2 2 1 Asterias rubens

A 1 mL syringe fitted with a 21 gauge needle was used to extract coelomocytes

from the inter-arm' region on the sea star (Fig. 2.6). The average volume collected

was 0.6 mL. The needle was removed from the syringe before transferring the

sample to a siliconised Eppendorf tube (to prevent the damaging the cells) and

held on ice prior to use. The samples were thoroughly mixed by vortexing prior to

use.

49

Fig. 2.6 Coelomocyte extraction from the inter-arm' region of >Asfer/as rubens.

Examples of the common cell types found in the coelomic fluid of A. rubens can be

seen in Figure 2.7. These cell types were common across all the echlnoderm

species used in these studies. Phagocytic amoebocytes are the most common

form of echlnoderm coelomocytes and have been found in every species studied

in the past (Smith, 1981) (Fig. 2.7a). For example, In A. rubens phagocytes were

found to constitute about 95 % of the total cell population (Pinsino et al., 2007) and

in Strongylocentrotus purpuratus they made up 66.3 ± 11.6 % (mean ± SD) of total

coelomocytes (Smith et al.. 1992). Historically, spherule cells have been given a

variety of names including morula cells, eleocytes, granulocytes and trephocytes

(Ratcliffe and Rowley. 1979) (Fig. 2.7b). Colourless spherule cells constituted

5.1 ± 4 . 6 % of total coelomocytes in Strongylocentrotus purpuratus, with red

spherule cells accounting for 14.8 ± 8.5 % (Smith et al., 1992).

50

(b)

S

Fig. 2.7 Giemsa stained cells from Astenas rubens coelomic fluid, (a) Phagocytic

amoebocytes (pa) are the most common cell types, fol lowed by (b) spherule cells

(s). Scale bar = 10 pm.

2.2 2.2 Paracentrotus lividus

A 1 mL syringe fitted with a 21 gauge needle was used to extract coelomic fluid

(average volume 0.5 mL) from the peristomal membrane which is located on the

aboral side of the urchin, around the Aristotle's Lantern (the jaw apparatus). The

needle was removed from the syringe before transferring the sample to a

sil lconised Eppendorf tube (to prevent damaging the cells) and held on ice prior to

use. These samples were thoroughly mixed by vortexing prior to use.

2 2 2.3 Ophiothnx fragilis

Difficulties were faced collecting sufficient coelomocytes from the brittle stars to

carry out the cellular and sub-cellular assays. Brittle stars did not contain very

much coelomic fluid in relation to body size, especially when compared to the sea

51

stars and sea urchins. Due to calcite skeleton of the animals, a needle and syringe

could not successfully be used to collect the cells. Instead, an individual brittle star

was placed in a clean Petri dish, a sharp blade was used to amputate an arm from

the central disc and the coelomic fluid was collected. Preliminary investigations

only yielded in the region 5 pL of coelomic fluid per animal, which was too small a

volume with which to carry-out cellular and subcellular assays. Therefore, only

behavioural/ physiological assays (clearance rate - Section 2.9.2 and righting time

- Section 2.10) were carried out with the brittle stars.

2.3 Cell membrane stability (neutral red retention assay)

The cell membrane stability (sometimes referred to as lysosomal membrane

stability) of the haemocytes/coelomocytes was determined by measuring the cells'

ability to retain neutral red (NR) dye (Babich and Borenfreund, 1992). This was

achieved through the implementation of the neutral red retention (NRR) assay.

The same protocol was implemented for all test species. That is. samples of

haemolymph. or coelomic fluid (50 pL) were incubated, in triplicate, for each test

organism, in a flat-bottomed microtitreplate in order to allow a monolayer of cells to

adhere to the wells. The plate had been pre-coated with 10 % (v/v) poly-i-lysine

solution, to aid cellular adhesion. After 50 min. non-adhered cells were discarded

and the plates washed with phosphate buffered saline (PBS) (see Appendix I). A

solution of 0.004 % neutral red dye (Sigma. U.K.) in distilled water (200 pL) (see

Appendix I) was added to each well and the cells incubated at 10 °C for 3 h. The

wells were washed and an acidified solution of 1 % acetic acid and 50 % ethanol

(see Appendix I) was added to resolubilise the dye. The plates were shaken for

52

l O m i n before reading the absorbance at 540 nm, in an Optimax tunable

microplate reader (Molecular Devices, USA), using Softpro Max (V.2.4.1) software.

Protein concentration was then determined following a recognised method, as a

surrogate for cell counts in each well (see Section 2.4). Results were presented as

optical density (at 540 nm) per mg of protein.

2.3.1 Validation of neutral red retention assay (NRR)

The NRR assay was validated in vitro in A. rubens, P. lividus and M. edulis using

hydrogen peroxide (H2O2) as a model toxicant (Richter-Landsberg and Vollgraf,

1998). The experimental procedures for the three species were identical (Fig. 2.8).

Following H2O2 exposure the same protocol as outlined in Section 2.3 was

followed to determine the cell membrane stability of the exposed cells. Briefly,

cells from 5 individual animals were collected and pooled. The cells were placed

onto a microtitre plate (in quintuplicate) and left to incubate. After 45 min excess

fluid was removed and replaced with H2O2 at concentrations of 0, 10, 50 and

100 pM for 10 min (on ice) (concentrations were selected to be similar to those

used by previous researchers (Mitchelmore et al., 1998; Cheung et al., 2006)).

The H2O2 was removed and the cells gently washed twice with PBS (to remove

any residual H2O2), the NRR assay was then carried out.

To interpret the results, statistical significance between means was determined

using one-way analysis of variance (ANOVA) and the data sets were analysed

with Statgraphics v5.1 software (Statsoft, USA). Differences between means which

were P < 0.05 were considered to be significant. The results of the validation study

(Fig. 2.9) showed significant a dose dependent decrease in cell membrane

53

stability across all the exposure concentrations for A. rubens (P < 0.001) and

M. edulis (P < 0.001) and for the two highest concentrations (50 and 100 pM) for

P. / /wc/us(P< 0.001).

Cells collected from 5 individuals and pooled

50 pL pipetted into 20 wells on microtitre plate (as 4 H2O2 concentrations, In

quintuplicate)

Cells incubated in plate for 45 min, then excess fluid removed. 50 pL of H2O2

(at 0, 10. 50, 100 pM) added to each well

H2O2 exposure for 10 min, on ice (in dark)

H2O2 removed, cells washed twice with PBS. NRR assay then carried out as per

Section 2.3

Fig. 2.8 Flow chart illustrating the experimental design for the in vitro validation of

the NRR assay in Asterias rubens, Paracentrotus lividus and Mytilus edulis.

54

I A. rubens I P. lividus I M. edulis

Control 10 50

H2O2 concentration (pM)

100

Fig. 2.9 Neutral red retention expressed as cell membrane stability (optical density

at 550 nm mg'^ protein) following in vitro exposure to H2O2 for Asterias rubens,

Paracentrotus lividus and Mytilus edulis. Concentrations are expressed in pM. The

values are means (n = 5) ± SEM. * indicates a significant difference from controls

at, or above, the 95 % confidence level.

2.4 Protein determination

Protein concentration was determined following a recognised method (Bradford,

1976). It was necessary to determine protein content in haemolymph coelomic

fluid following the NRR (Section 2.3) and phagocytosis (Section 2.5) assays as

protein concentration acted as a surrogate for cell concentrations in the wells of

55

the microtitre plate. Results from both the NRR and phagocytosis assays are

divided by protein concentration to normalise for cell number/ concentration in

each well of the microlitre plate. Protein concentration in haemolymph/ coelomic

fluid was determined using the Pierce bicinchoninic acid (BCA) reagent. Protein

standard solutions (0.1 - 2.0 mg mL'^) were prepared in distilled water from bovine

serum albumin stock solution (2.0 mg mL^) (Sigma, U.K.) (see Appendix I). BCA

reagent, 200 pL, was added to 10 pL of haemolymph/ coelomic fluid, along with

10 pL of each protein standard and incubated at 37 °C for 30 min. Absorbance

was determined spectrophotometrically at 550 nm in an Optimax tunable

microplate reader (Molecular Devices, USA), using Softpro Max (\/.2.4.1) software.

2.5 Immunocompetence (phagocytosis a s s a y )

The phagocytotic activity of haemocytes and coelomocytes was determined by

measuring the uptake of zymosan particles (from yeast, Saccharomyces

cerevisiae) dyed with neutral red dye (Pipe et al., 1995; Canty et al., 2007;

Dissanayake et al., 2008; Hannam et al., 2009b). The same procedure was

followed for both A. rubens coelomocyte and M. edulis haemocyte samples. That

is, haemolymph/ coelomic fluid (50 pL) was pipetted in triplicate onto a microtitre

plate (pre-coated with 10 % (v/v) poly-L-lysine solution) and left for 50 min to allow

the cells at adhere. Following incubation, non-adhered cells were removed and

50 pL of dyed zymosan suspension (50 x 10^ particles mL"^) was added and

incubated with the cells for 30 min at 10 °C. The reaction was stopped by adding

100 pL of Baker's Formol Calcium (see Appendix I) which fixed the cells. The plate

56

was cenlrifuged (70 x g) for 5 min, the supernatant removed and the cells washed

twice with 100 of PBS (see Appendix I). Acidified ethanol (1 % acetic acid and

50 % ethanol in distilled water) (see Appendix I), 100 pL, was added to resolubilise

the dye. The phagocytic uptake of the dyed zymosan particles was determined

spectrophotometrically (at 550 nm) against a standard curve of zymosan

suspensions (50 - 1.56 x 10^ parlicles mL"^). Protein concentration was

determined (Section 2.4) and the results were presented as absorbance of stained

zymosan particles phagocytosed per mg of protein.

2.5.1 Validation of phagocytosis assay

The phagocytosis assay was validated in vitro in A. rubens and M.edulis using

H2O2 as a model toxicant. The experimental procedure for the two different

species was identical (Fig. 2.10). Following H2O2 exposure the same protocol as

outlined in Section 2.5 was followed to determine the phagocytotic activity of the

exposed cells. Briefly, cells from four individuals were collected and pipetted onto

a microtitre plate in quadruplicate. Following 45 min incubation, excess fluid was

replaced with 50 pL of H2O2 (at 0. 10. 50 and 100 pM).

To interpret the results, statistical significance between means was determined

using one-way analysis of variance (ANOVA) and the data sets were analysed

with Statgraphics v5.1 software (Statsoft, USA). Differences between means which

were P < 0.05 were considered to be significant. The results (Fig. 2.11) showed

significant a dose dependent decrease in cell membrane stability across all the

exposure concentrations for both A. rubens (P < 0.001) and M. edulis {P = 0.036).

57

Cell samples collected from 4 individuals and pooled

50 pL pipetted into 16 wells on microtitre plate (as 4 H2O2 concentrations, in

quadruplicate)

Cells incubated in plate for 45 min. then excess fluid removed. 50 pL of H2O2

(at 0, 10, 50, 100 pM) added to each well

H2O2 exposure for 10 min. on ice (in the dark)

H2O2 removed, cells washed twice with PBS. Phagocytosis assay then carried out

as per Section 2.5

Fig. 2.10 Flow chart illustrating the experimental design for the in vitro validation of

the phagocytosis assay in Asterias rubens and Mytilus edulis.

58

CU CO

o o ^

"5 05

CO 0

2 Q.

CU Q . O

o E >^

N

2.5

2.0

1.5 H

1.0

0.5

0.0

I i

I ' rubens

X

Control 10 50

H2O2 c o n c e n t r a t i o n ( | j M )

1

100

Fig. 2.11 P h a g o c y l o l i c act iv i ty e x p r e s s e d as z y m o s a n par t ic les p h a g o c y t o s e d

(x 10^ m g prote in '^ ) fo l low ing in vitro e x p o s u r e to H2O2 for Asterias rubens a n d

Mytilus edulis. C o n c e n t r a t i o n s a re e x p r e s s e d in p M . T h e va lues a r e m e a n s (n = 4 )

± S E M . * ind ica tes a s ign i f icant d i f f e rence f rom con t ro ls at . or a b o v e , the 9 5 %

c o n f i d e n c e leve l .

2.6 Micronuclei induction

Mic ronuc le i ( M n ) a re c a u s e d by bo th c h r o m o s o m e b r e a k a g e s a n d sp i nd l e

a p p a r a t u s dys func t i on ( c h r o m o s o m e s fai l ing to cor rec t ly s e g r e g a t e at mi tos is ) . T h e

m i c r o n u c l e u s a s s a y d e t e r m i n e s the a b u n d a n c e of t hese M n a n d , the re fo re ,

59

eva lua tes c h r o m o s o m a l d a m a g e . M n are morpho log i ca l l y ident ica l to , but smal ler

t han nuc le i ( e x a m p l e s of Mn are s h o w n in F ig . 2 .12) . T h e y h a v e the fo l lowing

a t t r ibu tes :

I. M n normal ly have the s a m e s ta in ing in tensi ty as m a i n nuc le i , but they m a y

b e m o r e in tensely s t a i n e d .

II. M n are not phys ica l ly l inked to, or ove r l ap wi th the m a i n nuc le i .

III. M n m a y not over lap .

IV. A s Mn are non- re f rac t i le they c a n eas i ly b e d i f fe ren t ia ted f r o m a n y ar te fac ts ,

such as s ta in ing par t ic les ( F e n e c h , 2 0 0 0 ) .

T o p repa re M n s l ides, 2 0 0 pL of c o e l o m i c f lu id / h a e m o l y m p h w a s gen t l y sp read

on to a m i c r o s c o p e s l ide w h i c h had b e e n p r e - c o a t e d wi th 10 % (v /v ) poly-L- tys ine

so lu t ion ( to aid cel lu lar a d h e s i o n ) . T h e s l ides w e r e left for 30 m i n ; e x c e s s f luid

d ra ined and the s l ides left to air dry . T h e s l ides w e r e then f ixed in m e t h a n o l for

15 m i n , fo l l owed by s ta in ing w i th G i e m s a sta in (5 % (v /v) so lu t ion ) , for 20 m in .

Excess s ta in w a s r e m o v e d by r ins ing tw ice in dist i l led w a t e r a n d o n c e the s l ides

had air d r ied a covers l ip w a s m o u n t e d us ing D P X . T h e s l ides w e r e c o d e d and

r a n d o m i s e d , then e x a m i n e d unde r the m i c r o s c o p e for the induc t ion of M n . O n e

t h o u s a n d cel ls f rom e a c h s l ide w e r e sco red a c c o r d i n g to the cr i ter ia desc r i bed

p rev ious ly (Ven ier et aL, 1997; F e n e c h , 2 0 0 0 ; J h a et a l . , 2 0 0 5 ) . E x a m p l e s of M n

p resen t in M. edulis h a e m o c y t e s a n d in A. rubens c o e l o m o c y t e s a re s h o w n in

F igure 2 .13 .

60

(a) (b) (c)

Fig . 2 .12 T h e a p p e a r a n c e of s o m e typ ica l M n . (a) A cel l w i t h o n e M n p resen t , (b)

T w o M n p resen t , one c lose to . bu t no t o v e r l a p p i n g the m a i n nuc le i , (c) A ce l l w i th

two M n w h i c h a re c lose to o n e ano the r , bu t no t o v e r l a p p i n g .

(a)

Fig . 2 .13 E x a m p l e s of M n p r e s e n t in (a) a Mytilus edulis h a e m o c y t e a n d in (b) a n

Asterias rubens c o e l o m o c y t e . T h e M n are s h o w n by t h e a r r o w s . Sca le

ba r = 10 p m .

61

2.7 Cellular viability (Eosin Y staining)

Prior to carrying out the Comet assay (Section 2.8) the viability of the cells (i.e.

coelomocytes or haemocytes) were determined by Eosin Y staining. This assay is

based on the principal that healthy cells with intact plasma membranes can

exclude Eosin Y, but injured cells will take up the dye (Lowe and Pipe, 1994). A

stock solution of 2 mg mL'^ of Eosin Y stain (Sigma, UK) was prepared in

physiological saline (Appendix I). A 40 pL sample of coelomic fluid/ haemolymph

was pipetted into an Eppendorf and 2 pL of Eosin Y solution was added. The tube

was gently vortexed and the contents transferred to a haemocytometer and the

cells observed under a light microscope (total magnification x 400). A minimum of

100 cells were scored as being either alive (both colourless and clear, or with a

slight green colouration) or dead (distinctly stained red) (Fig. 2.14). Only cell

samples with a cellular viability of > 70 % were used in the Comet assay (Tice et

al., 2000; Jha et al. .2005).

(a) (b)

Fig. 2.14 Mussel haemocytes stained with Eosin Y. The cell on the left (a) is viable

(green), the cell on the right (b) is unviable/ dead (red). Scale bar = 10 pm.

62

2.8 The Comet a s s a y

T h e a lka l ine C o m e t a s s a y w a s a d a p t e d f r o m a p ro toco l desc r i bed by S ingh et a l .

( 1 9 9 9 ) a n d a d a p t e d for m e a s u r i n g geno tox i c i t y in aqua t i c o r g a n i s m s by m a n y

resea rche rs (D ixon et a l . . 2002 ; Jha . 2 0 0 8 ; Frenzi l l i et a l . . 2 0 0 9 ) . Br ief ly , s l ides

w e r e p re - coa ted w i th n o r m a l mel t ing po in t ( N M P ) a g a r o s e ( s e e A p p e n d i x I).

C o e l o m i c f luid or h a e m o l y m p h , 100 \JL, w a s cen t r i fuged at 200 x g for m in a n d the

supe rna tan t r e m o v e d . T h e cel ls w e r e r e - s u s p e n d e d w i th 200 pL of low me l t ing

point ( L M P ) a g a r o s e ( see Append i x I) a n d two s e p a r a t e 85 pL v o l u m e s of the L M P

a g a r o s e / cel l s u s p e n s i o n w e r e p laced o n the N M P a g a r o s e coa ted s l ide. T h e

s l ides w e r e left at 4 °C to a l low the L M P a g a r o s e to set. T h e s l ides w e r e i ncuba ted

in lysis so lu t ion ( see A p p e n d i x I) for 1 h in the dark at 4 °C to b r e a k d o w n the

cel lu lar m e m b r a n e s . T h e s l ides w e r e w a s h e d in d ist i l led wa te r 3 t imes for 2 m in

and p laced in a hor izonta l ge l e lec t rophores is c h a m b e r . T h e c h a m b e r w a s f i l led

wi th e lec t ropho res i s buf fer ( s e e Append i x I) a n d the s l ides left for 2 0 min ( to a l low

t ime for the D N A to u n w i n d ) pr ior to e lec t rophores i s , w h i c h w a s car r ied out at 20 V

for 30 m i n . T h e e lec t rophores is p rocess fo rced the D N A f r a g m e n t s a w a y f rom the

nuc leo id co res . Af ter e lec t rophores is the s l ides w e r e s u b m e r g e d in ch i l led

neu t ra l i s ing buf fer for 10 min (see A p p e n d i x I). T h e rep l ica te m ic roge ls o n the

s l ides w e r e e a c h s ta ined wi th 30 \JL of e th id ium b r o m i d e so lu t ion (a DNA-spec i f i c

f l uo rescen t d y e ) (see A p p e n d i x I) and sco red unde r an ep i f l uo resence m i c r o s c o p e

(Le ica , D M R ) us ing the Kome t 5.0 image -ana l ys i s so f twa re (K ine t i c Imag ing ,

L ive rpoo l , U K ) ( e x a m p l e 'Come t ' images a re s h o w n in Fig. 2 .15) . T h e s l ides w e r e

c o d e d a n d r a n d o m i s e d a n d 50 cel ls w e r e s c o r e d per rep l ica te ( the re fo re 100 cel ls

per s l ide) . P e r c e n t a g e tail D N A is the p a r a m e t e r p resen ted in this s tudy as a

m e a s u r e of s ing le -s t rand D N A b reaks , a s it is t hough t to be the m o s t re l iab le

63

p a r a m e t e r a n d it e n a b l e s in te r - l abora to ry r e f e r e n c e (Co l l i ns . 2 0 0 4 ; K u m a r a v e l a n d

J h a . 2 0 0 6 ) . T h e s a m e e x p e r i m e n t a l p r o c e d u r e w a s f o l l o w e d for al l of t h e tes t

s p e c i e s i nves t i ga ted a n d th is p ro toco l w a s s u m m a r i s e d in F igu re 1.3.

(a) ( b )

(c)

•

Fig. 2 .15 Mytilus edulis h a e m o c y t e s i l lus t ra t ing i nc reas ing leve ls o f D N A d a m a g e

as d e t e c t e d by the C o m e t assay , w i th (a) r e p r e s e n t i n g u n d a m a g e d ce l ls , t h r o u g h

to (d) exh ib i t i ng the d a m a g e ( i .e. the la rges t ' C o m e t ta i l ' ) . S c a l e b a r = 50 p m .

2 8 1 The modified Comet assay

Fo l l ow ing the lysis s t age of the C o m e t a s s a y an e n z y m e d iges t i on s t e p w a s

ca r r ied ou t to ta rge t ox i d i sed b a s e s (Co l l i ns et a l . . 1 9 9 3 ; D u s i n s k ^ a n d Co l l i ns ,

1996 ; R e e v e s et a l . . 2 0 0 8 ) . T h e e n z y m e s u s e d w e r e f o r m a m i d o p y r i m i d i n e D N A

g l y c o s y l a s e ( F P G ) a n d e n d o n u c l e a s e III ( E n d o III). T h e e n z y m e s conve r t o x i d i s e d

p u r i n e s ( a d e n i n e a n d g u a n i n e ) a n d p y r i m i d i n e s ( cy tos ine a n d t h y m i n e ) ,

respec t i ve ly , in to D N A s ing le s t rand b r e a k s . T h r e e s l i des w e r e p r e p a r e d fo r e a c h

64

o r g a n i s m ; o n e e n z y m e buf fer cont ro l (no e n z y m e t rea tmen t ) , one s l ide w a s t rea ted

wi th E n d o III a n d the third s l ide t reated w i th F P G . Fo l l ow ing lysis the s l ides w e r e

w a s h e d th ree t imes in e n z y m e react ion bu f fe r ( s e e A p p e n d i x I) at 4 °C . T h e n

50 pL of e i ther e n z y m e reac t ion buf fer ( r e fe rence s l ides) or 1.5 pg m L ' ^ of F P G , or

E n d o II I , w a s app l i ed to e a c h ge l , a long w i th a covers l ip , pr ior to i ncuba t i on at

37 °C for 4 5 m i n in a humid i t y chamber . Immed ia te l y a f ter e n z y m e d iges t ion the

s l ides w e r e p laced in a n e lec t rophores is c h a m b e r as per Sec t ion 2.8. T h e C o m e t

sco res of the e n z y m e - t r e a t e d s l ides w e r e c o m p a r e d to app rop r ia te re fe rence

s l ides ( t hose t rea ted w i th e n z y m e react ion buf fer w i th no e n z y m e ) . T h e s a m e

expe r imen ta l p r o c e d u r e w a s fo l lowed for all the test spec ies Invest iga ted a n d this

pro toco l w a s s u m m a r i s e d in F igure 1.3.

2.8.1.1 Validation of modified Comet assay

T h e mod i f i ed C o m e t a s s a y w a s va l ida ted in vitro in A. rubens, P. lividus a n d

M.edulis us i ng H2O2 as a mode l tox icant (Co l l ins et a l . , 1995 ; Co l l ins et a l . , 1997 ;

Andreo l l et a l . . 1999 ; G ie lazyn et a l . . 2003 ; C h e u n g et a l . , 2 0 0 6 ) . T h e expe r imen ta l

p rocedu re for the th ree d i f ferent spec ies w a s ident ica l (F ig . 2 .16) . Fo l l ow ing the

H2O2 e x p o s u r e s tep the s a m e protoco l as ou t l i ned in Sec t i on 2.7.1 w a s fo l l owed to

d e t e r m i n e the ox ida t i ve D N A d a m a g e to the e x p o s e d cel ls. T h e stat is t ica l

s ign i f i cance b e t w e e n the m e a n s of the resu l ts w e r e d e t e r m i n e d us ing o n e - w a y

ana lys is of va r i ance ( A N O V A ) and the data se ts w e r e a n a l y s e d w i th S ta tg raph ics

v5.1 so f twa re (Sta tsof t , U S A ) . D i f fe rences b e t w e e n m e a n s w h i c h w e r e P < 0 .05

w e r e cons ide red to be s ign i f icant . The resu l ts o f the va l idat ion s tudy s h o w e d a

c lear d o s e d e p e n d e n t r e s p o n s e for all three spec ies ( P < 0 .001 ) (F ig . 2 .17) .

65

Cel ls co l lec ted f rom 5 ind iv idua ls and poo led

Cel l suspens ions i ncuba ted w i th 5 0 pL of H2O2 (at 0. 10, 50 . 100 p M ) in s l l i con ised

cen t r i fuge tubes for 10 m i n , o n ice ( in the da rk )

A f te r 10 m in exposu re , t ubes s p u n (at 200 x g for 2 m in ) and supe rna tan t r e m o v e d .

Cel l pel lets w a s h e d tw ice in P B S

Cel l pel lets then r e - s u s p e n d e d in L M P a g a r o s e . Modi f ied C o m e t a s s a y t h e n carr ied

out as per Sec t ion 2.8.1

F ig . 2 .16 F low char t i l lust rat ing the expe r imen ta l d e s i g n for the in vitro va l ida t ion of

the mod i f ied C o m e t a s s a y in Asterias rubens, Paracentrotus lividus a n d Mytilus

edulis.

6 6

(a) 60

50

< 40 z Q TO I-

20

10

0

Buffer E n d o III Buffer F P G

(b)

Control H2O2 concentration (pM)

(C) 70

100

ContrnI

Buffer E n d o III Buffer F P G

ContrnI

H2O2 concentration

Buffer E n d o Buffer F P G

100 H2O2 concentration (jjM)

Fig. 2.17 Percen tage tall D N A fol lowing in vitro exposure to H2O2 for (a) Asterias rubens. (b) Paracentrotus lividus and (c)

Mytilus edulis. Concent ra t ions are expressed in p M . The values are means (/7 = 5) ± S E M . * indicates a signi f icant d i f ference f rom

contro ls at, or above , the 95 % conf idence level.

2.9 Clearance rate

Clea rance ra te (a lso k n o w n as the feed ing ra te ) w a s d e t e r m i n e d pr ior to

h a e m o l y m p h / c o e l o m o c y t e co l lec t ion . T h e assay p rov i ded an ind ica t ion of the

overa l l hea l th of a n o r g a n i s m ( W i d d o w s , 1978 ; W i d d o w s and Sa lke ld , 1993) .

2.9.7 Mytilus edulis

The c lea rance rates ( C R ) w e r e d e t e r m i n e d us ing a stat ic s y s t e m based o n

p rev ious work ( W i d d o w s a n d Sa lke ld . 1993) . Br ief ly, m u s s e l s w e r e p laced into

ind iv idual 4 0 0 m L ta l l - form beake rs , f i l led w i th 350 mL s e a w a l e r (at 15 °C) . E a c h

beaker con ta ined a 12 x 6 m m t r iangular magne t i c fo l lower (F isher®) a n d w a s

p laced on a 15 -po in l m a g n e t i c st i rr ing p la te (IKA® R O 15) . A beake r ac t ing a s a

sys tem b lank w a s a lso p laced on each st i r r ing p late; t hese we re h a n d l e d the s a m e

as the expe r imen ta l beakers , wi th the excep t i on that t hey d id not h o u s e a m u s s e l

(F ig . 2 .18) . T h e musse l s w e r e left to acc l ima t i se ( in o rde r for t h e m to o p e n their

va lves) for 15 m in pr ior to the add i t ion of 500 pL o f Isochrysis galbana a lga l

s u s p e n s i o n (supp l ied by Var i con A q u a So lu t ions L td . , Ma lve rn , UK) , to g i ve c.

15 ,000 a lga l ce l ls mL"^ in e a c h vesse l . A 20 mL a l iquot of water w a s immed ia te l y

r e m o v e d f rom the vesse ls us ing a g lass sy r inge a n d aga in af ter 30 m in had

e l apsed . T h e s a m p l e s w e r e ana l ysed o n a Beckman^"^ Coul ter Par t ic le Size a n d

C o u n t Ana l yse r (Z2) , w i th a 100 p m aper tu re tube f i t ted and set to coun t par t ic les

b e t w e e n 4.0 a n d 10.0 p m in d iame te r (F ig . 2 .19) .

68

T h e C R (L h"^) o f the m u s s e l s w a s c a l c u l a t e d u s i n g t he f o l l ow ing e q u a t i o n

( C o u g h l a n . 1 9 6 9 ) :

C R = V ( l o g e C i - l o g e C2) /t

W h e r e C R = c l e a r a n c e ra te (L h^ ) , V = v o l u m e of w a t e r (L ) .

C i a n d C2 = a lga l concen t ra t i on at b e g i n n i n g a n d e n d o f t ime in te rva l t (h) .

Blank Blank

Fig. 2 .18 B e a k e r s u s e d to d e t e r m i n e t he c l e a r a n c e ra te o f Mytilus edulis a r r a n g e d

a c r o s s two 15-po in t m a g n e t i c st i r r ing p la tes ( IKA ® R O 15) . M u s s e l s a re v is ib le on

the b a s e o f e a c h o f t he e x p e r i m e n t a l v e s s e l s a n d t he e m p t y 'b lank ' b e a k e r s a re

s e e n at the f ront o f e a c h s t i r r ing p la te .

69

1 .

Fig 2 .19 Beckman^*^ Cou l te r Par t ic le S ize a n d C o u n t A n a l y s e r (Z2 ) u s e d to c o u n t

a lga l par t ic les in s a m p l e s co l l ec ted du r i ng the c l e a r a n c e ra te a s s a y .

2 9.2 Ophiothnx fragilis

Bri t t le s tars w e r e p l aced into ind iv idua l 2 L b e a k e r s , f i l led w i t h 1.5 L s e a w a t e r (at

15 °C ) . T h e b e a k e r s w e r e a e r a t e d by m e a n s o f a n air p u m p fed to i nd i v idua l

P a s t e u r p ipe t tes w h i c h w e r e h o u s e d in a uni t p l a c e d o v e r e a c h b e a k e r (F ig . 2 .20 ) .

T h e d e g r e e of ae ra t i on w a s iden t i ca l in e a c h b e a k e r a n d a d j u s t e d so that it w o u l d

70

k e e p Isochrysis galbana a l g a e in s u s p e n s i o n , bu t no t t oo v i g o r o u s tha t it w o u l d

in te r fe re w i th the s u s p e n s i o n f e e d i n g b e h a v i o u r o f t he br i t t le s t a r s ( f low ra te c.

5 L h"^). C l e a r a n c e ra tes w e r e a s s e s s e d w i th a c o n c u r r e n t s y s t e m b lank b e a k e r (a

b e a k e r not hous ing a br i t t le s tar ) . T h e bri t t le s ta rs w e r e lef t t o acc l ima t i se for

3 0 m i n pr ior to t he add i t i on of 5 0 0 p L of Isochrysis galbana a lga l s u s p e n s i o n

( supp l i ed by V a r i c o n A q u a S o l u t i o n s L td . , M a l v e r n , U K ) , to g i ve c i r ca 3 0 , 0 0 0 a lga l

ce l ls m L ^ in e a c h v e s s e l . A l i q u o t s w e r e r e m o v e d f o l l ow ing 2 h. t h e n the C R w e r e

t hen d e t e r m i n e d as per t he m u s s e l s (Sec t i on 2 .9 .1 ) .

Bank

Fig 2 2 0 Expe r imen ta l s e t - u p for d e t e r m i n i n g the c l e a r a n c e ra tes of br i t t le s ta rs .

T h e a lga l s u s p e n s i o n in e a c h v e s s e l is kep t in s u s p e n s i o n v ia a P a s t e u r p ipe t te

d r i ven by a n air p u m p . T h e lef t m o s t b e a k e r is a s y s t e m b l a n k a n d br i t t le s ta rs a re

v is ib le o n the base of the o the r b e a k e r s .

71

2.10 Righting behaviour

Righ t i ng behav iou r in e c h i n o d e r m s p e c i e s g i ves a n ind ica t ion o f t h e ove ra l l hea l th

o f the o r g a n i s m . R igh t ing t i m e s w e r e d e t e r m i n e d pr ior to c o e l o m o c y t e co l l ec t i on .

T h i s a s s a y w a s ca r r ied ou t w i t h s e a s ta rs , s e a u rch ins a n d br i t t le s ta rs , w i th the

s a m e e x p e r i m e n t a l p ro toco l f o l l o w e d for e a c h s p e c i e s . E a c h a n i m a l w a s p l a c e d in

i nd i v idua l a q u a r i u m con ta i n i ng c l e a n s e a w a t e r (at 15 °C) (F ig . 2 .21 ) . T h e

d i m e n s i o n s of the a q u a r i u m a n d v o l u m e of the w a t e r w e r e s u c h tha t the a n i m a l

c o u l d no t t ouch the s ides (or t he s u r f a c e o f the w a t e r ) , a s th is w o u l d h a v e ass i s ted

t h e an ima l ' s r ight ing behav iou r , t h e r e b y lead ing to e r r o n e o u s resu l t s .

F ig . 2 .21 A q u a h u m used to d e t e r m i n e r ight ing b e h a v i o u r (w i th a par t ia l l y r i gh ted

Asterias rubens).

72

T h e an ima l w a s inver ted a n d p laced centra l ly on the base of the tank (n i t r i le

g l oves w e r e w o r n , to p reven t any c h e m o s e n s o r y con tam ina t i on (Za f i r i ou , 1972) ) .

T h e t ime taken for the an ima l to ful ly re tu rn onto its abo ra l s ide is m e a s u r e d , w i th

the a m b u l a c r a of al l the a r m s in con tac t w i th the b a s e of the tank. A n e x a m p l e of a

typical r ight ing r e s p o n s e s e q u e n c e in A. rubens c a n be s e e n in F igu re 2 .22 . Th is

p rocedu re w a s repea ted a total of th ree t imes and the m e a n t ime ca l cu la ted . F ive

m inu tes w e r e permi t ted for e a c h a t tempt , if the an ima l had not comp le te l y

se l f - r ighted af ter this t ime d id not r ight ' w a s reco rded .

73

Fig. 2.22 Ser ies of images i l lustrat ing the r ight ing response in Asterias rubens ( the sequence of images runs f rom (a) to (1)), Sca le bar = 10 m m .

2.11 Determination of cadmium levels in tissues of organisms using

analytical techniques

2.11.1 Mytilus edulis dissection

Following experimental exposure to cadmium (Chapter 5), body compartments

were dissected from mussels to determine the amount of cadmium (Cd) uptake.

To this end, the gills and gut (digestive gland) were removed.

7 (c)

•J

Fig. 2.23 (a) An opened Mytilus edulis showing the location of the gills (red

arrows) and the gut (black arrow), (b) Mussel with gills removed, (c) Mussel with

both the gills and gut (digestive gland) removed, (d) The body compartments on

slides prior to drying to a constant weight.

75

All dissection equipment and microscope slides were acid-washed (5 % HNO3

for at least 2 h) prior to use. For every animal dissected two microscope slides

were labelled and pre-weighed. The valves of each mussel were prised open

using scissors then the adductor muscles cut with a scalpel and the mussel

opened up fully (Fig. 2.23a). Both gills and the gut were removed from the

mussel, washed with Milii-Q water and placed onto the appropriate microscope

slide (Fig. 2.23b. c and d). These slides were placed in a drying oven (100 ±

4 °C) until a constant weight was reached.

2.11.2 Asterias rubens dissection

The same procedure as for M. edulis (Section 2.11.1) was followed with the

exception that different body compartments were analysed along with the gut

(stomach). As a surrogate for the gill tissue of the mussel, the tube feet were

dissected, as respiration occurs through a variety of structures in echinoderms

including the tube feet (Ratcliffe and Rowley, 1979). Additionally, the pyloric

caeca was analysed as biomonitoring studies have shown that this body

compartment is a good bioindicator of heavy metal exposure, especially Cd

(Temara et al., 1998; Danis et al., 2004b). The sea star was placed aboral side

upwards and scissors were used to remove the upper dermis of the central disc,

to reveal the cardiac stomach (gut) (Fig. 2.24b). Once the gut was removed, a

scalpel was used to amputate the arms of the sea star and the ambulacral ridge

which runs along the centre of each arm was separated from the outer dermis

of each arm. The tube feet were dissected from each ambulacral ridge and the

pyloric caecum was separated from the outer dermis of the arms (Fig. 2.24c, d

and e).

76

(a)

/

Fig. 2.24 (a) Asterias rubens specimen prior to dissection, (b) Outer dermis

removed from central disc revealing the cardiac stomach (black arrow), (c)

Specimen with cardiac stomach removed (black arrow), (d) Tubed feet along an

ambulacral ridge, (e) Pyloric caeca visible on two dissected arms (black

arrows).

2.11 3 Tissue digestion

Following collection of the appropriate tissues the process of t issue digestion

was carried out. Briefly, in a fume cupboard, each piece of dried tissue was

placed in 20 mL polythene screw-top digestion vials (scintillation vials, Simport.

Canada) and 1 mL of concentrated HNOawas added. Samples were digested at

70 °C for 2 h in a water bath. Once digestion was completed (no brown fumes

evolved), the tubes were allowed to cool, then diluted with 4 mL of Milli-Q water.

Tissue digest aliquots were stored at 20 °C (± 3 °C).

77

2.11.4 Inductively Coupled Plasma Mass Spectrometry (ICP-MS) analysis

Metal analysis was carried out following a recognised method (Hill. 1999;

Federici et al., 2007). Cadnnium concentrations were measured by using

Inductivity Coupled Plasma Mass Spectrometry (ICP-MS). The software used

with ICP-MS was PQ Vision 4.1.2. (Plasma Quod PQ2 Turbo. Thermo

Elemental. Winsford. UK). To validate the metal analysis, 0.25 mL of two

internal standards were added to all samples. These standards were indium (In)

and thallium (Ti). Cd standards were also prepared in order to calibrate the

instrument before metal analysis. The standard solutions were made by using

1000 mg L" Cd stock solution, diluted in 2 % nitric acid. The Cd standards used

were 0. 10. 20. 40 ppb. In all of the Cd standards 0.25 mL of the two internal

standards were added.

78

Chapter 3:

Echinoderms {Asterias rubens) are more sensitive to known

genotoxins than marine molluscs {Mytilus edulis).

Genotoxic responses can be linked to cytotoxic and behavioural

(physiological) consequences.

The results of this chapter were published in Aquatic Toxicology 94. 68-76

(Canty el al.. 2009) and presented as a platform presentation at SETAC

Europe, Hague, 2007.

79

3.1 Introduction

As global population levels continue to grow there are ever increasing

anthropogenic inputs into the aquatic environment. Of these inputs genotoxic

pollutants are of increasing concern (Wurgler and Kramers, 1992; Claxton et al.,

1998; Jha, 2004). For instance, in 1994 the USA released approximately

0.3 billion kg of substances which were known rodent carcinogens into the

environment, 2 . 9 % of which were discharged directly into surface waters

(USEPA, 1996). Little is known about the effects of these genotoxic agents on

aquatic flora and fauna. Therefore, in recent years environmentally realistic

assays have been developed and studies have been carried out to investigate

the effects of potential genotoxic compounds on aquatic organisms using a

variety of biological responses, or biomarkers (Dixon e l al., 2002; Jha, 2004).

Nevertheless, there is a lack of published literature which has explored links

between genotoxic responses with other responses at higher levels of biological

organisation (Anderson and Wild, 1994; Jha, 2008).

Many different species have been used as test organisms in the environmental

monitoring and toxicity testing of xenobiolics. The aquatic species historically

used are phylogenelically diverse; they include plants (e.g. duckweed (Wang,

1990)), annelids (e.g. Platynereis dumerilii (Jha et al. , 1997)) and crustaceans

(e.g. Daphnia magna (McWilliam and Baird, 2002a) and Carcinus maenas

(Dissanayake and Galloway, 2004)). The blue mussel, Mytilus edulis, has been

frequently used as a sentinel species in many monitoring programmes and

research protocols as a representative of the molluscs (Widdows el al., 1995;

O'Connor. 1996; Dixon e l al., 2002). There is increasing concern that basing

environmental health assessments on one species alone may not be sufficiently

80

robust (Cairns Jr, 1983) and there is a growing need to adopt 'new' test

organisms in the field of aquatic ecotoxicology (BEL, 2002). Many current

biomarker studies only utilise a limited number of lest species that are often

found at contaminated sites and the responses of these organisms may not be

indicative of other species in the community (Jha, 2004; Cheung et al., 2006;

Jha, 2008).

New regulations in the European Union under the Registration, Evaluation,

Authorisation and Restriction of Chemicals (REACH; EC 1907/2006)

programme have recognised these weaknesses. The regulations have

recommended that in order to protect marine organisms from contaminants that

have a high probability of causing toxicity, testing with freshwater fish,

crustaceans and algae should be supplemented with molluscs and

echinoderms. The European Centre for Ecoloxicology and Toxicology of

Chemicals (ECETOC) recommendations to reduce and replace both fish and

amphibian use in chemical toxicity testing included the proposed use of

echinoderms (ECETOC. 2007). The ECETOC task force slated that there is

worldwide interest in developing and adopting intelligent testing strategies (ITS)

for the environmental risk assessment (ERA) of chemicals that may enter the

aquatic realm. There is a need to adopt new invertebrate models into the ITS

approach to ERAs, the proposed use of echinoderms was suggested for two

main reasons: the complete genome of the sea urchin has been mapped (the

purple sea urchin Strongylocentrotus purpuratus (Sodergren el al.. 2006)) and

this genomic data could prove invaluable in the development of very specific

toxicity assays; the close evolutionary links of the echinoderms, at the molecular

level, to fish and amphibians (Adoutte et a!.. 2000) may enable the use of the

81

former in the place of vertebrates in toxicity testing (ECETOC, 2007). An

additional argument in favour of the use of echinoderm species in the

preparation of ERAs for human pharmaceuticals is that it is now deemed

necessary to determine the possible effects of human pharmaceuticals on non-

target, and environmentally relevant, organisms (i.e. echinoderms) in ERAs

(Winter et al., 2009). The potential replacement of vertebrate species (fish) with

echinoderms would have further benefits in that it would be in line with the

philosophy of the 3Rs (reduction, refinement and replacement) in vertebrate

toxicology (Russell and Burch. 1959) as it would comply with the ethical

principal of reducing fish use in acute ecotoxicity testing (Jeram et al., 2005).

However, relatively little is known about the sensitivity of echinoderms

(especially their adult stages) to environmental contaminants (Hutchinson,

2007). Therefore, the relative sensitivity of echinoderms compared with more

traditional invertebrate test species is worthy of further ecotoxicological

investigation.

In this study, the common sea star, Asterias rubens, was selected as a test

organism as representative of the echinoderms. The common sea star is a

euryhaline species that is abundant in the intertidal and sublittoral zones. It has

a widespread distribution around the British Isles and can be found throughout

the Atlantic and Arctic coastal waters of Europe (Picton, 1993; Sarantchova,

2001). Although this echinoderm species has been utilised in biomoniloring

studies (Temara el al.. 1997b; Danis et al., 2004b) it has not been used

extensively in ecotoxicology.

82

The effects of contaminants on behaviour (e.g. predator avoidance,

reproductive and social behaviour) in different fish species have been

extensively studied (for review see Scott and Sloman, 2004). The

implementation of behavioural endpoints has been less common with other lest

organisms frequently used in ecotoxicological studies. Examples of behavioural

and physiological responses studied include: the disruption of natural swimming

behaviour in the mysid shrimp Americamysis bahia following cadmium (Cd)

exposure (Roast et al.. 2000; Roast et al.. 2002); the reduced feeding rates and

locomotory behaviour of the gammarid Gammarus pulex following Cd exposure

(Felten et al., 2008); the impairment of natural burrowing activity in the marine

amphipod Corophium volutator following chronic exposure to crude oil

contaminated sediments (Scarlett et al.. 2007); the adverse impact of exposure

to oil-derived products on the prey-localisation behaviour of the Australian

eleven-armed asteroid Coscinasterias muricata (Georgiades et al.. 2003); the

suppression of feeding rates in Daphnia magna caused by a range of pollutants

(including heavy metals and PAHs) (McWilliam and Baird. 2002b); oil-induced

disruption of the foraging behaviour of the asteroid C. muricata (Temara el al..

1999) and the effects of crude oils on the righting behaviour of the purple sea

urchin Paracentrotus lividus (Axiak and Saliba, 1981). As behaviour, a

phenolypic modification, is controlled by the genetic make-up of individual and

because it links physiological function with ecological processes, it serves as a

useful indicator of toxicity a l the whole organism level. Organism level

responses can provide sensitive and general stress indicators that can be

correlated with community-level (population) responses (Maltby, 1999).

83

However, no attempt has been made previously to link or integrate the

genotoxic effects of contaminants with effects at physiological or behavioural

levels. Such studies would not only provide the opportunity to elucidate the

potential mechanisms of physiological (or behavioural) modifications but could

also stimulate the use of such alterations without the need to use invasive

techniques (Scott and Sloman, 2004). In this context, as previously mentioned,

'clearance rate' (CR) has been used recently (sometimes as a component of

'scope for growth' - an indicator of the energy status of an organism which

takes into account the processes of feeding, food absorption, respiration and

excretion (Bayne, 1984)) in bivalve molluscs as a sensitive indicator of pollutant

exposure (Eertman et al., 1995; Widdows et al., 1995; Canty el al.. 2007;

Scarlett et al., 2008) and 'righting time' (RT) (defined as the time taken for an

animal placed upside down to return to its normal, upright, position) has been

implemented as an indicator of overall health of echinoderm species (Axiak and

Saliba, 1981; Davies el al., 1998; Temara et al., 1999).

Against the backdrop of above information, the aims and objectives of the study

were:

(a) To validate the biomarker responses to determine the effects of genotoxic

agents in the blue mussel {M. edulis) and the common sea star {A. rubens).

(b) To compare the relative sensitivity of the two organisms {M. edulis and

A. rubens) for observed biological responses at different levels of biological

organisation.

84

(c) To test for correlations between genotoxic responses and higher levels of

biological response (including behavioural and physiological responses at the

whole organism level) with an aim to validate their usefulness as a potential

non-invasive tool for toxicological testing and pollution monitoring.

In order to achieve these objectives, a direct acting reference genotoxic agent,

methyl methanesulphonate (MMS) and an indirect acting (i.e. requiring

metabolic activation) agent cyclophosphamide (CP) were selected. Methyl

methanesulfonate (MMS) was selected as it has been used previously as a

reference genotoxin in many studies, including human or mammalian cells

(Natarajan et al., 1983; Andreoli et al., 1999) and in marine invertebrates (Jha et

al., 2000a; Jha et al., 2000b; McFadzen et al.. 2000; Rank and Jensen, 2003;

Hagger et al., 2005a). MMS modifies DNA via nucleotide alkylation through

DNA strand cleavage created at random in the genome (Schwartz and Kmiec,

2005). Cyclophosphamide (CP) (N.N-bis(2-chloroethyl) tetrahydro-2H-1,3,2-

oxphosphorin-2-amine, 2-oxide monohydrate), a chemotherapeutic and

immunosuppressive agent, acts as a neurotoxicant. (Hammer et al., 1994;

Rzeski et al.. 2004; Xiao et al., 2007; Perini et al.. 2008). This compound is

transformed by intracellular enzymes to active alkylating metabolites which

cross link with DNA strands (Anderson et al., 1995).

The production and use of pharmaceutically active compounds increase year by

year and the occurrence of pharmaceuticals (and their metabolites) in the

aquatic environment is of increasing concern (Bound and Voulvoulis, 2004).

Due to the incomplete removal of these pharmaceutical compounds by sewage

treatment works they are found in sewage waste waters, surface waters and

85

even in potable (drinking) water (Sacher et al. , 2001; Stackelberg et al., 2004;

Jones et al.. 2005). Cyclophosphamide is no exception to these trends. Once

administered to patients, CP and its metabolites are renally excreted in

substantial proportions, with up to 10 - 20 % being excreted unchanged

(Anderson et al., 1995; Buerge et al., 2006; Johnson et al., 2008). The active

compounds and metabolites which have poor degradability have been detected

in waste water (sewage) at concentrations up to 100 ng L'\- as a result there is

a growing environmental concern about CP both with respect to human and

ecosystem health (Steger-Hartmann et al.. 1997; Buerge et al., 2006; Johnson

et al., 2008). While CP has been used as a reference genotoxic agent for the

development of genotoxicity assays using embryo-larvae of the marine annelid,

Platynereis dumerilii (Jha et al., 1996), little further data exist for the impact of

this pharmaceutical on the marine organisms.

3.2 Materials and methods

3.2.1 Chemicals

All chemicals and reagents were of analytical grade and obtained from Sigma

Chemical Co., Poole, Dorset. UK. unless otherwise stated. Test chemicals were

methyl methanesulphonate (MMS; CAS number 66-27-3; purity 99 % ) and

cyclophosphamide monohydrate (CP; CAS number 6055-19-2; purity > 97 % ) .

As per the published information (e.g. Johnson et al., 2008;

http://dohs.ors.od.nih.gov), both MMS and CP are readily soluble in water at

room temperature (MMS 250 g L"^ CP up to 50 g L'^).

86

3.2.2 Species collection

Mytilus edulis (35 - 45 mm in length) were collected during March 2006 from

Port Quin (Latitude 50.35° N; Longitude 4.52° W). they were held in conditions

stated in Section 2.1.4. Asterias rubens were collected during March and April

2006 from in and around Plymouth Sound, Devon. UK (Latitude 50.20° N;

Longitude 4.10°W), using SCUBA divers and trawling. They were held in

conditions stated in Section 2.1.1.

3.2.3 Exposure to genotoxins and determination of MTC

The experimental design for the 7 d static exposures has been diagrammatically

represented in Figure 3 .1 . The exposure scenario for the two species differed

slightly in the sense that whilst RT for A. rubens was determined following 3, 5

and 7 d exposure period, cytotoxicity and genotoxicity in this species were

determined only after 5 and 7 d. In contrast, for M. edulis CR, cytotoxicity and

genotoxicily were determined following 3. 5 and 7 d exposure. The MMS

concentrations were selected based on the previous work on P. dumerilii {JUa et

al., 1996) and to mirror ethyl methanesulphonate (EMS: a closely related

alkylating agent), concentrations used in previous work using mussels (Jha et

al., 2005) and thereby the concentration range used ( i.e. 18, 32 and 56 mg L"")

incorporated the semi-logarithmic scale routinely used in ecotoxicity testing.

This range was replicated for the CP concentrations. The exposures were

carried out at 15 ± 0.2 °C and food was withheld. The exposure vessels were

individually aerated using Pasteur pipettes and were sealed with Parafilm^'^.

The vessels were checked daily for mortalities and the parameters of

temperature, salinity, dissolved oxygen and pH measured. In common with

regulatory mammalian genotoxicity criteria, maximum tolerated concentration

87

(MTC) for both the species was taken as being the concentration at which

mortalities occurred (Mackay, 1995).

88

(a)

00 CO

1 lb) f \ A. nibens individually exposed in 4 L

vessels (4 replicates per concentration)

V J

9 M. edulis in 1.8 L vessels (2 replicates)

\ y

t 1 After 3 days righting time determined

(4 A. rubens per concentration)

V J

/ N After 3, 5 & 7 days

(3 M.edulis sampled per replicate, therefore n = 6)

V J t 1 After 5 & 7 days righting time determined


V J

f \ Clearance rates determined

(6 M.edulis sampled per concentration)

\ J coelomocytes collected haemolymph collected

Neutral red, micronucleus & Comet assays carried out


V J

f \ Neutral red, micronucleus & Comet

assays carried out (6 M.edulis sampled per concentration)

V J

Fig. 3.1 Flow chart illustrating the integrated experimental design adopted In the studies to evaluate the behavioural (physiological),

cytotoxic and genotoxic effects of methyl methanesulphonate and cyclophosphamide in (a) Asterias rubens and (b) Mytilus edulis.

3.2.4 Determination of righting times (RT)

In order to assess the 'overall health' of the sea stars at the whole organism

level, the righting time (RT) of each animal was determined. The RTs of the sea

stars were determined prior to coelomocyte sample collection after 5 d and 7 d

(no cellular or sub-cellular assays were conducted following 3 d of the MMS or

CP exposures for the sea stars). Four sea stars per concentration were used.

The procedure for the RT assay was detailed in Section 2.10.

3.2.5 Determination clearance rate (CR)

As a surrogate for the RT assay in the sea stars, the clearance rates of the

mussels were determined after 3, 5 and 7 d (6 mussels per concentration). This

was carried out prior to haemolymph sample collection, as per the method

described in Section 2.9.1.

3.2.6 Collection of coelomocyte and haemocyte samples

Coelomic fluid and haemolymph were collected from the sea stars and mussels,

respectively, as detailed in Sections 2.2.1 and 2.2.2.1. Coelomic fluid was not

collected after 3 d of both sea star exposures as each individual was sampled

repealed during these studies. The coelomic fluid/ haemolymph samples were

held on ice until being used for cytotoxicity and genotoxicity assays. Samples

were thoroughly mixed by vortexing prior to use.

3.2.7 Determination of cellular membrane stability using NRR assay

The cellular membrane integrity of the coelomocytes and haemocytes were

assessed by the neutral red retention assay (NRR). The cell samples were in

triplicate for all individuals. This assay was carried out as per Section 2.3.

90

3.2.8 Determination of cytogenetic damage using Mn assay

T h e n u m b e r s of i nduced m ic ronuc le i ( M n ) in e a c h les t o r g a n i s m w e r e

d e t e r m i n e d w i th the Mn assay . T h e a s s a y w a s car r ied out as de ta i led in

Sec t i on 2.6. O n e s l ide w a s p r e p a r e d for e a c h o r g a n i s m a n d the s l ides w e r e

sco red a c c o r d i n g to the cr i ter ia d e s c r i b e d p rev ious l y (Ven ie r et a l . , 1997 ;

F e n e c h . 2 0 0 0 ; Jha et a l . , 2 0 0 5 ) .

3 .2 .9 Determination of DNA strand breaks using the Comet assay

Sing le s t rand D N A b reaks w e r e quan t i f i ed us ing the a lka l ine C o m e t a s s a y

( p r o c e d u r e is de ta i led in Sec t ion 2 .8) . T w o m ic roge ls per ind iv idua l w e r e

p r e p a r e d . A s requ i red as a p re requ is i te to ca r ry ing ou t the C o m e t assay , it w a s

n e c e s s a r y to d e t e r m i n e the ce l lu lar v iabi l i ty of the h a e m o c y t e s a n d

c o e l o m o c y t e s . Eos in Y d y e a n d a l ight m i c r o s c o p e w e r e used a s de ta i l ed in

Sec t i on 2.7 - on ly cel l s a m p l e s w i th a v iabi l i ty of > 70 % w e r e u s e d in the

C o m e t assay .

3.2.10 Statistical analysis

Stat is t ica l s ign i f i cance b e t w e e n m e a n s w a s d e t e r m i n e d us ing o n e - w a y ana lys i s

of v a r i a n c e ( A N O V A ) . All va lues a re p rov ided as m e a n s ± S E M . D i f f e rences

b e t w e e n m e a n s w h i c h w e r e P < 0 .05 w e r e c o n s i d e r e d s ign i f icant . Al l da ta se ts

w e r e a n a l y s e d us ing S ta tg raph i cs v5.1 so f twa re (Sta tSof t , U S A ) . A n y

co r re la t ions b e e n var iab les w e r e d e t e r m i n e d us ing the P e a r s o n co r re la t i on

coef f i c ien t .

91

3.3 Resul ts

3.3.7 Water quality parameters

W a t e r qual i ty p a r a m e t e r s de te rm ined be fo re a n d a f ter e x p o s u r e pe r iod w e r e in

the no rma l accep tab le range . T e m p e r a t u r e ranged f r o m 14.8 to 15.2 °C , sal in i ty

f rom 34 to 35 P S U , d isso lved o x y g e n leve ls r e m a i n e d a b o v e 70 % sa tu ra t ion

a n d p H f rom 8.0 to 8.2.

3 .3 .2 Determination of MTC

T h e cumu la t i ve mor ta l i t ies for bo th spec ies af ter e x p o s u r e to M M S a n d C P h a v e

b e e n s u m m a r i s e d in T a b l e 3 . 1 . For the M M S e x p o s u r e , the M T C s w e r e

32 m g L'^ a n d 18 m g L"* for M. edulis a n d A. rubens, respect ive ly . Fo l l ow ing C P

e x p o s u r e , the M T C for M. edulis w a s c o n s i d e r e d to be > 1 8 0 m g L"^ as no

mor ta l i ty w a s o b s e r v e d at this concen t ra t i on in m u s s e l s , c o m p a r e d to 56 m g L"^

for A. rubens.

92

Tab le 3.1

Percen tage survival for Asterias rubens and MytHus edulis fo l lowing 7 d exposure to cyc lophosphamide (Exper iment 1) and methy l

me thanesu lphona te (Exper iment 2) .

CD

Species

Exp. A rubens M. edulis

No. Treatment Survival (%) Treatment Survival (%)

(mg L- ) 3 d 5 d 7 d (mg L- ) 3 d 5 d 7 d

Exp. 1 SW Control 100 100 100 SW Control 100 100 100 C P - 18 100 100 100 CP - 18 100 100 100 C P - 3 2 100 100 100 C P - 3 2 100 100 100

CP - 56* 100 100 50 C P - 5 6 100 100 100 CP - 100 NT NT NT CP - 100 100 100 100

C P - 1 8 0 NT NT NT CP -180* 100 100 100

Exp. 2 SW Control 100 100 100 SW Control 100 100 100

MMS-18* 100 75 0 MMS - 18 100 100 100

M M S - 3 2 0 0 0 MIVIS-32* 100 100 50 M M S - 5 6 0 0 0 MMS - 56 100 50 17

* M T C deno ted by under l ined va lues; NT: not tes ted; SW: seawater

3.3.3 Biomarker responses following MMS exposures

3.3.3.1 A. rubens

Due to h igh mor ta l i t ies wh ich occu r red fo l low ing 5 d e x p o s u r e at 18 m g L'^ a n d

subsequen t l y at 3 d wi th h igher concen t ra t i ons ( i .e. 32 a n d 56 m g L ' \ T a b l e

3.1) , l imi ted b iomarke r s tud ies cou ld be p e r f o r m e d . T h e t ime t a k e n for the

an ima ls to r ight s h o w e d a s ign i f icant i nc rease b e t w e e n the cont ro l a n d

18 m g L'^ for 3 and 5 d ( P < 0 .05) . T h e cont ro l t imes w e r e 124 ± 8 s for bo th

days , c o m p a r e d to 223 ± 7 s and 365 ± 30 s, respec t i ve ly for 18 m g L"V

C o m p a r e d to sea wa te r ( S W ) cont ro ls , the m e a s u r e m e n t of cy to tox ic i ty a s

de te rm ined by the N R R a s s a y s h o w e d n o ef fect ( P = 0 .962) . Wh i l s t the N R R

assay s h o w e d no e f fec ts , bo th the Mn a n d C o m e t a s s a y s d e m o n s t r a t e d a

s igni f icant d i f f e rence b e t w e e n contro l va lues a n d at 18 m g L'^ a f te r 5 d (F i g .

3.2) . Mn fo rma t ion w a s i nduced fo l lowing 5 d e x p o s u r e at 18 m g L" \ 13.7 ± 3.5

c o m p a r e d to 5.8 ± 0.9 for the cont ro ls . T h e pauc i t y of b i omarke r da ta (as a

resul t of the h igh d e g r e e o f mor ta l i t ies) m e a n t that It w a s not poss ib le to

de te rm ine any cor re la t ions for this e x p o s u r e .

3.3.3.2 M. edulis

In c o m m o n w i th A. rubens, wh i le n o cy to tox ic e f fec ts (as d e t e r m i n e d by N R R

assay ) w e r e o b s e r v e d c o m p a r e d to con t ro ls , the c l e a r a n c e rate ( C R ) , Induct ion

of Mn and D N A s t rand b reaks ( i .e . C o m e t a s s a y ) s h o w e d a s ign i f icant

d i f fe rence on al l samp l i ng d a y s ac ross all e x p o s u r e concen t ra t i ons (F i g . 3.3) . A s

i l lustrated in F igure 3.4, the re w e r e re lat ive ly s t rong cor re la t ions b e t w e e n C R

and Mn (R = -0 .690 , P = 0 .013 ) a n d b e t w e e n C R a n d % tail D N A (R = -0 .728 .

P = 0 .007) . Fu r the rmore , the re w a s a lso a s t rong cor re la t ion b e t w e e n M n

94

Induct ion and D N A s t rand b reaks ( i .e. % tall D N A ) , a s de tec ted by the C o m e t

a s s a y ( P < 0 . 0 0 1 . R = 0.988).

95

to

250

200

CD

I 150

• | 100 ? ir

Q.

C

o o •o c

50

0

25

20

15

10

Day 3 Day 5 Day 7

IIMR.M_M[V1IV1 M Control 18 32 56

MMS concentration (mg L"')

KL

T-

M MMM. Control 18 32 56

MMS concentration (mg L' )

1.2

Q 0.8

0.6 E S 0.4

- Q 0) O U ^ 0.2

0.0

35

30

25

< Q 20 CO

15

10

5

0

M MMM. Control 18 32 56

MMS concentration (mg L'')

1

KL mi M MMJy MMM Control 18 32 56

MMS concentration (mg L"^)

Fig. 3.2 Behav ioura l (physio logical ) , cytotoxic and genotox ic ef fects In Asterias rubens exposed to M M S for 3. 5 and 7 d. Concent ra t ions are expressed as mg L'V The va lues are means ± S E M . * indicates a signi f icant d i f ference f rom contro ls at, or above , the 95 % conf idence level. NT, not tested; M, 100 % mortal i ty; NR, an imal fai led to right fo l lowing 3 x 600 s a t tempts .

CD

<u 0.5

50

"53 o 40 o o o i_ 30 Q) Q.

ion 20

o 3

•D 10 _C 10

C

Control 32 56 MMS concentration (mg L )

* * • * * -k

Control MMS concentration (mg L ' )

60

50

< 40 Z Q ^ 30 I -

20

10

0

Control MMS concentration {mg L' )

* * * * * * * * *

Control MMS concentration (mg L

Fig. 3.3 Behav ioura l (physio logical ) , cytotoxic and genotox ic in Mytilus edulis exposed to M M S for 3. 5 and 7 d. Concent ra t ions are exp ressed as mg l ' \ The values are means +/- S E M . * indicates a signif icant d i f ference f rom controls at, or above, the 95 % conf idence level.

1.5

1.2 • R = 0.278

P= 0.385

1.08 1.1 1.12 1.14 1.16 NRR (optical density mg protein- )

R =-0.690 P= 0.013

o

0 10 20 30 40

Mn (per 1000 cells)

o

R =-0.728 P= 0.007

10 20 30 40 50

Comet(%tailDNA)

CD 00

c 0)

m u

z

1.16

1.14

1.12

1.1

1.08

R =-0.415 P= 0.180

10 20 30 40

Mn (per 1000 cells)

R = -0.425 P = 0.168 .

0 10 20 30 40 50

Comet (% tail DNA)

R = 0.988 ^ P< 0.001 =S 30

o 20

0 10 20 30 40 50

Comet(% tail DNA)

Fig. 3.4 Linear regression analysis illustrating correlations for behavioural (physiological), cytotoxic and genotoxic effects in Mytilus

edulis following 7 d exposure to MMS. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

3.3 .^ Biomarker responses following CP exposures

3.3.4.1 A. rubens

T w o mor ta l i t ies occu r red fo l low ing 7 d e x p o s u r e at 56 m g L" (Tab le 1). T h e r e

w a s a s ign i f icant i nc rease in RT af ter 5 a n d 7 d at 32 a n d 56 m g L*V In c o m m o n

w i th M M S e x p o s u r e s . N R R s h o w e d n o s ign i f icant d i f f e rence c o m p a r e d to S W

con t ro ls ( P = 0 .9980 a n d 0 .7723 fo l l ow ing 5 a n d 7 d . respec t i ve ly ) . S ign i f i cant

e f fec ts w e r e h o w e v e r de tec ted fo l low ing 5 a n d 7 d e x p o s u r e s at 18, 32 and 56

m g L"^ for the Mn a n d C o m e t assays . Induc t ion of M n inc reased fo l l ow ing 5 a n d

7 d e x p o s u r e at the top two concen t ra t i ons (F ig . 3 .5) . Cont ro l , or base l ine,

f r equenc ies for M n w e r e 2.8 ± 0.9 per 1000 cel ls o n 5 d . c o m p a r e d to 9.8 + 2.5

and 15 + 2.4 for 32 and 56 m g L ' \ respec t i ve ly . A f te r 7 d , these va lues w e r e ,

3.8 ± 1 .1 , 12.5 ± 1.9 and 15.5 ± 0.5. T h e C o m e t a s s a y s h o w e d a s ign i f icant

i nc rease in % tail D N A for the s a m e t ime po in ts . Af ter 5 d , cont ro l va lues w e r e

7.1 ± 1.1 %, c o m p a r e d to 13.5 ± 1.7 % a n d 17.8 ± 2.3 % for 32 a n d 5 6 m g L'\

O n 7 d the va lues w e r e 6.3 ± 1.6 for t he cont ro ls a n d 17.8 ± 1.1 % a n d 24.1 ±

4 .0 % for 32 a n d 56 m g L" \ The re w a s a m o d e r a t e l y s t rong cor re la t ion b e t w e e n

R T a n d M n induc t ion fo l low ing 7 d e x p o s u r e (R = 0 .734 , P = 0 .038 ) a n d

re la t ive ly s t r ong cor re la t ions b e t w e e n R T a n d % tail D N A (R = 0 .910 . P = 0 .002 )

a n d b e t w e e n M n induc t ion a n d % tail D N A (R = 0 .926 , P = 0 . 0 0 1 ; F ig . 3.6) .

9 9

o o

w 500

Control CP concentration (mg L' )

<J 20

Control 18 32 56 CP concentration (mg L"')

Control 18 32 56 CP concentration {mg L'^)

Control 18 32 56

CP concentration (mg L'')

Fig. 3.5 Behav ioura l (physio logical ) , cytotoxic and genotox ic ef fects in Asterias rubens exposed to C P for 3, 5 and 7 d. Concent ra t ions are expressed as mg L'V The values are means ± S E M . * indicates a signif icant d i f ference f rom controls at, or above , the 95 % conf idence level. NT, not tested.

^ 200

0.8 0.9 1 1.1 1.2 NRR (opticaldensity rngprotein-^)

fc

E

t

1.2

1.1

0.9

0.110 0.795

0.81- , 0 5 10 15 20

Mn (per 1000 cells)

0.663 P= 0.073

cr 200

0 5 10 15 20 Mn (per 1000 celts)

fc

Q. O)

E

•t I 15 OC z

1.2

1.1

1

0.9

0.8[

R = 0.331 P= 0.424

0 5 10 15 20 25 30 Connet(%tail DNA)

R = 0.866 P= 0.005

^ 300

^ 200

0 5 10 15 20 25 30 Cometf% tail DNA)

R = 0.926 P= 0.001

0 5 10 15 20 25 30 Comet (% tail DNA)

Fig. 3.6 Linear regression analysis illustrating correlations for behavioural (physiological), cytotoxic and genotoxic effects in Asterias

rubens following 7d exposure to C P . The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

3.3.4.2 M. edulis

N o mor ta l i t ies occu r red e v e n at the h ighes t concen t ra t i on of 180 m g L"^

(Tab le 1). In cont rast to M M S e x p o s u r e s , the C R a n d N R R a s s a y s s h o w e d no

ef fect ac ross all e x p o s u r e concen t ra t i ons and du ra t i ons (F ig . 3 .7) . T h e r e w a s .

howeve r , a s igni f icant d o s e a n d t i m e - d e p e n d e n t Inc rease for t he induc t ion of

M n a n d D N A s t rand b reaks (F ig . 3.7) . In add i t i on , t he re w a s a s t r ong cor re la t ion

b e t w e e n Mn Induct ion a n d D N A s t rand b reaks , as de tec ted by the C o m e t a s s a y

( P < 0 . 0 0 1 . R = 0 .788) (F ig . 3 .8) .

102

o

2.0

1.5

1.0

O 0.5

0.0 Control 18 32 56 100 180


it * -k ^, 1, * * * * * * * * * * 20

Control 18 32 56 100 180

CP concentration (mg L"')

!Q (1) i5 o « a

m 5

Q) O

1.4

1.2

1.0

0.8

0.6

0.4

0.2

0.0

60

50

40

30

20

10

0

i . ^ m 1 m,A L j . l.d

Control 18 32 56 100 180

CP concentration {mg L"^)

* * * * * * * * * * * * * * *

Control 18 32 56 100 180


Fig. 3.7 Behav ioura l (physio logical ) , cytotoxic and genotoxic ef fects in Mytilus edulis exposed to CP for 3, 5 and 7 d at 15 ± 0.2 °C. Concen t ra t ions are exp ressed as mg L'V The values are means ± S E M . * indicates a signif icant d i f ference f rom contro ls at, or above , the 95 % conf idence level.

R = -0.381 P=0.119

0 9 08 0.9 1 1.1 1.2

NRR (optical density mg protein- )

1.4

1-3|

1.2;

1-1'

1;

0.9:

R = 0.100 P = 0.693

0 3 6 9 12 15 Mn (per 1000 cells)

1.4

1.3

1.2

1.1

1

0.9

R = 0.220 P= 0.380 .

10 20 30 40 Comet (% tail DNA)

o 4:

•ffi 1.2,

E 1.T

I 0.9

DC 0.8

i 0

R =-0.371 P= 0.129

3 6 9 12 15 Mn (per 1000 cells)

R =-0.342 P= 0.165

0 10 20 30 40 Comet(% tail DNA)

R = 0.900 ^ 1 2 ^*=0001 8 .

10 20 30 40

Comet (% tail DNA)

Fig. 3.8 Linear regression analysis illustrating correlations for behavioural (physiological), cytotoxic and genotoxic effects in Mytilus

edulis following 7d exposure to C P . The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

3.4 D i s c u s s i o n

T o da te the use fu lness of behav iou ra l (phys io log ica l ) b i o m a r k e r s in

eco tox ico log ica l s tud ies have not b e e n ful ly d e t e r m i n e d . A n a t tempt w a s m a d e

to inves t iga te the s ign i f i cance of two c o m m o n l y u s e d behav iou ra l a n d

phys io log ica l r e s p o n s e s In e c h i n o d e r m s and b iva lve mo l l uscs ( i .e. r igh t ing

behav iou r a n d c lea rance rate, respec t i ve ly ) . L imi ted s tud ies have d e m o n s t r a t e d

that these r e s p o n s e s a re in f luenced in a concen t ra t i on a n d e x p o s u r e

t i m e - d e p e n d e n t m a n n e r us ing a range of chem ica l c o n t a m i n a n t s . T h e r ight ing

response of the sea urch in Paracentrotus lividus w a s impa i red fo l low ing top ica l

e x p o s u r e to c rude oil (Axiak a n d Sa l iba . 1981) and the ant i -paras i t i c a g e n t

i ve rmec t in d id the s a m e to A. rubens (Dav ies et a l . . 1998) . T h e P A H s

b e n z o ( a ) p y r e n e a n d f l uo ran thene ( E e r t m a n et a l . , 1995) and a l k y l b e n z e n e s

(Scar le t t et a l . , 2 0 0 8 ) r educed the c l ea rance rate of M. edulis. Broad ly , wh i ls t

r ight ing behav iou r w a s a f fec ted fo l low ing lowes t concen t ra t i on of M M S ( in the

surv iv ing o r g a n i s m s ) a n d at the two h ighes t concen t ra t i ons of C P u s e d , the

c l ea rance ra te w a s a f fec ted on ly by M M S e x p o s u r e (at all the c o n c e n t r a t i o n s )

and not by a n y concen t ra t i on of C P u s e d .

It has to be e m p h a s i z e d that a d i rec t c o m p a r i s o n of t hese two behav iou ra l or

phys io log ica l r e s p o n s e s wi th d ive rse m e c h a n i s m s (and p u r p o s e ) in

phy logene t i ca l l y d i f ferent g r o u p s of o r g a n i s m s (one f ree l iv ing, the o ther sess i le )

is not log ica l . B a s e d o n th is s tudy , it a p p e a r s however that in g e n e r a l a d i rect

ac t ing g e n o t o x i n (i.e. M M S ) has a m o r e p r o n o u n c e d e f fec t at the w h o l e

o r g a n i s m level in bo th spec ies . Th is w a s a lso very we l l co r re la ted w i th the

mor ta l i ty ra te of the o r g a n i s m s , espec ia l l y in A. rubens, fo l low ing M M S

105

exposu re . Th is is very ev iden t w h e n the M T C v a l u e s of the two spec ies a r e

c o m p a r e d ( M M S : 1 8 m g L " a n d 32 m g L '^ C P 56 m g L"^ and > 1 8 0 m g L" for

A. rubens a n d M. edulis. respec t i ve ly ) .

T h e te rm M T C is s y n o n y m o u s w i th the te rm m a x i m u m to lera ted d o s e ( M T D )

used in the a rena of m a m m a l i a n tox ico logy. T h e c o n c e p t of M T C is bet ter su i ted

for use aquat ic tox ico logy t han M T D , as c h e m i c a l e x p o s u r e o c c u r s t h r o u g h

i m m e r s i o n in mos t aqua t i c tox ico logy s tud ies ( H u t c h i n s o n et a l . , 2 0 0 9 ) . G i ven

the relat ive impo r tance g i ven to the d e t e r m i n a t i o n of M T D in m a m m a l i a n

tox ico logy, it is surpr is ing to no te that he re has b e e n little work in the field o f

aqua t i c tox ico logy that has ut i l ised the M T C c o n c e p t ( Jha et a l . . 2 0 0 0 b ) . S o m e

s tud ies have adop ted the pr inc ipal (for e x a m p l e : ( J h a et a l . . 2 0 0 0 a ; Jha et a l . ,

2 0 0 0 b ; Hagge r et a l . . 2002 ; W in te r et a l . . 2 0 0 8 ) ) a n d these p resen t s tud ies

a t t emp ted to add ress this i ssue further.

T h e concep t of M T D w a s f i rst a d o p t e d in the ear ly 1970s in m a m m a l i a n

carc inogen ic i t y s tud ies and s ince this t ime there h a s b e e n d i s a g r e e m e n t on t he

exac t def in i t ion of the te rm (Buck ley a n d Dora to . 2 0 0 9 ) . T h e endpo in t for M T C

de te rm ina t i on in these s tud ies w a s a n i m a l mor ta l i t y . O the r au tho r s h a v e

s u g g e s t e d d i f fe rent endpo in t s for the de te rm ina t i on of M T D v a l u e s , such a s

b o d y we igh t a n d h i s topa tho logy (Fo ran , 1997) , but t h e s e e n d p o i n t s , a l t h o u g h

su i tab le for m a m m a l i a n in vivo s tud ies w e r e not su i tab le for this b o d y o f wo rk .

T h e b iomarke r techn iques u s e d w e r e of a non -des t ruc t i ve na tu re , the re fo re

h is topa tho logy cou ld not have been car r ied out , a n d the re lat ive ly sho r t du ra t ion

of t hese s tud ies wou ld have m e a n t that any we igh t loss and h i s topa tho log ica l

c h a n g e s in the test o r g a n i s m s w o u l d h a v e b e e n neg l ig ib le . It has b e e n

106

suggested that when assessing the genotoxicity of substances to aquatic

invertebrates behavioural, physiological, developmental and morphological

endpoints may be used to determine MTC (Jha et al.. 2000a). Although the

work presented here used the endpoint of mortality to determine MTC, the

results of the studies show that it would be more appropriate in future work to

use the behavioural (physiological) endpoints utilised here with echinoderms

and molluscs (RT and CR, respectively) when determining MTC in these two

lest organisms.

Cyclophosphamide, being an indirect acting genotoxin (as mentioned in Section

3.1). requires metabolic activation (which is enzyme dependent) prior to causing

a toxic effect. Figure 3.9 shows the metabolic pathway, and associated

metabolites, of CP in a model vertebrate system. Carboxyphosphamide is

formed by the oxidation of CP. which leads to the formation of nornitrogen (a

potent alkylating agent). The second pathway shown in Figure 3.9 leads to the

production of acrolein and phosphoramide mustard (PAM). The latter, PAM. is

the main cytotoxic metabolite of CP and it is responsible for the antineoplastic

(tumour suppressing) activity of CP (Anderson et al.. 1995). Acrolein (an

aldehyde which is ubiquitous in the environment) is believed to be involved in

the toxic side-effects of the medicinal use of CP (McCarroll et al., 2008). Studies

which have found acrolein to be mutagenic and/ or carcinogenic include: The

v79 assay using Chinese Hamster cells found that acrolein had mutagenic

properties (Smith et al., 1990) and dietary exposure of acrolein to adult rats

revealed that it initiated carcinogenesis in the bladder of rodents (Cohen et al.,

1992). Despite these findings, it is thought this metabolite is not itself a

mutagen, but it contributes to the carcinogenic properties of the parent

107

compound. CP, by nature of acrolein's cytotoxic properties (Cox, 1979). Cox

(1979) found acrolein Induced bladder damage (cystitis) in adult rats 48 h after

intraperitoneal injection of a CP analogue which only metabolised into acrolein

(no PAM was produced in vivo by this CP analogue).

Cl

N-^P >

Cl ^ OH Cl.

M fciwv'nat.c V

N - P . • N - P

O — Cl Cl Cl

Cydo(>r>ospnamicc (C^"t i.Hycio.yc>rtODftcsDl«nnioc 1-Kcioci-clcr^ospramKJc

Cl

N - P .CHO

a

Cl.

Cl

N - P

NH.

o n

Phospttca^iice Musia-a :H*.M|

+

^ , C H O

Cl

r . - / > i id :n , li>.>>.ti<>i!ii:i> N — P

) ^ ,COOH , ' o ^

Cl CnrDoxyshosshn'i-.iee

Cl.

NH

Cl

Fig. 3.9 Metabolic pathway of cyclophosphamide in a vertebrate system

(McCarroll e l a l . , 2008).

108

Cyclophosphamide was found to be more toxic at the individual level (as

detected by behavioural or physiological assays) in the echinoderm lest

organism compared to the mollusc. Given that CP requires metabolic activation,

this v^ould suggest that A. rubens has a better, or more efficient, mixed function

oxygenase system compared to that of M. edulis. Given the echinoderms close

evolutionary proximity to the vertebrates (Adoutte et al., 2000; Sodergren et al.,

2006; ECETOC. 2007) it could be expected that sea stars would have a more

efficient enzyme system with which to metabolise xenobiotics into reactive

electrophilic substances compared with that of the mussel. Therefore, toxicity al

the organism level would be more evident in A. rubens following exposure to a

compound such as CP.

While behavioural (or physiological) and genotoxic responses in both the

organisms showed significant effects, cytotoxicity was not found to be induced.

In this context, cellular membrane stability (as measured by NRR assay) in

haemocytes of M. edulis has been found to be significantly correlated with the

Comet response following exposure to Iribulyltin (TBT) under laboratory

conditions (Hagger el al., 2005b) and complex organics under field studies

(Rank el al., 2007). While comparing the relative sensitivity of mussels

(M. edulis) and fish (the Corkwing wrasse, Sympfiodus melops) following

exposure to styrene, it has also been shown that the sensitivity of haemocyles

is greater than peripheral erythrocytes of fish and destabillsatlon of the

lysosomal membrane is significantly associated with the Comet response

(Mamaca et al., 2005). The apparent lack of response for NRR assay in the

present study could be attributed to several factors. It should be noted that the

NRR assay technique used in the present study differs from the technique used

109

in these previous studies. The NRR technique utilised in this study determined

the total membrane stability of the cells (coelomocytes or haemocytes) through

the uptake of NR dye. This technique has been successfully used by many

authors (Olabarrieta et al., 2001 ; Gomez-Mendikute et al., 2002; Asensio et al.,

2007; Canty et al., 2007; Scarlett et al.. 2008; Hannam el al., 2009b) on a range

of cell types. Other authors (like those discussed here) have simply used NRR

uptake to determine the membrane stability solely of the cellular lysosomes.

Therefore caution has to be advised when comparing and contrasting these

NRR assay results as the techniques differ.

Both the genoloxic agents were able to induce genetic damage in the test

organisms, with A. rubens being more sensitive than M. edulis. In line with the

general observations that lower levels of organisation require less dose or

concentration to produce a biological response (Jha, 2008), the effects of two

genotoxins were more apparent at the DNA level. There is a paucity of

published data regarding the effects of genotoxins on echinoderms, but there is

a suite of previous studies which have investigated the effects of these

genotoxins on bivalve molluscs {e.g. M. edulis). Exposure to MMS led to

induction of cytogenetic damage (i.e. sister chromatid exchanges (SCEs) and

chromosomal aberrations (Cabs)) in embryo-larvae of mussels (Jha et al.,

2000a; Hagger et al., 2005a). The Comet assay has also been used to

determine the genotoxicity of MMS to gill cells and haemocytes in M. edulis

(Rank and Jensen, 2003) following exposures (up to 33 mg L*^) for 4 d. The

study suggested haemocytes were more sensitive than gill cells, with highest

level of damage recorded (25 to 30 % tail DNA), being comparable to the

results in these studies. Rank and Jensen (2003) however did not observe a

110

dose-response; in fact they noted a surprisingly low level of genoloxic damage

at the higher exposure levels and hypothesised that this may be due to high

concentration interfering with the processes which lead to DNA damage. In

annelids. MMS exposure to Platynereis dumerHii embryos also induced a clear

dose-response relationship for cytogenetic damage (i.e. SCEs and Cabs) at a

concentration range of 3.5 to 110 mg L" (Jha et al., 1996).

Despite being an environmentally relevant contaminant, there has been only

limited study pertaining to CP induced toxicity and genotoxicity in aquatic

organisms. CP exposure to P. dumerilii embryos induced a clear dose-response

relationship for the induction of cytogenetic damage (SCEs and Cabs) at a

concentration range of 9 - 28 mg L" (Jha et al., 1996). In addition, limited

studies using fish have suggested that intraperitoneal injections of CP to the

fathead minnow, Pimephales promelas, of up to 400 mg kg'^ failed to

significantly enhance the frequency of Mn in erythrocytes sampled from

peripheral blood and spleen samples, despite this dose being the MTC for the

species (Winter et al., 2007). Rodriguez-Cea (2003) exposed three species of

fish under in vivo conditions to CP at 25 and 50 mg L" for 72 h. The Mn assay

showed the trout to be the most sensitive of the species, followed by the

minnow, but the eel was not sensitive to the compound. Since early life stages

of the aquatic organisms are considered to be more sensitive, this may have

been due to the age of the lest organisms, given that both the trout and

minnows were juveniles, whereas the eels were immature adults (Rodriguez-

Cea et al., 2003). Unlike the current study, however, no dose-dependent effect

was observed in the sensitive species. While evaluating relative sensitivity of

fish and mammals, following an intraperitoneal injection of 20 mg kg"^ (the MTC)

111

of CP in Tilapia rendalli, and of 30 mg kg ' into the Swiss Mouse, there was a

significant increase in Mn induction (Grisolia, 2002). In this context, it has been

shown that compared to mussels (M. edulis), sea urchins (Strongylocentrotus

droebachiensis) are more sensitive to the induction of DNA strand breaks

following exposure to crude oil (Taban et al., 2004).

Except for the preliminary study by Taban et al. (2004) with adult sea urchins

and bivalves, and several studies using embryo-larval of sea urchins and

mussels (Pagano et al., 1996; His et al., 1999; Bellas et al. . 2005) there has

been no other study in which the relative sensitivity of adult echinoderms have

been concurrently compared at different levels of biological organisation, under

similar experimental conditions with any other organism. Taking into account

available information and comparing the results obtained in the present study

showing induction of genetic damage at relatively lower concentrations

(compared to concentrations required for adult life stages of fish and mice), it

could be suggested that echinoderms are relatively more sensitive than bivalve

molluscs. Furthermore, because of their close phylogenetic relationship, they

could serve as a potential substitute for other aquatic vertebrates (e.g. fish) for

toxicological studies.

This study suggested strong correlations between induction of DNA strand

breaks and the cytogenetic damage in both organisms. This is in line with other

studies where strong correlations between these two genotoxic parameters

have been demonstrated in fish, molluscs and mammalian cells both under in

vitro and in vivo conditions (Raisuddin and Jha, 2004; Nagger et al., 2005b; Jha

et al., 2005; Villela et al., 2007). For example, positive correlations were found

112

for the induction of Mn and percentage tail DNA as detected by the Comet

assay in Chinese Hamster ovary (CHO) cells and rainbow trout gonad cells

following exposure to sodium arsenate and arsenite (Raisuddin and Jha, 2004);

in adult M. edulis following TBT exposure (Hagger et al., 2005b) and in the

golden mussel, Limnoperna fortunei following exposure to contaminated field-

collected water samples (Villela et al., 2006).

In some studies targeting the erythrocytes of fish (which have some inherent

limitations as mentioned earlier) however, such correlations have not been

found (Belpaeme el al., 1996; Chang et al., 2005). Interestingly, the genoloxic

responses also showed good correlations with behavioural or physiological

responses in both the organisms. II is however to be remembered that

genotoxic responses were determined in coelomocytes and haemocytes of the

organisms (i.e. at the cellular level), whereas the behavioural responses were

measured al the whole organism level. Given the open vascular system of the

organisms, it could be assumed thai other target cells could be affected (either

directly or indirectly) leading to changes at the whole organism level.

The use of coelomocytes and haemocytes could be considered analogous to

the situation in mammalian or human health arena where use of surrogate cells

(e.g. lymphocytes and bone marrow cells) are widely used to evaluate the toxic

potential of environmental contaminants which could lead to adverse health

consequences at the whole organism or individual level. In this context,

although in this study, as previously mentioned, a cytotoxic effect was not

observed, a correlation between cytotoxic effect and genotoxic effect (as

113

determined by Mn assay) In different cell types (I.e. digestive gland and gill

cells) of mussels has been found (Kalpaxis et a!., 2004).

Given that behavioural effects could be directly controlled by the nervous

system, it is now emerging that neurotoxicity has become a frequent effect of

genotoxic and chemotherapeutic agents. Cyclophosphamide and other

chemotherapeutic agents (cisplatin, methotrexate, vinblastin and Ihlotepa) were

found to be neurotoxic both in vitro and in vivo (RzeskI et al., 2004). In vitro

exposure of these agents to rat neurons resulted in a concentration-dependent

neurotoxic effect. In vivo exposure, using juvenile rats (via intraperitoneal

injection), caused an array of brain lesions less than 24 h after injection of the

lest compounds (Rzeski et al., 2004). The genotoxins methylazoxymethanol

and nitrogen mustard induced reduced cellular viability and increased DNA

damage in nerve cells (granule and astrocyte cells) prepared from the cerebella

of young mice (Kisby et al.. 2006). Maternal injection of CP has been found to

Induce neuronal apoplosis and DNA damage In developing rat embryos (Xiao et

al.. 2007).

Echinoderms, as with mammals, have the ability to synthesize, store and

release neurotransmitters (e.g. acetylcholine, serotonin, dopamine etc.) and

possess homologous populations of receptors and their downstream signalling

cascades (Buznikov et al., 1970; Guslafson and Toneby, 1970; Buznikov et al.,

2007). It Is therefore not surprising that in addition to genotoxic responses, the

agents used in the present study also induced behavioural and physiological

responses, with the echinoderms being more sensitive than the molluscs. Given

that approximately one third of all industrial discharges could be genotoxic

114

(Claxton et al., 1998) and taking into account the vast amount of anthropogenic

discharges reaching the hydrosphere, ecotoxicological consequences of these

contaminants could be immense.

In addition to being neurotoxic, genotoxins can also act as cytotoxic,

immunotoxic and endocrine disrupting agents (Jha, 2008). For instance, TBT

exposure to P. dumerilii embryo-larvae resulted not only in a genotoxic effect

(as measured by SCEs and Cabs), but also a cytotoxic effect (proliferative index

(PRI)) (Hagger et al., 2002). In the human health arena, exposure to

environmental agents including genotoxins have been correlated to

developmental disorders (e.g. learning disabilities, attention deficit disorder) in

which interaction between environmental factors and genetic susceptibility play

an important role (Branum et al., 2003; Landrigan et al., 2004).

To conclude, whilst validating and employing a range of non-destructive

biomarkers at different levels of biological organisation, the relative sensitivities

of two ecologically relevant marine invertebrates have been evaluated. While a

small number of studies are available in literature where a range of biological

responses have been evaluated in a single species (e.g. Regoli et al., 2004),

there is a paucity of information available in the literature in which the sensitivity

of two groups of organisms have been concurrently compared. This study

indicated that when compared to marine bivalves, asteroids are more sensitive

to direct and indirect-acting genotoxicants.

This study also suggested that in addition to clearance rate in mussels, righting

behaviour in echinoderms has the potential to serve as a sensitive indicator of

115

behavioural (or physiological) alterations following exposure to genotoxins.

which could simultaneously act as neurotoxicants. In addition, while a direct link

was found between genotoxic and behavioural responses, these studies also

suggested that expression of different biological responses might not always be

linked or correlated. This further strengthens the notion that depending upon

levels of exposure, properties of contaminants and the target cells, toxicological

properties of any substance could be manifested in a variety of ways (Hagger et

al., 2002; Jha, 2008). Further studies will strengthen the use of adult

echinoderms as sensitive bioindlcator species for environmental studies. These

will Include the use of an environmentally relevant polycyclic aromatic

hydrocarbon (PAH), i.e. benzo(a)pyrene and the implementation of a recovery

period following in vivo exposure to determine If any deleterious effects are

irreversible.

116

Chapter 4:

A nnodel polycycl ic aromat ic hydrocarbon (benzo(a)pyrene) can

cause behavioural (physiological) , cytotoxic and genotox ic

consequences in three different ech inoderm spec ies (Asterias

rubens, Paracentrotus lividus and Ophiothrix fragilis) and a bivalve

mol lusc (Mytilus edulis).

These consequences at different levels of biological organisat ion will

be reversible.

117

4.1 Introduction

Polycyclic aromatic hydrocarbons (PAHs) are a group of more than 100

chemicals that are formed during the incomplete burning of coal, oil, gas, wood,

or other organic substances, such as tobacco (ATSDR, 1995). They are

omnipresent throughout the environment In the air, water, and soil. Levels of

PAHs In the world's oceans are increasing every year due to anthropogenic

activities (e.g. oil spills and the combustion of fossil fuels) (Woo et al., 2006).

Estimated world inputs (in million tonnes per year) of PAHs into the marine

environment are summarised in Table 4 .1 .

PAHs are a group of compounds which are very lipophilic and consist of 2, or

more, fused benzene rings (Readman et al., 2002). Benzo(a)pyrene (B(a)P) Is

an example of a PAH, which is a ubiquitous pollutant and an indirect acting

mutagen (Kim and Hyun. 2006). B(a)P is classed as one of 33 priority pollutants

by the European Union (annex II of the Directive 2008/105/EC) and one of 129

priority pollutants by the United States Environmental Protection Agency

(USEPA). The structure of B(a)P consists of a benzene ring attached to a

pyrene molecule (Fig. 4.1).

118

Table 4.1 Estimated global input of PAHs into the marine environment (from

Clarke, 2001).

Source Input (million tonnes year'^) Total Transportation Tanker operations 0.163 Tanker accidents 0.162 Bilge and fuel oil 0.524 Dry docking 0.009 Scrapping of ships 0.002 Non-tanker accidents 0.020 Atmospheric emissions 3.750

4.630 Fixed installations Coastal refineries 0.10 Offshore production 0.05 Marine terminals 0.03

0.180 Other sources Municipal wastes 0.70 Industrial wastes 0.20 Urban run-off 0.12 River run-off 0.04 Atmospheric fall-out 0.30 Ocean dumping 0.02

1.380 Natural sources 0.250

Total 6.440

Fig. 4.1 A ben2o(a)pyrene molecule (consisting of a benzene ring attached to a

pyrene molecule).

119

PAHs require metabolic activation before they cause mutagenic or carcinogen

effects. In mammalian systems, PAHs are initially activated into arene oxides by

cytochrome P450-depenadent monooxygenases. These arene oxides can

spontaneously isomerise into phenols or be converted into trans-dihydrodiols by

epoxide hydrases. When the trans-dihydrodiols are oxidised into diolepoxides,

the resultant compounds are believed to react with other metabolites to initiate

malign cell transformation (Jacob, 1996). The sequence of B(a)P metabolic

activation in vertebrates is detailed in Figure 4.2.

MonocnrygoniiMt 10)

^utcnridstlon OH

1 o Qluni

ghjtBtrilon or protilru HO H

avya«na Prolan

ydrast w th DNA

Fig. 4.2 The sequence of metabolic activation of B(a)P in vertebrates. (From

Jacob. 1996).

120

It is thought that invertebrate metabolism of PAHs follows a similar pathway as

that which occurs in vertebrate systems (Livingstone, 1991). Crustaceans have

been found to metabolise hydrocarbons faster than molluscs and fish

metabolise B(a)P faster than marine invertebrates (Bihari and Fafandel, 2004).

The Impact of B(a)P on a range of aquatic organisms has been investigated by

many researchers, these include: flounder, Paralichthys olivaceus (Woo et al.,

2006), European green crab, Carcinus aestuarii (Rossi et al., 2000), scallop,

Chlamys farreri {Pan et al., 2008), zebra mussel, Dreissena polymorpha (BInelli

et al., 2008), green-lipped mussel, Perna viridis (Ching et al.. 2001; Siu et al..

2004; Fang e l al., 2008; Slu e l al., 2008) and Mediterranean mussel, Mytilus

galloprovincialis (Venier el al., 1997; Canova et al.. 1998; Gomez-Mendikute el

al., 2002; Marigomez and Baybay-Villacorla, 2003). These studies have

encompassed a wide range of assays which operate at different levels of

biological organisation, from whole organism level (heart rate (Fossi et al. ,

2000)), to cellular level (NRR (Gomez-Mendikule et al., 2002; Fang el al. ,

2008)), through to sub-cellular level (Mn induction (Venier el al., 1997; Slu et al.,

2004; Slu el al., 2008) and the Comet assay (Bihari and Fafandel, 2004; Siu el

al.. 2004; Slu et al., 2008)). Examples of selected in vivo benzo(a)pyrene

biomarker studies with the adult stages of different aquatic organisms are given

in Table 4.2.

121

Table 4.2 Examples of selected in vivo benzo(a)pyrene biomarker studies with the adult stages of different aquatic test species.

Phyla Species Biomarker References Chordata Dicentrarchus labrax

Gambusia affinis Paralichthys olivaceus Serranus scriba

Cytochrome P450 induction DNA single strand breaks Comet assay DNA single strand breaks

(Gravato and Santos, 2002) (Bihah and Fafandel, 2004) (Woo etal . , 2006) (Bihari and Fafandel, 2004)

Crustacea Carcinus aestuarii

Daphnia magna

Maja crispate

Mixed function oxidase activity Alkaline unwinding assay Micronucleus induction Heart rate DNA damage Parental fecundity DNA single strand breaks

(Fossi et al.. 1996) (Fossi et al.. 1996; Fossi et al.. 2000) (Fossi et al., 2000) (Fossi eta l . , 2000) (Atienzar et a!., 2002; Atienzar and Jha, (Atienzarand Jha, 2004) (Bihari and Fafandel, 2004)

2004)

Mollusca Chlamys farreri

Chlamys islandicus Dreissena polymorpha

Mytilus galloprovincialis

Perna viridis

Aryl hydrocarbon hydroxylase activity Alkaline unwinding assay Cellular membrane stability Micronucleus induction Comet assay DNA adduct formation DNA single strand breaks Micronucleus induction Lysosomal structural changes Cellular viability DNA adduct formation DNA strand breaks Micronucleus induction Comet assay Lysosomal stability

(Pan eta l . . 2008) (Pan eta l . . 2008) (Camus et al.. 2002) (Binelli et al., 2008) (Binelli e ta l . . 2008) (Canova eta l . , 1998) (Bihari and Fafandel, 2004) (Venier eta l . . 1997) (Marigomez and Baybay-Villacorta. 2003) (Gomez-Mendikute et al., 2002) (Ching eta l . , 2001) (Ching eta l . , 2001) (Siu et al., 2004; Siu et al.. 2008) (Siu et al., 2004; Siu et al., 2008) (Fang eta l . , 2008)

These previous studies have covered a range of marine invertebrate species,

but there is a paucity of studies which have investigated the effects of B(a)P on

echinoderm species. To address this problem, the present study not only used

the blue mussel Mytilus edulis as a test organism, but also the common sea star

>As/er/as rubens, the purple sea urchin Paracentrotus lividus and the brittle star

Ophiothrix fragilis. The assays utilised evaluated the potential effects at different

levels of biological organisation from the whole animal level (righting time and

clearance rate), to cellular level (the neutral red assay) and sub-cellular levels

(micronucleus induction and the Comet assay). The experimental design

incorporated a recovery period to determine if any deleterious effects observed

at the different levels of biological organisation were reversible.

The aims and objectives of this study were:

(a) To determine if the relative sensitivity to B(a)P at different levels of biological

organisation in three different echinoderm species {Asterias rubens,

Paracentrotus lividus and Ophiothrix fragilis) compared to bivalve molluscs

{Mytilus edulis).

(b) Through the implementation of a recovery period into the experimental

designs, to assess whether any deleterious effects (at any of the levels of

biological organisation) were irreversible?

123

4.2 Materials and methods

4.2.1 Chemicals


Chemical Co.. Poole, Dorset, UK. The test chemical investigated was

benzo(a)pyrene (B(a)P; CAS number 50-32-8; purity > 96 % ) and the solvent

vehicle used was dimethyl sulphoxide (DMSO; CAS number 67-68-5).

Appropriate weights of B{a)P were dissolved in DMSO and added to the

exposure vessels to provide nominal concentrations of 0 .1 , 1.0 and 10 pg L'\

Final DMSO concentrations were 0.005 % (v/v) in all test vessels.


Asterias rubens and O. fragilis were collected during January and February

2007 from in and around Plymouth Sound, Devon, UK (Latitude 50.20° N;

Longitude 4.10°W). using SCUBA divers and trawling. They were held under

the conditions slated in Sections 2.1.1 and 2.1.3. respectively. Paracentrotus

lividus were collected in Bantry (Co. Cork, Eire) (Latitude 51.41° N; Longitude

9.28° W) and supplied by Dunmanus Seafoods Ltd (Co. Cork, Eire) in January

2007. They were held as per Section 2.1.2. Mytilus edulis were collected during

April 2007 from Port Quin (Latitude 50.35° N; Longitude 4.52° W) they were held

in conditions outlined in Section 2.1.4.

4.2.3 Exposure to B(a)P

The experimental design for the series of B(a)P exposures incorporated a

sampling day on the third and seventh exposure days, then the animals were

placed in clean seawater (with 0.005 % DMSO) for a further three days to

assess their recovery. Five replicates at each concentration were used for all

124

species. The exposure vessels (7 L glass aquaria for the sea stars and urchins,

1.5 L glass beakers for the brittle stars and mussels) were spiked daily with the

appropriate nominal concentration of B(a)P solution (during the 7 d exposure

period). The experimental design for the static exposures is diagrammatically

represented in Figure 4.3. The exposure scenario for the four species differed

slightly because, as previously stated (see Section 2.2.2.3) it was not possible

to collect sufficient coelomic fluid from O. fragilis to carry out the cytotoxic and

genotoxic assays. Therefore only behavioural (RT) and physiological (CR)

assays were carried out on the three sampling days for this species. In the case

of the sea stars, sea urchins and mussels the complete suite of behavioural

(physiological), cytotoxic and genotoxic assays were carried out on each

sampling day.

Sublethal nominal concentrations of B(a)P, 0 .1 , 1.0 and 10 pg L'^ were used.

These concentrations were selected to be similar to those used in previous in

vivo studies with mussels (Siu et al., 2004) and crabs (Fossi et al., 2000); and

also to be environmentally realistic. In the open coastal waters of Hong Kong

B(a)P concentrations range from 10 to 40 pg L' (Ching et al.. 2001) and in

freshwater streams of the USA the maximum concentration of B(a)P detected

was 0.24 pg L' (Kolpin et al., 2002). The exposures were carried out at

15 ± 0.3 °C and food was withheld during the exposure period. The exposure

vessels were individually aerated using Pasteur pipettes and were sealed with

Parafi lm™. The vessels were checked daily for mortalities and the parameters

of temperature, salinity, dissolved oxygen and pH were measured.

125

(a) (b) (c)

Sea stars/ urchins individually exposed in 7 L vessels

(5 replicates per concentration)

3 mussels exposed in 1.5 L vessels (5 replicates per concentration)

Brittle stars individually exposed in 1.5 L vessels

f5 replicates per concentration)

After 3 & 5 days exposure & 3 days recovery, 1 animal sampled per replicate

(n = 5)

After 3 & 5 days exposure & 3 days recovery. 1 mussel sampled per

replicate (n = 5)

After 3 & 5 days exposure & 3 days recovery. 1 animal sampled per

replicate (n = 5)

Righting times determined (n = 5)

Clearance rates determined {n = 5)

coelomocytes collected haemolymph

Neutral red. micronucleus & Comet assays carried out

(n = 5)

collected

Neutral red, micronucleus & Comet assays carried out

(n = 5)

Righting times determined (^ = 5)

Clearance rates determined {n = 5)

Fig. 4.3 flow chart illustrating the integrated experimental design adopted in the studies to evaluate the behavioural (physiological),

cytotoxic and genotoxic effects of B(a)P in (a) sea stars/ urchins, (b) mussels and (c) brittle stars.


Righting times (RT) of the echinoderms were determined prior to coelomocyte

sample collection (5 sea stars, urchins or brittle stars per concentration) as

detailed in Section 2.10.

4.2.5 Determination of clearance rate (CR)

In the case of the mussels, CRs were determined (5 mussels per

concentration), prior to haemolymph sample collection, as per the method

described in Section 2.9.1. For the brittle stars (n = 5) clearance rates were

determined as per Section 2.9.2.


Coelomic fluid and haemolymph was collected as per Sections 2.2.1, 2.2.2.1

and 2.2.2.2. The coelomic fluid/ haemolymph samples were held on ice until

used for cytotoxicity and genotoxicity assays. The samples were thoroughly

mixed by vortexing prior to use.


The cellular membrane stability of the mussel haemocytes and the sea star and

sea urchin coelomocytes (all n = 5) were determined using the NRR assay. This

assay was carried out as per Section 2.3 and triplicate samples were analysed

for each individual organism.

127

4.2.8 Determination of cytogenetic damage using Mn assay

This assay was carried out as detailed in Section 2.6 for the mussels, sea stars

and sea urchins. One slide was prepared for each test organism and the slides

were randomised and coded prior to scoring under a light microscope according

to the criteria described previously (Venier et al., 1997; Fenech, 2000; Jha et

al., 2005).

4.2.9 Determination of DNA strand breaks using the alkaline Comet assay

The frequencies of single strand DNA breaks in the mussels, sea stars and sea

urchins (all n = 5) were quantified with the alkaline Comet assay. One slide

(each with 2 microgel replicates) was prepared for all test organisms. This

assay was carried out as per Section 2.8. As required as a prerequisite for the

Comet assay, it was necessary to determine the cellular viability of the

haemocytes and coelomocytes. To this end, Eosin Y dye and a light microscope

were used as detailed in Section 2.7 - only cell samples with a viability of >70 %

were used in the Comet assay.

4.2.10 Statistical analysis

Statistical significance between means was determined using one-way analysis

of variance (ANOVA). All values are provided as means ± SEM. Differences

between means which were P < 0.05 were considered significant. All data sets

were analysed using Statgraphics v5.1 software (StatSoft, USA). Any

correlations been variables were determined using the Pearson correlation

coefficient.

128

4.3 Results


Water quality parameters measured during all the exposures were found to be

within the normal acceptable range. Temperature ranged from 14.9 to 15.3 "^C,

salinity from 33 to 35 PSU, dissolved oxygen levels remained above 70 %

saturation and pH from 8.0 to 8.3.

4.3.2 Biomarker responses following B(a)P exposures

4.3.2.1 Asterias rubens

The righting time for the sea stars showed no significant increase over the 7 d

exposure. The NRR assay showed an increase in cell membrane stability

following 3 d exposure (P < 0.05), which returned to control levels after 7 d

B(a)P exposure (Fig. 4.4). Significant induction of Mn were found after 3 d and

7 d and these remained statistically significant even after 3 d recovery period in

clean sea water. Genotoxic effects were evident with increased percentage tail

DNA after 7 d and this remained increased compared to the control values

following the recovery period. A moderately strong correlation was found

between Mn induction and percentage tail DNA (R = 0.731, P < 0.001; Fig. 4.8).

4.3.2.2 Paracentrotus lividus

No significant physiological effects (as determined by the righting assay) or

cytotoxic effects (NRR assay) were detected (Fig. 4.5). Increased Mn induction

occurred after 3 d and 7 d (both P < 0.001) and remained elevated following 3 d

recovery (P < 0.005). The Comet assay detected an increased amount of tail

DNA after 3 d (P < 0.005) and 7 d (P < 0.001), but after 3 d in clean sea water

there was no significant difference from the control values (P > 0.005). The

129

NRR and Comet assays showed a relatively weak correlation (R = 0.280,

P = 0.031; Fig. 4.9). A moderately strong correlation was found between the two

genotoxic parameters, Mn induction and percentage tail DNA (R = 0.726.

P < 0.001; Fig. 4.9).

4.3.2.3 Mytilus edulis

As with P. lividus, no physiological effects (clearance rate) or cytotoxic effects

(NRR assay) were detected (Fig. 4.6). The Mn assay showed a significant

increase in Mn numbers after 3 and 7 d (both P < 0.001), but after the recovery

period the number of Mn returned to control levels. Increased tail DNA was

induced after 3 and 7 d (both P < 0.001) and remained above the amounts

detected in control mussels following 3 d in clean sea water (P > 0.005). In a

similar fashion to the other test organisms, a moderately strong correlation was

found between Mn induction and percentage tail DNA (R = 0.834, P < 0.001;

Fig. 4.10).

4.3.2.3 Ophiothrix fragilis

Only behavioural (physiological) effects were investigated in the brittle stars.

The time taken for the animals to self right was significantly increased after 7 d

exposure (P > 0.005), but these times returned to control levels after the

recovery period (Fig. 4.7). The clearance rates remained unaffected following

exposure to B(a)P. It is worth noting that the overall mean control clearance rate

of the brittle stars was 0.171 ± 0.014 L h'\ compared to a value 1.31

± 0.062 L h" for the mussels. This difference is accounted for by the difference

anatomy and physiology of the two species. That is, the brittle stars suspension

feed using their tube feet, where as the mussels possess a powerful siphon

130

pump which is much for efficient. No correlation was found between righting

time and clearance rates In O. fragilis (Fig. 4.11).

131

300

250

30 1

cells

)

25

o o o 20 •

(per

15 • c g

indu

ct

10 -

c 5 •

Day 3 Day 7 Recovery Day 3

Control 0.1 1

B(a)P concentration (pg L'')


Control 0.1 1 10

B(a)P concentration (pg L" )

0.7

0.6

0.5

0.4

0.3

30

25

20


^ 0.2

Control 0.1 1 10


I Day 3 ' Day 7

• Recovery Day 3

Control 0.1 1 10

B(a)P concentration (pg L*')

Fig. 4.4 Behavioural (physiological), cytotoxic and genotoxic effects in Asterias rubens exposed to B(a)P for 3 and 7 d and following a 3 d recovery period (In clean seawater). Concentrations are expressed as pg L"V The values are means ± SEM (n = 5). * indicates a significant difference from controls at, or above, the 95 % confidence level.

250

00

o. c o

' i •o c c

25

20


P 100

Control 0.1 1 10



Control

B(a)P concentration (fjg L" )

1.5

c

o

30


Control B(a)P concentration (pg L' )

Day 3 Day 7 Recovery Day 3 * *

Control B(a)P concentration (pg L"')

Fig. 4.5 Behavioural (physiological), cytotoxic and genotoxic effects in Paracentrottis lividus exposed to B(a)P for 3 and 7 d and following a 3 d recovery period (in clean seawater). Concentrations are expressed as pg L" significant difference from controls at, or above, the 95 % confidence level.

The values are means ± SEM (n = 5), * indicates a

2.0

CO

a c o u 3

•D C

c 5

15

12

3

0 ^

• Day 3 J Day 7

• Recovery Day 3

1.8

Control

B(a)P concentration (|jg L" )


* •

* *

T

I (ML Control 0.1 1 10

B(a)P concentration (pg L"')

30

25

• Day 3 J Day 7 • Recovery Day 3

Control 0.1 1 10

B(a)P concentration (pg L"')

I _ Day 3 Day 7 Recovery Day 3

• * * * * *

Control 0.1 1 10 B(a)P concentration (pg L"')

Fig. 4.6 Behavioural (physiological), cytotoxic and genotoxic effects in Mytilus edulis exposed to B(a)P for 3 and 7d and following a 3 d recovery period (in clean seawater). Concentrations are expressed as pg L'\ The values are means ± SEM (/7 = 5). * indicates a significant difference from controls at. or above, the 95 % confidence level.

E

C

125

100

I Days 3 Day 7

I Recovery Day 3

Control 0.1 1 10

B(a)P concentration (pg L' )

B

c CD

O

0,30

0.25

0.20

0.15

0.10

0.05

0.00

mmm Oay 3 I 1 Day 7

Recovery Day 3

Control 0.1 1 10

B(a)P concentration (yg L'^)

Fig. 4.7 Behavioural (physiological) effects in Ophiothrix fragilis exposed to B(a)P for 3 and 7 d and following a 3 d recovery period (in

clean seawater). Concentrations are expressed as pg L'V The values are means + SEM (n = 5). * indicates a significant difference from

controls at, or above, the 95 % confidence level.

400

300

.^200

100

0

R =-0.014 P = 0.916

0 0.25 0.5 0.75 1 NRR (optical density mg proteirr )

400

300

200

R = 0.239 P = 0.066

0 10 20 30 40 Mn (per 1000 cells)

400 'R = 0.186 p = 0.156

300 • . • •

m • vt 200

QC

100 f

0 '• - •

0 5 10 15 20 25 30 Comet r% tail DNA)

R = 0.107 P = 0.416

0 25 - r r ;

Q: cr: z 0 10 20 30 40

Mn (per 1000 cells)

I 1

|o.75 E ^ 0.5

J 0.25

K 0

R = 0.126 P = 0.339

— : —

40

0 5 10 15 20 25 30 Comet (% tail DNA)

R = 0.731 P < 0.001

0 5 10 15 20 25 30

Comet(% tail DNA)

Fig. 4.8 Linear regression analysis illustrating correlations for behavioural (physiological), cytotoxic and genotoxic effects in Asterias rubens following 7 d exposure to B(a)P. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

400

300

^ 200

100

oL

R =-0.078 P = 0.552

0 0.5 1 1.5 2

NRR (opticaldensity mg protein- )

I 2 O

o) 1.5

0.5

R = 0.122 P= 0.352

•

0 5 10 15 20 25 Mn (per 1000 cells)

400

300

200

100

R = 0.217 P = 0.096

0 5 10 15 20 25 Mn (per 1000 cells)

R =0.280 P= 0.031

5> 1.5

I

0 10 20 30 40 Comet (% tail DNA)

400

300

R = 0.146 P= 0.266

^ 200

0 10 20 30 40 Comet (% tail DNA)

R =0.726 P< 0.001

0 10 20 30 40 Comet (% tail DNA)

Fig. 4.9 Linear regression analysis illustrating correlations for behavioural (physiological), cytotoxic and genotoxic effects in Paracentrotus lividus following 7 d exposure to B(a)P. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

2

^ 1.5

5 1 DC " 0.5

R = 0.211 P= 0.873

0 0.5 1 1.5 2 NRR (optical density mg protein- )

2

1.5

O 0.5

oL R =-0.003 P= 0.986

0 5 10 15 Mn (per 1000 cells)

2

^ 1.5

1 Q:

R =-0.051 P= 0.700

0 5 10 15 20 25 30 Comet(% tail DNA)

I 1.5 t c 1

0.5

g 0

-

R = 0.098 P = 0.457

0 5 10 15 Mn (per 1000 cells)

I ^ |» 1.5

.« 0.5 R = 0.010 P= 0.937

o 8

15

10

I 5

R = 0.a34 P < 0.001

1 ' ' • ' 1 •' " 1 • • • • 1.

• • • •

*

"i. • •. 1 • . . . 1 • • . .

0 5 10 15 20 25 30 Comet (% tail DNA)

0 5 10 15 20 25 30

Comet (% tail DNA) Fig. 4.10 Linear regression analysis illustrating correlations for t>ehavioural (physiological), cytotoxic and genotoxic effects in Mytilus edulis following 7 d exposure to B(a)P. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

42,

150

100

50

oL

R =-0.066 P = 0.616

:.

0 0.1 0.2 0.3 0.4

C R ( L h - ^ )

Fig. 4.11 Linear regression analysis illustrating correlation for behavioural

(physiological) effects in Ophiothrix fragilis following 7 d exposure to B(a)P. The


limits.

4.4 Discussion

The results of these integrated studies illustrate that genotoxic effects are nriore

evident in echinodenns and nrK>lluscs than behavioural (physiological) or

cytotoxic effects following short term exposure to B(a)P. as both the Mn and

Comet assays showed deleterious effects of the PAH in the sea stars, sea

urchins and mussels. The only significant behavioural (physiological) and

cytotoxic effects were in the righting tinr es of the brittle stars (following 7 d

exposure) and the cell membrane stability (NRR assay) of the sea stars

(following 3 d exposure). The genotoxic consequences detected by the Mn

assay were not reversed following the 3d recovery period in the sea stars or

139

urchins, but these consequences were partly reversed in the mussels.

Conversely, the consequences measured by the Comet assay were not

reversed In the sea stars and mussels following the recovery period, but the sea

urchins did recover.

Little data is present in the literature which reports the physiological or

behavioural effects of B(a)P on marine invertebrates. The physiological Impact

of B(a)P on the crab Carcinus aestuarii, as measured by heart rate, has been

studied (Fossi et al., 2000). Following a 10 d exposure to 1 ppm (1 mg L'^) a

significant increase In the mean rate of heart beats was detected, but, no

explanation was given for the occurrence of this physiological response. There

is a small body of work which has investigated the effects of PAHs in general on

the behaviour (physiology) of echinoderms. A mixture of PAHs in the form of oil

drill cuttings (a by-product of oil drilling which consisted of a complex mixture of

oil-based drilling muds, emulslflers, dispersanls and stabilisers (Newton and

McKenzie, 1998a)) was found to impair the burrowing behaviour of two brittle

star species, namely Amphiura chiajei and A. filiformis. These brittle stars

exhibit a natural tendency to burrow Into soft sediments, but when presented

with sediments high in PAH content (up to 1,200 ppm) this intrinsic burrowing

behaviour was impaired (Newton and McKenzie, 1998a). The foraging

behaviour of the eleven-armed asteroid Coscinasterias muricata was disrupted

following exposure to a complex mixture of PAHs in the form of a water-

accommodated fraction (WAF) of crude oil (Georgiades et al., 2003). The ability

of the sea star to locate a food item (in the form of a mussel) was reduced

following exposure to WAF with a total PAH content of 1.8 mg L" for 4 d,

although this behavioural response returned to control levels following a

140

recovery period of 7 d. The researchers suggest that this behavioural response

was either as a result of the PAH mixture disrupting the motor responses of the

tube feet, or due to the suppression of the sea stars' metabolic processes

brought about by hydrocarbon narcosis (Georgiades et al., 2003). The

Australian sea star, Patiriella exigua has been shov^n to actively avoid crude oil

contaminated sediments in laboratory studies (Ryder et al., 2004). It was noted

that the sea stars exposed to the highest levels of contaminated sediments

(initial oil concentration of 114.000 pg g'^ in the sediment, falling to 800 pg g'^

after 32 d) were unable to right themselves when they became inverted in the

test vessels. It was postulated that this was due to a narcotic effect (i.e. the tube

feet became less responsive, that is. they possibly enter a state of stupor) of the

PAHs (Ryder et al.. 2004).

In the studies presented here the righting behaviour of the sea stars and sea

urchins remained unaffected by B(a)P, suggesting that either the exposure

duration, or concentrations were insufficient to induce a narcotic effect on the

tube feet of these asteroids and echinoids. The reduced self-righting ability of

the brittle stars after 7 d exposure indicated that either the metabolic processes

of the organisms were Impaired or narcosis occurred In the tube feet, resulting

In a reduced capacity for the tube feet to assist In the righting mechanism.

Cellular membrane stability (including that of the lysosomes). as measured by

the NRR assay, was only adversely affected in the sea stars in (the sea urchins

and mussels remained unaffected). This reduction was only evident following

3 d exposure across all concentrations (0.1. 1.0 and 10 pg L'^) and returned to

the negative control levels following 7 d exposure and 3 d recovery. This

141

recovery, whilst still being exposed to the contaminant, suggests that either the

cells of the sea star were being regenerated, the cellular concentrations of

B(a)P had reached a maximum after 3 d and therefore decreased by the 7 d

sampling point, or that the rate of B(a)P metabolism had increased. Lysosomal

integrity (as assessed by the NRR slide technique) was investigated following

18 d and 42 d exposure of the green-lipped mussel Perna viridis to 100 pg L'

B(a)P (Fang et al., 2008). The lowest lysosomal integrity in the green-lipped

mussels was observed after 6 d ( 5 1 % reduction in initial levels) of the 18 d

exposure. A complete recovery in lysosomal integrity was observed after 20 d in

clean seawater (Fang et al., 2008). This showed that the lysosomal damage in

the mussels was reversible and the integrity of the lysosomes adapted to the

presence of the xenobiolics. In vitro studies with mussel haemocytes (MytHus

galloprovincialis) exposed to 0.5 - 40 pg mL'^ B(a)P for 1 h found a clear

dose-response reduction in the amount of NR taken up by the cells as

measured by the NRR assay (Gomez-Mendikute et al., 2002).

Significant induction of Mn occurred in the coelomocytes of the sea stars and

sea urchins were detected after 3 d exposure at all concentrations and Mn

frequency remained significantly increased after the 3 d recovery period.

Conversely, the Mn frequency detected in the mussels returned to control levels

following the recovery period. So although all three species studied were

susceptible to Mn induction when exposed to B(a)P, the mussel haemocytes

were better equipped to recover from Mn induction than the echinoderm

coelomocytes. Previous work has shown that Mn were induced in both the gill

cells and haemocytes of M. galloprovincialis as a consequence of in vivo

exposure to B(a)P for 2 d (Venier et al., 1997). Dose dependent increases in Mn

142

numbers were evident at concentrations from 50 to 1000 pg L" with higher

frequencies in the agranular haemocytes compared to the main gill cells.

Increased Mn frequency were observed in the haemocytes of the Carcinus

aestuarii ai[er a 10 d exposure to 1 mg L" of B(a)P (Fossi et al., 2000).

Increased levels of single strand DNA breaks (as quantified by the alkaline

Comet assay) were evident In the sea stars following 7 d exposure and

remained elevated following the 3 d recovery period. Increased levels were also

found in the urchins following 3 d and 7 d, but returned to baseline levels after

the recovery period. The mussels appeared to be the most sensitive of the three

speciess as the degree of DNA breaks showed a significant Increase after 3 d

and remained so at the end of the recovery period. Past work has found that

concentrations of 0.5 and 3.0 pg L" B(a)P for 6d caused an increase In the

frequency of DNA strand breaks In the gill tissue and digestive gland of the

scallop Chfamys farreri as detected by the alkaline unwinding assay (similar to

the Comet assay) (Pan et al., 2008). There was then a recovery to control levels

in the gill tissue, but not In the digestive gland following an additional 14 d

exposure. No increase in DNA strand breaks was detected at any time point

during 20 d exposure at 10 pg L'V There was no increase in DNA damage as

detected by the alkaline unwinding assay in the haemocytes of the crab

Carcinus aestuarii following a 10 d exposure to 1 mg L' of B(a)P (Fossi et al..

2000).

B(a)P needs to be metabolised in vivo in order to produce a toxic effect.

Although metabolism of PAHs is not fully understood in invertebrates, the

process is thought to be similar to that of vertebrate species (Livingstone,

143

1991). Figure 4.12 represents a schematic illustrating the four possible

molecular pathways involved in the activation of B(a)P into DNA reactive

intermediates In a model mammalian system under in vivo conditions.

b e n z o l a j p y r c n c

7 . B - c p o x i d c

r iyafc!ys s

7 .8 -d io l

:::r..v- v'^iur.

S - m c t h y l - i n l c f m c r l i ; i l c

/

mcnoGrygcnniion c a t e c h o l

s c m i q u i n o n c a n i o n r a d i c a l R O S

q u i n o n c s

o n c - o i e c t f o r oti-jauo- R r O S O i O H O P O ^ O H O C O C H . .

© d i o l o p o x t d c s r a d i c a l c a t i o n s b c n z y l i c c a r b c n i u m ton:

ON'M L E S l O r J S ( O N A a d d u c i s . o-icair.'a a a - n a g e dcpllr^.^a:lOn s J r a r a c r e a k s -

T ' J M O R I N ' l T l A l l O N

Fig. 4.12 Four possible pathways for the activation of B(a)P leading to DNA

lesions in vivo In a mammalian cell system (from Canova et al.. 1998).

144

Pathway 1 is based on the 'bay-region theory' in which epoxides, phenols, diols

and tetrols are formed through the monoxygenation of B(a)P by cytochromes

P450. Diols produced in pathway 1 can be dehydrogenated by dihydrodiol

dehydrogenase to produce catechols, quinones which can undergo redox

cycling to produce reactive oxygen species (ROS). DNA reactive radical cations

may also be formed by one electron oxidation (pathway 3). This pathway may

be mediated by cytochromes P450. The final potential cause of B(a)P activation

leading to DNA lesions are benzylic carbenium ions. These ions are formed

when a 6-methyl-intermediate of B(a)P undergoes either one electron oxidation

or formation of methyl ester derivatives (Canova et al., 1998).

Cytochromes P450 are a large and versatile group of proteins which are

involved in Phase 1 metabolism of many lipophilic xenobiotics including PAHs

(Rewitz et al., 2006). This metabolism is involved in both the detoxification and

activation into toxic derivatives (Stegeman, 1985). Evidence of P450 enzyme

activity and associated mixed-function oxygenase (MFC) activities have been

found in Cnidaria, Annelida. Mollusca, Arthropoda and Echinodermata

(Livingstone, 1991). The differential phylogenetic sensitivity of the organisms

tested in these studies may be as a result of their differing abilities to metabolise

B(a)P. Workers have found that M. edulis can metabolise 20-30 % of the B{a)P

they are exposed to, compared to polychaete worms Nereis spp. which can

metabolise 86-96 % (Rust et al.. 2004).

There is a paucity of data regarding the role and function of P450 in

echinoderms. Although there is evidence that cytochromes P450

monooxygenase system is found in asteroids, echinoids and holothuroids and it

145

Is not only Involved in the detoxification of xenoblotics, but also the metabolism

of steroids (den Besten, 1998). Cytochromes P450, along with other

components of the monooxygenase (MO) system, have been detected in

asteroids (A. rubens and Marthasterias glacialis), holothurians {Holothuria

forskali), and echinoids {Strongylocentrotus sp.) (Payne and May, 1979; den

Besten et al., 1990). As the B(a)P hydroxylase activity In the pyloric caeca of A.

rubens was found to be inhibited by compounds known to be Inhibit vertebrate

cytochromes P450, It was concluded that the echlnoderm mixed-function

oxygenase (MFC) system has more similarities to that found in vertebrates and

crustaceans than that of the molluscs (den Besten et al., 1990).

These integrated studies carried out on a range of echlnoderm species and a

bivalve mollusc found that B(a)P did cause deleterious effects at different levels

of biological organisation. These effects were most noticeable at the

sub-cellular level, as measured by the Comet assay. This merits further

investigation with the implementation of the modified Comet assay - a technique

which not only measures single strand DNA breaks, but also detects oxidative

DNA damage. This technique used in conjunction with a measure of

immunocompetence (the phagocytosis assay) would add to the paucity of data

available detailing any possible links between oxidative stress and

immunocompetence In marine invertebrates, especially the echinoderms.

146

Chapter 5

A heavy metal (cadmium) can cause behavioural (physiological) and

cytotoxic consequences in echinoderms (>As/enas rubens) and

molluscs {Mytilus edulis), and reversible immunotoxic, genotoxic and

oxidative stress consequences in both species.

147

5.1 Introduction

Heavy metals (a group of metals consisting of arsenic, cadmium, chromium,

cobalt, copper, lead, mercury and nickel) are one of the most widespread and

serious forms of environmental pollution (Pruski and Dixon, 2002). Heavy

metals enter the biosphere through volcanic activity, the weathering of rocks

and as a result of human activities including the release of sewage, mining and

the combustion of fossil fuels (Depledge et al., 1994). Most types of pollution,

regardless of its initial source, ultimately finds its way into the aquatic realm and

heavy metals are no exception. Cadmium (Cd) is one such heavy metal and it is

considered a priority pollutant due to its toxicity (USEPA, 1987). Cadmium has a

range of consumer and industrial uses. It is used in the protective plating of

steel; as a stabiliser in the production of poly-vinyl chlorides (PVCs); as a

pigment in glasses and plastics; and as an electrode material in nickel-cadmium

batteries and as a component in various alloys (WHO, 1992). Cadmium is a

nonessential metal, meaning that it is not essential to any organism (Muller et

al.. 1998). Although Cd is present in seawater at trace levels (Erk et al., 2005),

anthropogenic inputs can result in elevated levels in coastal marine

environments. The total input of Cd into the marine environment is estimated at

8000 tonnes per year, about half of which is believed to from anthropogenic

sources (Coles et al.. 1995).

There is a large pool of literature which has investigated the effects of Cd on a

range of aquatic organisms (for example; the water flea, Daphnia magna (den

Besten and Tuk, 2000); the aquatic clam, Corbicula fluminea (Tran et al.. 2001;

Tran et al., 2002); the Pacific oyster. Crassos/rea gigas (Geret el al., 2002); the

freshwater gammarid, Gammarus pulex (Felten et al.. 2008); turbot,

148

Scophthalmus maximus (Bolognesi et al., 2006); carp, Cyprinus carpio (Zhang

et al., 2007); Nile tilapia, Orecochromis niloticus (Almeida et al.. 2009) and the

white shrimp, Litopenaeus vannamei (Chang et al., 2009)). These studies have

investigated the effects of Cd at a wide range of levels of biological

organisation, from physiological and behavioural responses, through to

cytotoxicity, immunotoxiclty, oxidative stress and genotoxicity.

This study utilised the adult stages of the common sea star Asten'as rubens and

the blue mussel Mytilus edulis to determine the effects of Cd at different levels

of biological organisation. Increased levels of MTs have been quantified (den

Besten et al., 1989; Temara et al., 1997b; Erk et al., 2005) and reduced cellular

immune response (Coleur et al., 2005b) following Cd exposure in the sea star.

Toxicology studies with mussels that investigated the effects of Cd have

implemented a range of endpoints, including genotoxicity (Bolognesi et al.,

1999; Emmanouil et al., 2007), cytotoxicity (Dalllanis, 2009) and the induction of

MTs (Beblanno and Langston. 1991; Geret et al., 2002; Erk et al., 2005). There

is a lack of information available regarding the effects of Cd in the two named

species in terms of whole organism effect (physiological and/or behavioural

effects), cytotoxicity, immunotoxicity, Induction of oxidative stress and

genotoxicity. Therefore the aims and objectives of this study were:

(a) To determine If Cd exposure induced a response at higher levels of

biological organisation (i.e. the whole organism and cellular levels) in either of

the two test species.

149

(b) To determine if there is a relationship between any immunocompetence,

genotoxicity and/ or oxidative damage.

(c) To Investigate the recovery capability of the organisms to any incurred

immunotoxicity and/ or oxidative stress.

In order to achieve these objectives a series of in vivo experiments were

conducted. The first studies assessed the effects of a five day exposure of Cd

to A. rubens and M. edulis using behavioural / physiological (righting time and

clearance rates, respectively for the two species) to determine any deleterious

effect at the whole organism level. The neutral red retention (NRR) assay was

used along side these assays to determine any cytotoxicity of the heavy metal.

Chemical analysis In the form of Inductivity Coupled Plasma Mass Spectrometry

(ICP-MS) was carried out on exposure water samples and different body

compartments of the test species to determine any bioaccumulatlon. The

second suite of studies used the same exposure conditions, with the exception

of the implementation of a one and a five day 'recovery period' following a five

day Cd exposure. The assays employed were the phagocytosis assay (to

determine immunocompetence) and the modified Comet assay (to determine

genotoxicity and oxidative stress). By sampling on the recovery days it was

possible to determine the respective repair capabilities of the two species with

regard to immunotoxicity. genotoxicity and oxidative stress caused by the Cd

exposure.

150

5.2 Materials & methods

5.2.7 Chemicals


Chemical Co. (Poole, Dorset, UK). Cadmium (Cd) was used in the form of

calcium chloride hydrate (CdCb hydrate, CAS number 7790-78-5; purity

>98 % ) , it was readily soluble in water at room temperature, 133 mg mL'^

(http;//dohs.ors.od.nih.gov), therefore no solvent vehicle was required. All

glassware was acid-washed (5 % HNO3 for at least 2 h) and triple rinsed in

deionised water prior to use.


Asterias rubens were hand collected by SCUBA divers during November 2008

from in and around Plymouth Sound, Devon. UK (Latitude 50.20° N; Longitude

4.10°W). They were 60 - 100 mm in diameter and were held under the

conditions stated in Section 2.1.1. Adult M. edulis (40 - 45 mm in length) were

collected during November 2008 from Port Quin (Latitude 50.35° N; Longitude

4.52° W), they were held in conditions outlined in Section 2.1.4.

5.2.3 Exposure of sea stars and mussels to cadmium

Due to logistical constraints the exposures for both species were separated into

two individual studies (therefore four separate exposures in total). In all cases

the exposure concentrations used were identical; seawater control, 100, 180

and 320 pg L" of Cd. The Cd concentrations were selected based on those

used in previous work using mussels (Bolognesi et al.. 1999; Geret et al., 2002;

Emmanouil et al., 2007) and sea urchins (Warnau et al., 1995), whilst still

incorporating the semi-logarithmic scale routinely used in ecotoxicity testing.

151

The first set of exposures for both species ran for 5 d. After the 5 d exposure

period the RT of the sea stars (Section 5.2.4) and CR of the mussels (Section

5.2.5) were determined. Coelomocyte and haemocyte samples were then

collected and the cellular membrane stability of the test organisms were

determined using the NRR assay (Section 5.2.7). Exposure water samples were

collected at the initiation of the study and following 5 d to determine Cd

concentrations at the start and termination of the exposure (Section 5.2.10.1).

Following collection of the coelomocyte and haemocyte samples different body

compartments were dissected from the two species in order to determine Cd

body burdens (Section 5.2.10.2). This experimental protocol can be seen in Fig.

5.1a and 5.1b (n = 5 in all cases). The second suite of exposures incorporated

two recovery periods (1 d and 5 d recovery) following an initial 5 d exposure to

Cd. On the three different sample days the phagocytosis assay (Section 5.2.8)

and modified Comet assay (Section 5.2.9) were used to determine the

immunocompetence, genotoxicity and oxidative stress, respectively {n = 3). The

protocol for these experiments can be seen in Fig. 5.2a and 5.2b.

The exposures were carried out at 15 ± 0.5 °C and food was withheld during the

exposure period. The exposure vessels were individually aerated using Pasteur

pipettes and were sealed with Parafi lm'". The vessels were checked daily for

mortalities and the parameters of temperature, salinity, dissolved oxygen and

pH measured.

152

(a) Sea stars individually exposed in 7 L vessels [n = 5)

Aqueous Cd concentrations determined at start and after 5 d of exposure (n = 5)

Righting times determined after 5 d (n = 5)

coelomocytes _ • _ collected

Neutral red assay carried out (n = 5)

Body compartments dissected (gut, pyloric caeca & tube feet) for Cd body burden analysis

(n = 5)

(b) Mussels individually exposed in 1 L vessels in = 5)

Aqueous Cd concentrations determined at start and after 5 d of exposure (/? = 5)

Clearance rates determined after 5 d (n = 5)

haemocytes collected

Neutral red assay carried out (/7 = 5)

Body compartments dissected (gut & gills) for Cd body burden analysis {n = 5)

Fig. 5.1 Flow chart Illustrating the Integrated experimental designs adopted in the studies to evaluate the behavioural (physiological) and

cytotoxic effects of cadmium (as well as body burdens) in (a) sea stars and (b) mussels.

153

(a) Sea stars Individually exposed in 7 L vessels (3 replicates per concentration)

Following 5 d exposure, 1 d & 5 d recovery 3 sea stars sampled per concentration

coelomocytes collected

Phagocytosis and modified Comet assays carried out (/? = 3)

(b) Mussels Individually exposed in 1 L vessels (3 replicates per concentration)

Following 5 d exposure, 1 d & 5 d recovery 3 mussels sampled per concentration

haemocytes collected

Phagocytosis and modified Comet assays carried out (n = 3)

Fig. 5.2 Flow chart Illustrating the integrated experimental designs adopted In the studies to evaluate the Immunotoxic, genotoxic and

oxidative stress effects of cadmium in (a) sea stars and (b) mussels.

154


Righting times (RTs) of the sea stars were recorded following 5 d exposure to

Cd. RT was determined prior to coelomocyte sample collection (5 sea stars per

concentration) as detailed in Section 2.10. The mean of three times per animal

was calculated.

5.2.5 Determination clearance rate (CR)

Clearance rates (CRs) were determined (5 mussels per concentration), prior to

haemolymph sample collection at the end of the 5 d exposure (as per the

method described in Section 2.9.1). The results were presented in L h" \


Coelomic fluid and haemolymph samples were collected as per Sections 2.2.2.1

and 2.2.1, respectively. The coelomic fluid/ haemolymph samples were held on

ice until used for the cytotoxicity, immunocompetence, genotoxicity and

oxidative stress assays. The samples were thoroughly mixed by vorlexing prior

to use.


The cellular membrane stability of the sea star coelomocytes and the mussel

haemocytes (both n = 5) were determined using the NRR assay following the

5d exposure. This assay was earned out as per Section 2.3 and triplicate

samples were assessed for each individual organism and the mean optical

density (at 550 nm) mg'^ protein was determined.

155

5.2.8 Determination of immunotoxicity using phagocytosis assay

The phagocytic activity of coelomocytes and haemocytes was determined by

measuring the uptake of zymosan particles dyed with neutral red dye (as

detailed in Section 2.5). This activity was quantified following 5 d exposure, 1 d

and 5 d recovery periods. Triplicate samples were assessed for each individual

organism (n = 3 for both species) and the mean calculated.

5.2.9 Determination of genotoxicity and oxidative stress using the modified

Comet assay

Determination of the induction of DNA single strand breaks (SSBs) and alkali

labile sites of coelomocytes and haemocytes were determined using the

modified Comet assay, as per the method described in detail in Section 2.8.1.

This was carried out following 5 d exposure, 1 d and 5 d recovery periods, with

three organisms per concentration. Three slides (each with 2 microgel

replicates) were prepared for all the test organisms, one slide was an 'enzyme

buffer' control (to detect single strand DNA breaks), one slide Endo III enzyme

treatment (to detect oxidised pyrimidlnes) and a FPG enzyme treatement slide

(for the detection of oxidised purines). As a required prerequisite for the

modified Comet assay, it was necessary to determine the cellular viability of the

coelomocytes and haemocytes. To this end, the Eosin Y dye exclusion assay

was used as detailed in Section 2.7 - only cell samples with a viability of

> 70 % were used in the modified Comet assay.

156

5.2.70 Chemistry (ICP-MS) analysis

5.2.10.1 Aqueous cadmium concentrations

At the start and following 5 d of the initial sea star and mussel exposures water

samples were taken from all of the exposure vessels. To this end, 10 mL of the

exposure water was removed and placed in polythene screw-top digestion vials

(scintillation vials, Simporl, Canada) (acid-washed ( 5 % HNO3 for at least 2 h)

prior to use). The samples were stored at 4 ''C prior to Inductivity Coupled

Plasma Mass Spectrometry ( ICP-MS) analysis as detailed in the next section

(Section 5.2.10.2).

5.2.10.2 Cadmium body burdens

Prior to Cd body burden analysis it was necessary to dissect the appropriate

body compartments from the sea stars (gut, pyloric caeca and tube feet) and

the mussels (gut and gills). The methodology followed was detailed in Sections

2.11.1 and 2.11.2. Once the tissue samples were dried to a constant weight

they were placed in polythene screw-top digestion vials and digested in 1 mL

concentrated HNO3 in a water bath (70 °C for 2 h). Once cooled, the samples

were diluted with 4 mL of Milli-Q water. Tissue digest aliquots were stored at

room temperature prior to ICP-MS analysis (this process is described In Section

2.11.3) . The software used with ICP-MS was PQ Vision 4.1.2 (Plasma Quod

PQ2 Turbo, Thermo Elemental. Winsford, UK). The data for aqueous Cd

samples was given In pg L' and the tissue digest data in mg kg'^ (Section

2.11.4) .

157

5.2,11 Statistical analysis

Statistical significance between means was determined using one-way analysis

of variance (ANOVA). All values are provided as means ± SEM. Differences

between means which were P < 0.05 were considered significant. All data sets

were analysed using Statgraphics v5.1 software (StatSoft. USA). Any

correlations been variables were determined using the Pearson correlation

coefficient.

5.3. Resul ts


Water quality parameters measured daily during all the exposures were in the

normal acceptable range. Temperature ranged from 14.5 to 15.5 °C, salinity

from 34 to 35 PSU, dissolved oxygen levels remained above 70 % saturation

and pH from 8.1 to 8.4.

5.3.2 Chemical analysis

The concentrations of Cd in the exposure water of the M. edulis test vessels

decreased markedly from the start of the exposure through to the final day (5 d).

This trend was not followed in the A. rubens exposure vessels, where Cd

concentrations remained relatively consistent during the 5 d exposure period

(Fig. 5.3). On average the tissue burdens of the mussels were a factor of 10

times greater that those found in the sea stars (Figs. 5.4a and 5.4b). Linear

regression analysis found no significant relationship between the amount of Cd

'lost' during the 5 d exposure from the A. rubens vessels and the amount taken-

up by the three body compartments (gut, pyloric caeca and the tubed feet)

analysed (Fig. 5.5a - c). Conversely moderately strong correlations were

158

observed for the M. edulis vessels and both the gut (R = 0.831. P < 0.001) and

the gills (R = 0.632. P = 0.003) (Fig. 5.6a and b).

159

CD O

I I .

0) o c o o •D o

400

300

o 2 200

100

M. edulis dO M. edulis d5 A. rubens dO A.rubens d5

0 '! Control 100 180 320

Initial nominal Cd concentration (^g L' )

Fig. 5.3 Cadmium concentrations in exposure vessels at ttie beginning (0 d) and following end (5 d) of static exposure of Mytilus edulis

and Asterias rubens for 5 d to nominal concentrations of 100, 180 and 320 pg L' of cadmium. The values are ttie means ± S E M (rj = 5).

if) C o CO

c a; u

11 o <D 3

20 1

15

10

5

Gul 1 I Pyloric caeca H i Tube feet

(I))

Control 100 180 320

Cd concentration (pg L" )

to c o

c 0) (J

If u 0) D ifi {/)

350

300

250

200

150

100

50

Gut C = ) Gills

Control 100 180 320

Cd concentration (pg L'')

Fig. 5.4 Cadmium-body burdens for different body components In (a) Asterias rubens and (b) Mytilus edulis following 5 d exposure to

nominal concentrations of 100, 180 and 320 pg L' of cadmium. The values are the means ± SEM (/7 = 5).

(a)

32

^ 24

i I i 6F u

O M 0) 3 cr <

R = 0.250 P= 0.287

0 4 8 12 16 20 24 Tissue Cd burden (mg kg-i)

( b )

32 R = 0.232 P= 0.325

0 4 8 12 16 Tissue Cd burden (mg kg-')

1

g E SJ •o a (0 o 0)

(c)

32 - 2 4

c o

R = -0.004 P= 0.988

0 3 6 9 12 15 18

Tissue Cd burden (mg kg- )

Fig. 5.5 Linear regression analysis illustrating correlations for amount of Cd removed from aqueous solution in the exposure vessels over

the 5 d exposure period (pg L' ) and Cd body compartment burdens (mg kg' ) for the (a) gut, (b) pyloric caeca and (c) tubed feet In

Asterias rubens following 5 d exposure to Cd. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

(a) ( b )

g 100 o

80 Q) ^

E S

§ °

60

40

20

0

.1 • • ' ' T - 1 • • 1 ' ' ' ' 1 • * . ' • "

-

^ ' ^ a

• ' ' ^ * '" * "

^ /

y

R = 0.831 • -r-.''. . . P< 0.001

0 100 200 300 400 Tissue Cd burden (mg kg-"")

E 100 o

• 4 —

0) 80 > O

1

_ l E O)

3 . 60

O tion

40 C/) 3

O O 0)

(A 20 3

< 0

R = 0.632 P= 0.003

J , , L

0 30 60 90 120 150 Tissue Cd burden (mg kg-"')

Fig. 5.6 Linear regression analysis illustrating correlations for amount of Cd removed from aqueous solution in the exposure vessels over

the 5 d exposure period (pg L' ) and Cd body compartment burdens (mg kg" ) for the (a) gut and (b) gills in Myf//us edufe following 5 d

exposure to Cd. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

5.3.3 Behavioural (physiological) responses

All of the sea stars successfully self-righted within the allotted 600 s for every

RT assessment. The RT of the sea stars and CR of the mussels did not show a

significant Cd-induced effect across all concentrations (P = 0.838 and 0.180,

respectively) (Figs. 5.7a and b). Correlation analysis found no relationship

between the RT of the sea stars and the Cd body burdens measured in each of

the body compartments (RT : gul, R = -0.027, P = 0.912; Fig. 5.8a; RT ; pyloric

caeca. R = 0.133, P = 0.636; Fig. 5.8b; RT : tubed feet, R = -0.005, P = 0.985;

Fig. 5.8c). Equally, no significant correlations were found between the CR and

Cd body burdens in the mussels (CR : gut, R = 0.147, P = 0.535; Fig. 5.9a; CR ;

gill. R = -0.126, P = 0.596; Fig. 5.9b).

5.3.3 Cytotoxic responses

No significant cytotoxic response was observed in either species (sea stars

P = 0.275, mussels P = 0.749) as measured using the NRR (Figs. 5.10a and

5.10b). No correlations were found between the behavioural and cytotoxic

responses in either species {A. rubens: R = 0.069. P = 0.772; Fig. 5.11a,

M.edulis: R = -0.093, P = 0.696; Fig. 5.11b). Correlation analysis found no

relationship between the NRR of the sea stars and the Cd body burdens of the

body compartments (NRR : gut, R = 0.361, P = 0.118; Fig. 5.12a; NRR : pyloric

caeca, R = 0.036, P = 0.880; Fig. 5.12b; RT : tubed feet, R = -0.077, P = 0.747;

Fig. 5.12c). No statistically significant correlations were found between the NRR

and Cd body burdens In the mussels (NRR : gut, R = -0.192, P = 0.416; Fig.

5.13a; NRR : gill. R = -0.101, P = 0.673; Fig. 5.13b).

164

cn

(a)

200

160

I 120 i -

•I 80

40

i

Control 100 180 320 Cd concentration (ijg L' )

(b)

3.0

^ 2.5 1—

^ 2.0 Q) "5 0 o c ni \— nj Q) O

1.5

1.0

0.5

0.0 Control 100 180 320

Cd concentration ( jg L'')

Fig. 5.7 Behavioural (physiological) effects in (a) Asterias rubens (righting time) and (b) Mytilus edulis (clearance rate) exposed to

Cd for 5 d. Concentrations are expressed as pg l ' \ The values are means ± SEM (n = 5).

(a)

I-

0 4 8 12 16 20 24

Tissue Cd burden (mg kg-^)

( b )

220 • R =-0.027' 220

200 • P = 0.912..- 200

180 180

160 • • •

iT(s

;

160

140 A 140

120 120-

100 1 • 1 ' ' 100-

R = 0.113 P = 0.636

0 4 8 12 16

Tissue Cd burden (mg kg-i)

(c) 220

« R = -0.005 200 • P = 0.985...

180 M 1- 160

"T

140 •

120 ' • " V .

100 " i . . 1 . . I . . I . . 1 , . 1 . 1

0 3 6 9 12 15 18


Fig. 5.8 Linear regression analysis illustrating correlations for behavioural (physiological) effects and cadmium body compartment

burdens (mg kg' ) for the (a) gut, (b) pyloric caeca and (c) tubed feet in Asteiias rubens following 5 d exposure to Cd. The solid line

is a linear regression and the dashed lines represent 95 % confidence limits.

(a) (b )

0.5[- R = 0.147 P = 0.535

0 100 200 300 400

Tissue Cd burden (mg kg-^)

a: o

2.5

2

1.5

1

0.5

OL R =-0.126 P = 0.596

0 30 60 90 120 150 Tissue Cd burden (mg kg-i)

Fig. 5.9 Linear regression analysis illustrating correlations for behavioural (physiological) effects and cadmium body compartment

burdens (mg kg" ) for the (a) gut and (b) gill in Mytilus edulis following 5 d exposure to Cd. The solid line is a linear regression and

the dashed lines represent 95 % confidence limits.

(a)

CD

1.2 1

2 E 0.6

f 1 E 0.4

o 2. 0.2

0.0

(b)

Control 100 180 Cd concentration (pg L ')

320

1.0

0.8

^ E

E c 0 4

^ in

O O 0.2

0.0 Control 100 180

Cd concentration (pg L" ] 320

Fig. 5.10 Cytotoxic effects (the neutral red retention assay) in (a) Asterias rubens and (b) Mytilus edulis exposed to cadmium for

5 d. Concentrations are expressed as pg l ' \ The values are means ± SEM {n = 5).

CD

(a)

220 200

^ 180

(A

H 160

140

120

100

R = 0.069 P = 0.772

0.55 0.75 0.95 1.15 1.35

NRR (opticaldensity mgprotein-i)

( b )

2.5

0.5

0

R = -0.093 F = 0.696

0 0.3 0.6 0.9 1.2 1.5

NRR (opticaldensity mgproteln-i)

Fig. 5.11 Linear regression analysis illustrating correlations for behavioural (physiological) and cytotoxic effects in (a) Asterias

rubens and (b) Mytilus edulis following 5 d exposure to Cd. The solid line is a linear regression and the dashed lines represent 95 %

confidence limits.

(a) (b ) (c)

- J o

.E 1.35 a> 2 S 1.15 E

S 0.75 Q.

a: a:

0.55

R = 0.361 P = 0.118

s 0.95

0 4 8 12 16 20 24


.9 1.35

2 S 1-15 E

• f 0.95 c

1 O

0.55

R = 0.036 P= 0.880

0 4 8

Tissue Cd burden

12 16

(mg kg-1)

g 1.35 c 0) o Q. O) E

c

0)

(0

1.15

0.95

0.75

0.55

R = -0.077 P= 0.747

0 3 6 9 12 15 18 Tissue Cd burden (mg kg-*")

Fig. 5.12 Linear regression analysis illustrating correlations for cytotoxic effects and cadmium body compartment burdens (mg kg" )

for the (a) gut. (b) pyloric caeca and (c) tubed feet in Asterias rubens following 5 d exposure to Cd. The solid line is a linear

regression and the dashed lines represent 95 % confidence limits.

(a) (b)

1 c

1.5 Q) "5 k_ 1.2 Q.

E 0.9

c 0.6 0)

0.6 TD 16 O 0.3

V-»

0 a:

z

R =-0.192 P = 0.416

0 100 200 300 400


I Q.

E

c

0)

CO

o •s . a:

1.5

1.2

0.9

0.6

0.3

0

R =-0.101 P = 0.673

0 30 60 90 120 150


Fig. 5.13 Linear regression analysis illustrating correlations for cytotoxic effects and cadmium body compartment burdens (mg kg" )

for the (a) gut and (b) gill in Mytilus edulis following 5 d exposure to Cd. Ttie solid line is a linear regression and the dashed lines

represent 95 % confidence limits.

5.3.4 Immunocompetence responses

The phagocytosis assay showed a reduced immunocompetence response after

5 d for the sea stars at the 100 and 180 pg L' exposure levels when compared

to the control means (P = 0.031). This response was still impaired after 1 d in

clean sea water (P = 0.002). but returned to control levels after the 5 d recovery

period (P = 0.169) (Fig. 5.14a). The exposed mussel haemocytes showed a

reduced mean phagocytic capacity across all concentrations after 5 d

(P = 0.012), there was no difference between the concentrations. This response

returned to control levels following 1 d recovery (P = 0.467) and remained so

after 5 d recovery ( P = 0.972) (Fig. 5.14b).

5.3.5 Genotoxic and oxidative stress responses

Cellular viability in the coelomocyles and haemocytes (as measured by the

exclusion of Eosin Y) exceeded the required 70 % viability required for use In

the modified Comet assay. For the sea stars a significant increase in DNA

SSBs as measured by the buffer control Comet assay slides was observed after

the 5 d exposure period for the two highest concentrations (P = 0.0004; Fig.

5.15a). The percentage tail DNA of the exposed coelomocytes returned to

control levels after the I d recovery period and remained there after the 5 d

recovery period (Fig. 5.17a). Significant increase in oxidised pyrimidines (Endo

III enzyme) ( P = 0.025) were detected for the 180 and 320 pg L" exposed sea

stars and an increase in the amount of oxidised purines (FPG enzyme) (P =

0.034) were measured at all exposure concentrations (Fig. 5.15b and c). As

with the DNA SSB these returned to control levels following 1 d recovery in

clean seawater. Figure 5.16 illustrates the repair of DNA damage detected by

the modified Comet assay. The figure is based on the same data presented in

172

Fig. 5.15, but better illustrates the repair kinetics over the recovery periods of

the appropriate modified Comet conditions (buffer control. Endo II and FPG) as

opposed to the dose-effects illustrated in the former figure. A moderately strong

correlation was found between immunoloxic (phagocytosis assay) and

genotoxic responses (modified Comet assay - buffer control) following 1 d

recovery (R = -0.579, P = 0.049; Fig. 5.17b). No correlations were found for the

Endo III and FPG treated Comet slides and the phagocytic response of the sea

stars (Figs. 5.18 and 5.19).

There was no statistically significant increase in DNA strand breaks for the

M. edulis haemocytes following 5 d exposure ( P = 0.072), although the mean

values appeared higher in the exposed mussels compared to the controls. But,

following the 1 and 5 d recovery periods significances increases were observed

(P = 0.008 and 0.001) (Fig. 5.20a). The enzyme Endo III delected a significant

increase in oxidised pyrimidines following 5 d exposure and 1 d recovery

(P = 0.001 and 0.004). The values returned to control levels after 5 d recovery

(P = 0.103) (Fig. 5.20b). A significant increase in oxidised purines (FPG

enzyme) were also observed after 5 d exposure and 1 d recovery (P = 0.029

and 0.001, respectively). The values returned to control levels after 5d recovery

( P = 0.419) (Fig. 5.20c). Figure 5.21 shows the repair kinetics of DNA damage

delected by the modified Comet assay. Moderately strong correlations were

present between immunoloxic (phagocytosis assay) and genotoxic/ oxidative

stress responses (modified Comet assay) following 5 d exposure for all the

modified Comet assay conditions, i.e. buffer control (R = -0.604, P = 0.038;

Fig. 5.22a), Endo III (R = -0.696, P = 0.012; Fig. 5.23a) and FPG (R = -0.704,

173

P= 0.011; Fig. 5.24a). No statistically significant correlations were present for

any of the recovery periods.

174

(a)

(J -o "c

w °-(U O)

y E

U o E

N

2.5 1

^ 2.0

1.5

1.0

0.5

0.0

d5 ' ' Recovery d1 warn Recovery d5

Control 100 180

Cd concentration (jjg L' )

320

(b)

u -

ro o E N

2.5

X. 2.0

1.5

1.0

0.5 ^

0.0

I I Recovery d1 • • 1 Recovery d5

Control 100 180 320

Cd concentration (fjg L'')

Fig. 5.14 Immunotoxic effects in (a) Asterias rubens and (b) MytHus edulis exposed to cadmium for 5 d. Concentrations are expressed as

pg L'V The values are means ± SEM (/? = 3). * Indicates a significant difference from controls at, or above, the 95 % confidence level.

(a) 40

(b)

(c)

d5 I I Recovery dl

RecovervdS

Control 100 180 320

Control 100 180 320

Control 100 180 320

Cd concentration (pg L' )

Fig. 5.15 Genotoxic effects in Asterias rubens exposed to cadmium for 5 d and

following 1 d and 5 d recovery periods. Modified Gomel assay conditions: (a) enzyme

buffer control; (b) Endo III enzyme and (c) FPG enzyme. Concentrations are expressed

as Jg L'\ The values are means ± SEM {n = 3). ' indicates a significant difference from

controls at, or above, the 95 % confidence level.

176

(a)

(b)

40

35

30

< Q 25

^ 20

15

10

5

40

35

30

< Q 25

J 20

15

10

5

Control

d5 Rdl Rd5

Control 100 180

d5 Rd1 Rd5

(c) 35

30

< 25

^ 20

^ 15

10

Control

d5 Rdl Rd5

Fig. 5.16 Repair of single strand DNA breaks (a).oxidised purines (b) and oxidised pyrimidines (c) by Asterias rubens coelomocytes in vivo following 1 d (Rd1) and 5 d (Rd5) recovery periods after a 5 d exposure to cadmium (d5). Concentrations are expressed as pg L V The values are means ± SEM (n = 3).

177

( »

I a> c (0 s Q. 2 M Q. " E

Q . O

o E

2.5

2

1.5

1

0.5

• v .

R»-0.150 P= 0.643


(b)

R =-0.579 P= 0.049

w Q. 0.8

1

N 10 14 18 22

Comet(% tail DNA)

(c)

2 ^ 1.3

o E

0 5

1 • • • 1 • • • 1 J

• •

/ .* R = 0.380

• j — ^ — - • -

P= 0.223 • 1 — - — » — • — . — • — » '

10 12 14 16 18 Comet (% tail DNA)

Fig. 5.17 Linear regression analysis illustrating correlations for immunotoxic (phagocytosis assay) and genotoxic responses (modified

Comet assay buffer enzyme control treatoDent) in Asterias rubens following (a) 5 d exposure to cadmium, (b) 1 d recovery and (c) 5 d

recovery. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

(«)

2.5

E >> N

R«-0.118 P=0.714

« o» 15


2.0 0,6

o E

0 4

R = -0.460:

• •

•

P=0132

— .. fc—«


1.3

' . • • • T • • • T J

< • •

R = 0.116 •

"l • • • 1 • 1 •

P= 0.720 . i • • . 1 . • . 1 . . . 1

8 10 12 14 16 18

Comet (% tail DNA)

Fig. S.18 Linear regression analysis illustrating correlations for immunotoxic (phagocytosis assay) and genotoxic responses (modified

Comet assay Endo III enzyme treatment) in Asterias rubens following (a) 5 d exposure to cadmium, (b) 1 d recovery and (c) 5 d recovery.

The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

C30 O

TO Q.

(b) (c)

R =-0.277 P = 0.384


R =-0.165 P = 0.609

9> a> 0.8

Q - O 0.6

in CL 0)

Comet (% tail DNA)

O

I 0 5

R =-0.203 P= 0.528

10 14 18 22 Comet (% tall DNA)

Fig. 5.19 Linear regression analysis illustrating correlations for immunotoxic (phagocytosis assay) and genotoxic responses (nwdified

Comet assay FPG enzynte treatment) in Asterias rubens following (a) 5 d exposure to cadmium, (b) 1 d recovery and (c) 5 d recovery.


(a) 50

40

* *

< Q TO

(b) 80

60

< Q ^ 40

20

d5 Recovery d1 Recovery d5

0 -L

) 70 1

60

50 < Q 40 • CD

1 - 30

20

10

0

Control 100 180 320

* * * * * *

Control 100 180 320

* * * * * *

Control 100 180 320 Cd concentration {pg L'')

Fig. 5.20 Genotoxic effects in Mytilus edulis exposed to cadmium for 5 d and following 1 d and 5 d recovery periods. Modified Comet assay conditions: (a) enzyme buffer control; (b) Endo III enzyme and (c) FPG enzyme. Concentrations are expressed as pg L'V The values are means ± SEM (n = 3). * indicates a significant difference from controls at, or above, the 95 % confidence level.

181

(a) 50

40

1 30

20

10 d5 Rdl Rd5

(b)

Q

I-

80

60

40

20

Control

d5 Rdl Rd5

Control 100 180 320

^ 30

Rd5

Fig. 5.21 Repair of single strand DNA breaks (a), oxidised purines (b) and oxidised pyrimidines (c) by Mytilus edulis haemocytes in vivo following 1 d (Rd1) and 5 d (Rd5) recovery period after a 5 d exposure to cadnfiium (dS). Concentrations are expressed as pg L'\ The values are means ± SEM (n = 3).

182

(a) (b)

R =-0.604 P= 0.038

E >« N 0 10 20 30 40 50

Comet (% tail DNA)

m<i. 0.9 Q.

R = 0.057 P=0.86

15 30 45 60

Comet (% tail DNA)

I CQ 'S

^ E

?-E

1.5

1.4

1.3

1.2

1.1

1

• R = 0.034 • P=0.915:

•

"l . . . . 1 . . . . t . . . . 1 . . . . r

10 15 20 25 30

Comet (% tail DNA)

Fig. 5.22 Linear regression analysis illustrating correlations for immunotoxic (phagocytosis assay) and genotoxic responses (nrwdified

Comet assay buffer enzyme control treatment) in Myfr/us edufe following (a) 5d exposure to cadmium, (b) 1 d recovery and (c) 5d

recovery. The solid line is a linear regression and the dashed lines represent 95 % confidence limits.

00 4: o

E

R =-0.696 P= 0.012

Si

9> 1.5

0 20 40 60 80

Comet (% tail DNA)

( O S P= 0.535 J

(c)

0 20 40 60 80

Comet (% tail DNA)

Q l

E

03

(/) 0) o (0 "X Q . O

m £ (D —" (0 O

1.5

1.4

1.3

1.2

1.1

1

•

T • • • • I • • • • 1 • • • • t • • • ' 1J

R = 0.029 P= 0.929 ^

•

•

' l . . . . 1 . . . . 1 . . . . 1 . . . . 1 • • • • 1 X

15 20 25 30 35 40 45

Comet (% tail DNA)

Fig. 5.23 Linear regression analysis illustrating correlations for immunotoxic (phagocytosis assay) and genotoxic responses (modified

Comet assay Endo III enzyme treatment) in Mytilus edulis following (a) 5 d exposure to cadmium, (b) 1 d recovery and (c) 5 d recovery.


CO

R = -0.704 P= 0.011

Q ) If)

0 20 40 60 80

Comet (% tail DNA)

(b)

It M Q .

03 (/) o E N

2 [ ' ' •—• ' ' ' • —

R = 0142 P= 0.660

15 . •

1

•

•

•

0 5 "i . . . . 1 . . . .

• * . . . . t . • • • r

15 25 35 45 55

Comet (% tail DNA)

0)

t-M a 1.3

• I I I s ' x 1.1 [

1 i o E N

R = 0.029 P= 0.929

10 20 30 40

Comet (% tail DNA)

Fig. 5.24 Linear regression analysis illustrating correlations for imnounotoxic (phagocytosis assay) and genotoxic responses (modified

Comet assay FPG enzyme treatment) in Wyft/us edu//s following (a) 5 d exposure to cadmium, (b) 1 d recovery and (c) 5 d recovery. The

solid line is a linear regression and the dashed lines represent 95 % confidence limits.

5.4 D iscuss ion

The suite of biomarkers used in this study clearly illustrated that the apparent

toxicity of the heavy metal cadmium varied with the level of biological

organisation Investigated. The degree of bioaccumulatlon of Cd in the mussels

when compared to the apparent lack thereof in the sea stars posed some

interesting questions.

The chemistry analysis of Cd levels in the test vessels at the end of the 5 d

exposure, coupled with the cadmium burden present in the different body

compartments of the two species yielded very interesting results. In the sea star

exposure vessels only a very nominal loss of Cd from the exposure water over

the 5 d period was measured. This was in stark contrast to the mussel vessels -

where there was on average mean loss of 48 % of Cd across all the exposure

vessels. Unsurprisingly, the amount of Cd accumulated in the tissues of the

mussels was substantially greater than that measured in the sea stars (a factor

of fifteen times higher). Control animals of both species showed very low Cd

concentrations in all the body compartments analysed. The gut (digestive gland)

of the mussels was found to have accumulated more Cd than the gills. This was

as expected because the gills are a more transient organ for the detection of the

accumulation of xenobiotics. were as the accumulation in the gut can indicate

longer term exposure. For the sea stars the tubed feet showed the least amount

of Cd uptake. This is to be expected as the tube feel were analysed as a

surrogate for the gill tissue analysed in the mussels, as the tube feet are the

primary site of oxygen uptake in asteroids (Lawrence. 1987) and are therefore a

very transient tissue for the detection of the bioaccumulation of xenobiotics. On

average across all the concentrations the gut and pyloric caeca showed a

186

similar amount of metal uptake in the sea stars. The pyloric caeca of A. rubens

has been shown to be a indicator of short-term (days to weeks) metal exposure,

with metals been rapidly taken up and then lost (Temara et al., 1998).

Following laboratory exposure to 50 pg L' Cd for sixteen weeks, Cd

accumulation was highest in the pyloric caeca compartment of A. rubens, then

the body wall, the stomach and least in the gonads with the highest

concentration being 10 pg g"' Cd in the pyloric caeca (den Besten et al., 1989).

The authors noted that compared to other marine invertebrates, sea stars

accumulated relatively low levels of Cd. The pyloric caeca of asteroids is

thought to be Involved in the digestion of food which passes to it from the gut,

the absorption of the products of digestion and the storage of nutrient reserves

(Anderson, 1953). This would explain why the gut and pyloric caeca showed

similar degrees of Cd bioaccumulatlon in this study. One possible explanation

for the apparent lack of uptake of Cd by the sea stars from the water column is

that the metal was taken up prior to the 5 d sampling period, yet was released

by the body compartments back into the water column. But, the initial

(undetected) Cd uptake was sufficient to adversely affect both the

immunocompetence and genotoxicity responses assessed in the sea stars.

The behavioural (physiological) endpoints studied for both species. RT

(A. rubens) and CR (M. edulis), did not yield any statistically significant results

following five day exposure to Cd, There Is a paucity of data on the effects of

heavy metals, especially Cd, on biological endpoints operating at the whole

organism level in aquatic invertebrates. Righting time of the cushion star

Asterina gibbosa was investigated following exposure to Cd (0.5, 1.0 and

5.0 pg L'^) for 2 and 7 d. The response was impended (that is RTs increased) at

187

all exposure concentrations following 2 d duration, but this trend was not

observed following the 7 d exposure period (Bowett, 2002). Sublethal

concentrations of sediment-bound zinc (Zn - a heavy metal, with similar

properties to those of Cd) was found to impede the burrowing response of the

brittle star Ophiophragmus filograneus (Clements et al., 1998). The brittle stars

were exposed to high levels of Zn in sediment (119 mg kg'^ dry sediment) for

21 d and their behaviour recorded at least every 48 h. As the duration of the

exposure increased the animals exhibited a suppressed tendency to burrow and

to adopt a natural feeding position in the sediment (i.e. central disc submerged

in the sediment with all five arms held prominently in the water column). The

burrowing behaviour of an isopod from the Baltic Sea has been investigated

after exposure to sediments spiked with Cd (Pynnonen, 1996). The natural

burrowing behaviour of Saduria entomon (a brackish water benthic scavenger)

was impaired when the isopod was presented with sediments containing

35 pg g"^ Cd. But, these findings could have been due to the detection and

avoidance of the sediment-bound Cd by the animal, rather than being as a

result of the metal impeding one the complex components of the physiological

instinctive burrowing response. The lack of any evident behavioural

(physiological) responses In this present study may have been due to the

exposure concentrations and duration used in the experimental design.

The lack of a cytotoxic effect in either the sea stars or mussels suggests that

both organisms were able to 'cope' with the metal by accumulating and

detoxifying the Cd in their lysosomes, therefore no deleterious effect was

observed for the cellular membrane stability of either the coelomocytes or

haemocytes. Cadmium was found to exert a cytotoxic effect on the lysosomes

188

of the asteroid Asterina gibbosa at concentrations as low as 0.5 [jg L' for 2 d

(Bowelt, 2002). The NNR microtitre plate technique when used as an indicator

of cell viability showed a LC50 of 400 pM (45 mg L'^) Cd following 24 h in vitro

exposure of M. galloprovincialis haemocytes (Olabarrieta et al., 2001).

Cadmium chloride at in vitro concentrations of 50 (5.62 mg L*^) for 1 h

reduced the amount of NR dye taken-up by the mussel haemocytes (Dailianis,

2009). The NR uptake of M. edulis haemocytes was found to be elevated

following a 7 d exposure to 400 pg L' Cd. indicating that there was actually an

increase in either the number of lysosomes and/ lysosomal compartment size in

the haemocytes (Coles et al., 1995).

The immunotoxic effect of the Cd, as detected by the phagocytosis assay, in

both species may be as a result of the lysosomes of the animals accumulating

the metal to detoxify it. Although both species had reduced immunocompetence

following 5 d exposure to the metal, the mussels returned to control levels after

just 1 d recovery in clean seawater, were as the sea stars required five days

recovery to return to control levels of immunocompetence. This indicates that

the mussels had a more efficient mechanism than the sea stars to reverse the

cadmium-induced immunotoxicity. Cadmium at concentrations of 40 and

400 pg L' had no statistically significant effect on the phagocytotic ability of

M. edulis haemocytes following in vivo exposure (Coles et al.. 1995). The NNR

microtitre plate technique when used as an indicator of cell viability showed a

LC50 of 400 pM Cd following 1 h. In vitro exposure of M. galloprovincialis

haemocytes to Cd at concentrations up to 2000 pM for 1 h stimulated the

phagocytotic of the cells in a non dose-dependent manner {Olabarrieta et al.,

2001). When A. rubens were fed Cd contaminated mussels the Immune cells in

189

the coelomic fluid showed a reduced phagocylotic ability and there was an

increase In the production of ROS (Coteur et al., 2005b). The activity of

lysozyme (an enzyme involved in immune response to a bacterial threat) has

been quantified in the epidermal mucus of the spiny sea star Marthasterias

glacialis following in vivo exposure to 5 mg L' of Zn (a heavy metal, with similar

properties to those of Cd) (Stabili and Pagliara, 2009). After a 2d exposure the

lysozyme-like activity of the mucus was inhibited, evident by the fact that the

reduced capacity mucus aliquots to clear (lyse) areas of the bacteria

Micrococcus lysodeikticus from a Petri dish. Once the sea stars had been

subjected to a 2 d recovery period (in clean seawater) this immune response

returned to control levels. Manganese (Mn), a heavy metal similar to Cd,

induced a suppressed immunotoxic response in A. rubens following a 5 6 in

vivo exposure al 15 mg L' (Oweson el al., 2008).

In the case of A. rubens, the modified Comet assay results indicated that the

sea stars were able to repair both DNA SSB (buffer control results) and

Cd-induced oxidative damage (Endo III and FPG enzyme treatments) after just

1 d recovery. The trend observed with the mussels was some what less clear.

The buffer control data showed a lag between the 5 d exposure period and the

increase in DNA SSB as indicated by percentage tail DNA. A significant

increase in genotoxic effect was not observed following the 5 d Cd exposure,

but was measured after I d recovery for all concentrations and this remained

elevated for the top two concentrations (180 and 320 pg L'^) on the 5 d

sampling period. The oxidative damage was not repaired (i.e. had not returned

to control levels) until the 5 d recovery sampling day. These results indicate that

there is a difference in the relative capabilities of these two phylogenetically

190

different species in their ability to cope with, and repair. DNA SSBs and

oxidative stress.

Oxidative stress is caused by an increase in production of ROS to such an

extent that the organism's antioxidant defence mechanisms are overwhelmed.

Reactive oxygen species are produced during normal metabolic processes

involving oxygen (Azqueta et al., 2009), but their production may increased as a

result of the presence of a toxicant. An organism's main way of coping with

oxidative DNA damage is through a repair pathway termed the base excision

repair (BER) pathway (Nilsen and Krokan, 2001). The BER pathway is initiated

by DNA glycosylases, which recognise and remove damaged bases (Hartwig,

1998). Once the damaged base is removed the gap is repaired by the addition

of a new base and the DNA helical structure is restored by DNA ligase (Azqueta

et al., 2009). A second repair pathway the nucleotide excision repair (NER)

pathway repairs much larger DNA lesions consisting of about 25 to 30 bases

(Azqueta et al., 2009). This process can involve in the region of 15 to 20

different enzymes and proteins in mammalian systems (Hartwig, 1998).

There Is only a very small body of previous work which has Investigated the

Cd-induction of DNA SSBs in marine bivalves and none In echlnoderms. Single

strand DNA breaks (as measured by the alkaline elutlon of DNA) were

significantly Increased following a 5 6 in vivo exposure of M. galloprovincialis to

184 pg L' Cd (Bolognesi et al., 1999). But, overall the authors concluded that

Cd was only a weakly genotoxic agent. A second study which employed the

modified Comet assay (FPG enzyme only) on M. edulis gill cells following a

10 d exposure to 10 and 200 pg L'\ found an increase in DNA SSBs at the

191

highest exposure concentration (Emmanouil et al., 2007). But, no increase in

FPG sensitive sites were measured, despite the oxidative potential of the metal

much to the surprise of the authors, given the oxidative potential of Cd. The

authors hypothesised that the increase in SSBs was caused by the inhibition of

the BER pathway (Emmanouil et al., 2007). When the results of this present

study are taken into account it is appropriate to hypothesise that the

echinoderms {A. rubens) are more tolerant than bivalves {M. edulis) to

genotoxic properties of Cd, whilst also possessing a better capacity to repair

oxidative damage. This could be due to the BER pathway in the bivalves being

more sensitive than that present in the echinoderms. Therefore the ability of the

mussels to repair the oxidative stress was suppressed when compared to that

of the sea stars - which was evident in the sea stars recovering from the

genotoxic and oxidative stress effects of the Cd after only 1 d in clean seawaler,

were as the mussels had not fully recovered following 5 d in clean sea water.

To conclude, Cd exposure did not induce behavioural (physiological) or a

cytotoxic responses in either the sea stars or mussels. But, immunoloxic,

oxidative stress and genotoxic effects were observed. The sea stars recovered

from the oxidative stress and genotoxic effects of the metal at a faster rate than

the mussels - indicating that the echinoderms possessed a more efficient

mechanism to cope with, and repair, this assault than the molluscs.

192

CHAPTER 6

General discussion

193

6. General d iscuss ion

The overall objective of these studies was to add the small volume of scientific

literature that has utilised the adult stages of echinoderms in toxicological

studies. Although a cross-section of echinoderms have been utilised in

biomoniloring studies, their use, to date, in laboratory-based ecotoxicological

studies have not been very comprehensive. Through the implementation of a

range of different non-destructive biomarker techniques on a selection of

echinoderm species following exposure to representative marine pollutants It

has been possible to fulfil this objective. By the concurrent use of a more

commonly used ecotoxicological test species, namely the blue mussel, Mytilus

edulis it has been possible to elucidate the experimental results in terms of the

relative sensitivity of the echinoderms to bivalve molluscs. As discussed in the

relevant section of the present study, generally the echinoderm test species

were found to be more sensitive to exposures to a cross-section of aquatic

xenobiotics.

One possible reason for the greater sensitivity of the echinoderms compared to

the molluscs could be considerable difference in surface area of the organisms.

For instance, when one compares the surface area of an adult sea star used to

that of an adult blue mussel, it is readily obvious that the echinoderm has a

much greater surface area than the mollusc. So. despite the fact that both

organisms have an open vascular system, and are therefore highly susceptible

to any aqueous contaminants, the sea stars are more likely to take-up more of

the contaminant than the mussels.

194

Respiration occurs over the total surface area of the sea star through

microscopic papulae (dermal branchiae) (Cobb, 1978) which are present across

the outer dermis of the animal and also through the tube feel (Smith, 1937). In

the mussels respiration only occurs in the gills (Bayne, 1976). Possible routes of

uptake would include absorption across the 'skin' or gills and by ingestion of the

surrounding water. As food was withheld during all exposures (in line with many

short-term regulatory aquatic toxicology studies), ingestion of 'contaminated'

exposure water can be excluded as a potential route of contaminant uptake.

A second reason which may account for the greater sensitivity of the

echinoderm species may be their more 'evolutionary advanced' capability to

metabolise xenobiotics. As mentioned previously, the echinoderms are more

closely related to the chordates (vertebrates) than the molluscs - and for that

matter, all of the other more frequently used aquatic invertebrate test species

used in toxicology (for example, the annelids and crustaceans). Therefore,

when the echinoderms were exposed to pollutants that required metabolic

activation (i.e. the pharmaceutical, cyclophosphamide and the polycycllc

aromatic hydrocarbon, benzo(a)pyrene) they would convert these compounds

into their toxic 'daughter' products at a faster rate than the mussels. As a result,

these metabolic by-products would result in deleterious consequences which

were much more evident In the echinoderms compared to the mussels -

despite the fact that the concurrent exposures were at the same concentrations

and duration for all the species examined.

It is important to keep In mind that whilst discussing these points of relative

sensitivity of the two phyla investigated, the only 'target' cells Investigated In the

195

present studies were the coelomocytes and haemocyles. That is, no other

potential target cells/ tissues that may have been adversely affected by the

contaminants were examined, for example the digestive gland or gonads. Given

the open vascular system of the organisms. It could be assumed that other

target cells/ tissues could be affected (either directly or Indirectly) by xenoblotics

leading to changes at higher levels of biological organisation. The use of

coelomocytes and haemocytes could be considered analogous to the situation

In mammalian toxicology, or in the human health arena where the use of

surrogate cells (e.g. lymphocytes or bone marrow cells) are widely used to

evaluate the toxic potential of environmental contaminants which could lead to

adverse health consequences at the whole organism or individual level. So, it

would be Interesting In future work to select other target cells/ tissues, in

addition to coelomocytes and haemocytes, to see what effect. If any. Is

observed.

The complete mapping of the sea urchin genome (the California purple sea

urchin, Strongylocentrotus purpuratus) was completed and published in 2006 by

The Sea Urchin Genome Project. Other commonly toxicity test organisms which

have had their complete genomes mapped include, the common house mouse

(Mus musculus) (Chinwalla et al., 2002). the sea squirt (Ciona intesiinalis)

(Dehal et al., 2002), the common fruit fly (Drosophila melanogaster) (Adams et

al., 2000) and the nematode worm Caenorhabditis elegans (Ainscough et al. ,

1998). But. the sea urchin genome was the first full genome available for a

nonchordate deuterostome (Davidson, 2006) and It opened the door for the

possible design of very specific toxicology assays using the sea urchin.

196

The 23,300 genes that made-up the complete sea urchin genome included

many genes that were thought only vertebrate innovations' and known only

outside of the deuterostomes (Sodergren et al., 2006). As both humans and

urchins have a shared ancestral evolutionary pathway, the sea urchin genome

contains a number of genes which are orthologs (genes present in two different

species that are similar to each other because both species originated from a

common ancestor) of human disease genes (Sodergren et al., 2006). A

cross-section of thirteen of these one hundred ortholog genes and their related

human diseases are shown in Table 6 .1 . The surprisingly high number of

urchin ortholog genes further reinforces the phylogenetic closeness of the

echinoderms to the chordates that other researchers have found (Smith and

Davidson, 1992; Littlewood et al., 1997; Littlewood el al., 1998; Adout le el al.,

2000).

The presence of these orlhologs for human diseases further reinforces the idea

that the use of echinoderms in aquatic toxicological studies could be of great

value, as it would be possible, for example, to screen pharmaceuticals and their

by-products for potential adverse side-effects using adults sea urchins in the

place of higher vertebrates (i.e. fish) in regulatory toxicity testing.

197

Table 6.1 Human disease genes and their sea urchin ortholog genes (adapted from Sodergren et al, 2006).

Human disease name Affected organ/ tissue/ cell

Human gene Sea urchin ortholog

Acute lymphoblastic leukaemia Blood Tec Sp-Tec

Alzheimer disease Brain Amyloid precursor protein

Sp-APP

Parkinson disease (type V) Brain UCHL1 Sp-UCHL

Cancer association Breast Ret finger protein 2 Sp-BRCA1

Deafness Ear ATP2B2 Sp-PMCA

Obesity Endocrine system PCSKI1 Sp-furin

Glaucoma Eye Myocilin Sp-Amassin

Respiratory distress syndrome Lung ABCA3 SP-ABCA3A

Huntington's disease Nerves Huntington (HD) Sp-Huntingtin

Spinal muscular atrophy Nerves, muscle Survival of motor neurone

Sp-SMNI

Diabetes and hyperprolnsulinemia Pancreas CPE Sp-carboxy-peptldase E-like

Male infertility Sperm ATP2B4 Sp-PMCA

Thyroditis Thyroid gland Thyroid hormone receptor Interactor 12

Sp-TRIP12

00

The different xenobiotics used in these studies were selected to be

representative of a wide-range of marine pollutants. But, all of these

laboratory-based experiments only examined the effects of a single contaminant

concurrently. While this approach was Invaluable In determining the suitably of

adult echinoderms as aquatic toxicology test species, it has to be remembered

that the organisms would never be exposed to single toxicants in the field.

Therefore, it would be very interesting in future work to examine the effects of

mixtures of pollutants on these test species. These mixtures could possibly take

the form of potycyclic aromatic hydrocarbon mixtures (for instance the water

accommodated fraction of crude oils) or effluent waste waters from sewage

treatment works.

199

Appendix I

Formulation of experimental reagents and solutions

200

Reagents for neutral red and phagocytosis a s s a y s

Neutral red (NRR) solution 0.004 g of neutral red powder (Sigma Chemical Co., Poole, Dorset, UK) in 100 mL of distilled water. Stored in an amber bottle at 4 °C.

Bakers Formal Calcium (BFC) 2 g NaCI 1 g C4H6Ca04 4 mL formaldehyde Made up to 1 L with distilled water. Stored at room temperature (20 ± 3 °C).

Acidified ethanol 1 mL glacial acetic acid 50 mL Ethanol

Made up to 100 mL with distilled water. Stored at room temperature (20 ± 3 °C).

Phosphate buffered saline (PBS) 1 phosphate buffered saline tablet (Sigma P-4417) dissolved in 200 mL distilled water. Stored at 4°C. Protein standards Protein standards (Sigma P-0834) with stock concentration of 2 mg mL"' diluted with the addition of distilled water to 0, 0.2, 0.6, 1.0, 1.4 and 2.0 mg mL"V Stored a t 4 ° C .

Reagents for Comet and modified Comet a s s a y

Preparation of Kenny's salt solution for low melting point (LMP) agarose For 1 L: NaCl - 23.37 g K C I - 0 . 6 7 0 9 g K2HPO4 - 0.12194 g NaHCOa- 0.16802 g Made up to 1 L with distilled water

LMP agarose 1.0 % solution - 10 mg LMP agarose/1 mL Kenny's salt solution.

Preparation of Tr is-Acetate-EDTA (TAE) Solution for normal melting point (NMP) agarose For 1 L: Tris-Acetate - 7.248 g EDTA - 2 mL 0.5 M solution Made up to 1 L with distilled water

NMP agarose 1.0 % solution - 1 0 mg NMP agarose/1 mL TAE solution.

201

L y s i s Solution For 1 L: NaCI - 146.4 g N a 2 E D T A - 3 7 . 2 g Tris Base - 1.2 g pH adjusted to 10 by addition of NaOH N-Lauroyl-sarcosine - l O g Distilled water added to give a finally volume of 890 mL - as solutions added prior to use increase final volume to 1 L. Stored at 4 °C. Immediately prior to use add: TRITON X-100 - 1 % (e.g. 0.5 mL per 50 mL) DMSO - 10 % (e.g. 5 mL per 45 mL)

Electrophoresis Buffer NaOH - 40.0 g Made up to 1 L with distilled water. Stored at room temperature (20 ± 3 °C). 200 mM EDTA solution - 0.5M EDTA stock solution diluted 2.5 times (e.g. I m L EDTA and 1.5 mL distilled water). Stored at room temperature (20 ± 3 °C).

Electrophoresis buffer prepared immediately prior to use. For 1 L electrophoresis buffer; NaOH solution - 300 mL EDTA solution - 5 mL Distilled water - 695 mL

Neutralisation Buffer For 1 L: TRIS - 48.44 g pH adjusted to 7.5 by addition of concentrated HCI. Stored at room temperature (20 ± 3°C) .

Enzyme reaction buffer (for the modified Comet assay ) 40 mM HEPES 0.1 M KCI 0.5 mM EDTA 0.2 mg mL"^ BSA

pH adjusted to 8.0 by addition of KOH. Stored frozen (at -20 °C).

Ethidium bromide staining solution Preparation of stock ethidium bromide solution from original 10 mg mL'^ solution: Stock solution Is made up of 2.0 mg mL'^ ethidium bromide in distilled water (giving a x5 dilution) in an amber vial. Stored at room temperature, in the dark (20 ± 3 °C). Working solution made by adding 0.05 mL of 2.0 mg mL"^ stock solution to 4.95 mL distilled water (giving a xlOO dilution of the stock solution) In an amber vial. Stored at room temperature. In the dark (20 ± 3 °C).

202

A p p e n d i x I I

S c i e n t i f i c c o n t r i b u t i o n

203

Ccn:»o:s lists avaHabi£ «t ScierceOireci

Aquatic Toxicology

j i M ^ i n t l h o m t p a g e : www.o l t&v le i comMocatB/aqut tOx

Linking genotoxic responses w i t h cytotoxic and behavioural or physiological consequences: Differential sensitivity of echinoderms ( A > K Y m y n/i)o/]>) and marine molluscs {Myulus •. Wii/is) Martin N. Canny . Thomas H. Muichinson**. Rebecca J. Brov;n^ Malcolm D. lones-". Awadhesh N.Jha-' •Vi,sf»i.»i;irL-.i: h-r. tvîi^^.^ tpvr' i j ' i '«:v-t ff.c• w n . w . t ^ : I Jt^TCJc-n

A R t I V L E I N i A B 5 r R A r

InirpJicd Jiwrj ior / r.Ldicj jcdrrtsrt! nulnplc btoirurVcr trstxnw^ in ihr sf j r.u As-.mii r j i fn ; j - r ihc WjT muiwl Mi iutii ftfL-ti cipovrd J urge of rorrrnL-jaons of dir.-î in i indirfcr jr:in£ p ' f i u t o i n i : mrlhjH tranhj-ic sjUbruir 'M\1S' j n d n'^fcptoipturciz^ iCT* an rT."iiunTicn'. j l ly ' C r t j r i onii-cjnior^hjrmjrûiif J l l . mprniwh',m ordci lodcicmirf if i:waip7*ss«Jf>.'':ir:rTiri:y nj~ nk-c«;rifiTr:i j : ihch rhirrlc-.Tiiorbrzlccic-j! j r g a n i v a i D a Trx-t-xr-rTi.Tirnu. driij-ri jirr^rd i ; r:»nrurrs'-lly r . - j ujlf b«.r:nirtrnclbchj-nQ-jijljndphi-«K)lr-g:f Jlrindriirnj 'nRhtns:mc"jr.tl'rIcj:j.imji« l;i

, Mr«. j rdmJ^vh . :ii'ir*'''^r.T!y ir. j d i r j c T t-rvintrjtrriiv r>i"urjl rrd rrurr!ii.-.T j ? u y . nduiT ••n< ' rwritTnjrlci. Mn. j r j W»4 iirand brejics: J i drTprnunrd by thr i'>nnn j i t j j • Thr pîrôl j iw i r ' uilcd itv drirrminarirtnof thr Tuwoum i n l r r j i r d CDnftnrrjticn' MTC . jtizt la g:?acxjiccvjlu.m.-p.Thf d MTC. Ji tJrirm ncdb>' ihclur.nvjloiihp wsarjims,nhowrtl v j « j n t o U ' mcir tCTi^r.vnhJr.mufM K ir« MMS ,15 J n d 33m£L triponnrly J i d CPiifi jnd iBOfr.jL ..-rspi-rnvriy • F iI tvih ^rvr.f* j r , J ctwminb.r.ioioxics^'wii not foun^ s tf stsnifKanity dutrri-ni crnvirrd iac::r.ir:li. Apî IY IFT- I .V MM^ pxpcnrf lo m s u n • w.Mc.n sho'Aed 1(KK rHnj1:T>- j i ^ . s ^ r rorwm'.rjticn! j l i r : hi nfiznAf^', CWJI daso rrsponsf rcljlismhi;?! wcir oi;icr.-ed for bc4t srrrtaxtniy rKlFKxr.tA n ojrh spcors l-:>l-k 'AT i? i^pcirrc :n CP. c'-irf C5rrPli[i:r.> wrr.* jiso frmnd bnwrc-. ihr N-rj^nrtitJi and pf.v*'N-yTjl rcii.pwws i - r cc.w:r djrrucr ir.rjcn s p t x i c t l u r i - M N B l ' K -0 .? j Ccrzn ST. t - O ' j ; nwiisrK MS V3 C!" R-0G9; CDrrr: -.-i. CR: K-0.73, Thi* intrjrjtcd jppfcutt.. jpph-.ng r j jn- invj! w j \s jv^ :oi:niuliJWoJt;'/tl.'ifrmir.pir* rt»pc>rws JI diff-'irn ir.rlôfbicki^'J. Drj;jriuuLr..indo'jU*s ihi.' (> L'ni iJl v j . u c c f bfhj ' . . i ; / j i j» J . I . ; p l -v^nk^j j rri-Jw-rpN n d.-H'rTnxir.fi r r r iMiimy J >"r„-Tt; j l -i -nur.ncfKf jrisTO jnd r.ijhfcjni^ jl«?rnrn*h'«inrr nf i r * : lu: i rg Jdul i o r i i : i j i c r m M r . crvir-.r.iTKrijr

1. inirodudion

7 -c «: :curfc iKc c( gcr f--!luur-.i in &c I ^ J J U : crtv.. r>.TirA-n: i< of i r c r c j i i n i C( nccrr. •'.V.ii'',.v jnd Krjmcr^. I'JCO' L l j : * t T cl J 1 , i9 ' jK.; ' - j JihiÎ C'ln^csjiK-n:!-,- in rcfcr: \-c-'ri,man>' d i ' i ro r i cma l l v rcj;.;iu- J i S J y i hx.v hcor. J,;vTl-^pcd and siudicr hive been c jrr icc r- : u- invc«iij:jlc ilic-ol'ir^:* M porcntisi j:i'n.i-roxic rofT.poundi J i u j t r jrjran.^rr" y^srg s V3:.civ c f h K i b r i c A l r,:ii>.mcior biarr.-rkc;! D.tc-r c: ; | . I W . J . J - J . ^ C C S ' . These lu^-ic '*u jcrwMiccr CfHj . j - . ' r tponjvi rn..>::;• t^rneJ • ui i-i (orriacc ictl i are used j% ar inJ;caTo- - Ict i t r p j c t c - i h c Z>ir.v;nu:i

i A A: - It.

^lirjS; Nrvrnrwicis, there t2\ keen, hmilec tV-'-i fkr. thi l,nVjj^-* hci^ncrn p^nc-'xic re^p-'ir^cs aivj rc^r-'fw^ i ' ^l]:^cr level* r j ' bisIciTi.-j' o i s i n i u t k r . n i-w of ihc key ^ bjc.".ivcs i f cc^ i.wrcMcgi.-Jl studies..*ii:drru nandWi lJ , l f»4 Jhj 2C(:«; .'.KfS.

Tratlil cnjily, iScbtuc n u i i r l MVHiU.'i'dî'li h j i b c r r j s c - l j - i scniinrl ipi.'cici in miny m-iru-.unnB prt prjmmc's jnd r r jc i ;ch p-i >• iM.i lsr.VidJj 'A-nel J L . Jl'O; ('"CvnnuL '.iJC«n: 3 w n ci 20ii2L hcwcvrr. '.here [s a gtjw.ns ccnccrn thai b i j : i s cnviionrr.cnia. hcjith A S i C i m e n i s cn c w jpcow rr.iy r.,it be !_f!icifrily :i b_sT: the mj in [:rt4j."cm be.n^ that : - r tcni b-Omar^cr j iuJics ir^ rr.tz.-SLicC m j l i n i i e d number of species. anJ then roporse? n u \ n n l-c ir.dK-atiwc -li oihct spccici ir. the mmun.iv iCbcung er J I . 200G. Jm. 2004, 200S\ New rc jul j i icns in ibe E-i tpcan Jr.i unJcr ;ne rtegisiraiKn, (Iva uJiKir, Autb'-nsj'.tjr. jr.d kci ini i ;oncf Chcn:i.-J i .RC^L!r1:Et i lK. '? j'XHilproi;; jr-.r-i^hj-.'efcrv'griisc^'.K•^ weaVnf« jf:d hjtve zca tnnwndcti ihai :.• fîrcvT marin*- ••'E-^**-

204

V> --rev

K ti:.:>* [cMinB vv:h i icfhwjicr fish aui.-j-'wn': i r d ihr. Jî be suf^plemenifdvMir. mbltuscj ancochiii^ Ciiir.t InJceJ jhc rrl^

im-crcbraic rrii jpccifi is vônh-y of further «-'LV.»;c.rli jjitj: imd: ICS d u t r c their .::osc p<sM:cn lo:hc :h i . r f i a ic * : r ! h c dcuierrtir-rnr line i.f Jevrlopnicr.! t.f tho animal li;i^d.:r:-~.Wouiir o: J L I'tiOO v.hkh t:ciijld hrlp in jv-jijing iinncct.'njrv utPi;rvcr:cb:o:cs i rii' lish.-.fx loxicobsf^jl siudirs far ein:cji i ca j in^ Rc>'j:tvt:'.v Is. l x i w c \ T r kncvvn Aatx ihc scr.sf.tviiyA f t:njn.'ce:m'- v r w . l a m m u l c r K i t j m i r u n n . • l l u i c h m K : ; : . 20i.»?i Kf.-cn:iy p-i^'iM'-cd v.c:lt cchinodcrms in.-ludrt ihci - f fcci i ^:rrxii\i ir,d v:^^t-rt •J-jtupting compcL'nda lO-A-csan c: al.- HiAî KtcF<c « jl_ 20i<': fusr i « 2007. ÔOEl scnoicAi: ciTivii iTaSar i-: j , . ? 0 M : r\if .-hrCTic P4jb srd hfai-5h>:L p r V i i r , . n u r - i n 'Cii'-MÛJc-, ft al- : :0X: Plmmoft jl_20tUJi. ir. ir.i« iiuJ>. thr rtTTimor if a iiof. ^s:^^tta rutsra WJS ictcctcd J i".?jn:[rr, J5 K hiz J w-idr d i t i n b u t i c n i h r >,;:shout thr.AiIanci.- j rHiArcr iCc;)^^:^! w j : c n 31 EurapciSaramdiava. 2UaiJ. Alth;wjfih i h i i cch:n^.icim has bfcn uiilisrc i r biafr ioni ionng itadirs tlvmara TI X . , U j n i s CI aL 20C-;i. it Kas no[ been used ocicmvcSv i r c.-Qiatt;cl.

tiTrcii of CD'niaminanisbn iHc bchjv:^jai Cf g - t i ^ ' ^ i av;«d-3n:e. rcproducm-c a^d w n l beKavicuri cf d.iTfreri nsh »pr.-'c* have b e e r studied extensively 'fol r c w c w see SeJ!I j r d 5lLn:jn 20(M1 Because bchx.'iour, j phcnc-ivnic ni.jJifi;j:ipn. li r.jfr.r..'.:«i by the genctkr n u k e - u p nf the irjin.nd,:al. and trrJU»c r, f j J i * phyiiclogiHlfunctHir vvtlh e<o(t.gii:jl prt;ri:!M-j. il s e n c ! - u i e -f j l indicator of loxiriry J ; the whclc cfpar-isr^ i r * t ' H'.-»i.cvci. TH.'-jiierr.pt has been mace previously olinVcrmtcjraie :hc^cr:^:..'n;i rffecrs of conumir j r t * w th cfTccij at f4:y^.i:l.-'fu'al ber .a^- . iLrx lewis-, w h i c h wuuid lead ti^'dctiinirnijt irr.pjct cn vi: j | lii'c f.rit-. •esses inrludtng the reprsdurtn-r pi.:cr.i;3. J I inr -wginî.-n'. :hs main -ibjccinr cf e c o i c A i r a b g i c i t stuJu-v. In addiiicn. such fi;:d-les n:it only provide cppontmtvio elucidjiepv:f nit^. rr.nrun'.u-na

physio}DCi»t cr bcha\iou:j3 inodi6ja:i,i:K :ould a'ii.) f::rr.U' Ijtc the use of s u f h i l ( c i J l i 3 t u M-i:h(-ui usiop.T.-u-ve te . -nr .q-es wilh jr. aim to uir appropniic behxvisural measures I3 rr.ininiisc c».poiiii< •fcomaniir.uua t c l c v c l s th« nt.iintain iuiiainihili:y uf ihe e m i r D n m e n t 'S:(!ii and Slomin. 30CH:. Iri ihis cpmcxt. ';'cjr> j n r c ni:e"CCR>hjsbcen3^pd irccnrlyiv-irci nici 3 f > ' .:: e• : pf scope for Eroxviii'i in bivjive mollusri as A irnsftr . 'C in J v j r ^ r ol puiluiini exposure CWiddaws et JI . . I 'JDS Ca:«>- Vt al. '_'0O7;. SjjiJeii et al^ 3D0S> arid rishtlns tirne i JefineJ a'l the ::mc taker fcr i r j n i n u l p b c c d upside dowm rojetun :.i ni > rmal, upright tSjCitijn} hai bccnimplcnicnted a^ an hucalcr ct cyera.L'hcJitt, »f cch.ncdemi speficJ'lAxtjK an j Saîba. l^SI Davici el J L llWS: Icmjra ct a L VJ'JOl

In the hackdrcptf abrt -e inicrrr - i i i -Jr" ' ' _ - « ; i x n v o i i.^-re :cr-a t vslldaie ihe biomarkcr re^p:rfiscb t>drtctnimeiHe effc^rivCjl ^'en-^ loAi; agents u> rrurinc masAcIs -IIV! cJu-'U; an J K-J l u r rtf^^fJSJ. •'b CMTjpjre the relative jensiiMty the ati jT 'irni ri r M rJulS. and A. fhtk'mi for i f b i e n c d rop*. nscs ai dJTcient .evtlj i C biolssical tHganitaiton: and Icjiesi f j r cf.irffTjruai bc:*Ai'o^>;rcnc-lusiirejponscswilhhijherlcvdstlhi. 'Jtjt. ' j^reiparisrsiin.-^vding' bervn-cural and ph>-5ic!ogicj| respcnscs j i the inanidiia. I .- \TI. to ntidaie iheii usctutnevs as a potpnual n'.rt-inv J i i v c i- A tj.i iuxt.'t*; I»S cal iciting and pcKniar moniioiinj

In Crdcr lo achieve : k c ob|crtiv?t, we lelcciec J cire--t j . ' i -in^ reference serw'cxic jpcni. meihyl n»c:h2rj;iu!-'i naif l^•V.i i i n j ar. indtrm a.nng^l.c. icqu.Tinj; m r u h - J K a:tivji:r<n* J i ^ n t . .•>x:li:phcq>h3mM3c(CPiLN'.V.S W31 sclecKdai i: hasbctTT. ' - s c d p r r viiJiiSi> as a refcrcnrc fcrcisxjni . - man>- siudipv.ir:;.^dinn a u r i a r ;ir mammal ianfd l i lN j : j ra ,anc ia l . . i l>a ; . i \nd-e i j l i c i j 3 . tCrry zr.c in marine imwnebrateitjhaei a l . IIJUC, SiJJpa.fc: HzrX and fynwii ? O 0 3 : Mj^gcr ft j ' J ipo . ' a .

: y . - r:t^-;;:*'-j'^'^« tCT • J ..-ir.:. ihcrapiui k and immunc»_p-piTT* l ivv i f cr.! -^ is as a t»r_-» i- \ti-ir.t '"HaniTic': ci a l , IDP-I; R /c ;k i fi aL 1 0 0 ^ : Iv i r ei i" . i , T . Thi-^ cc.rri^_r.d ij iTsnsfQrr::ed by inir.:*.c!lu 3r t-nrymj* :» j ; ; L• j'.K>1ai,ni: mffiab«-liiP5 vi-hich ::L-iS - - iViVj i t}: D%A saiand'- *jiJ;i>J:n et J L iCrS^J. Upcn appiica-i.cn.'.hcar.f Jndl,êmcta^. l.:.-cjrercnaIlve>:crcicdiniiibMimial Fr^-pi.-.i .n. up I-- I0-2I^'» .ycirc cV .T i i cd unihar.ged (Andciion n a L BL;CIJC CI JI. . ^ W V . ,loKn%cn :ei' a!.. 20081. The a;inTr rrrr.pc.-rds anc the rretptii ;itc» A i l h pocr dep-jdabiluy and high prrws>cn:c haw beer, ce icrc i l -n watte and ;iurfarc waiexs j i l•o^.:i:n::J:;•J:^ Jpio lOOnc. ' J i i r c s L l i thtfrcfiastpwingfT-n-1. -c rv - . ' j l . - i r r c m brtS -.-L-iin -^^pc j t i 3 n - jman a.id ecosysicn; h :a tn ^:ect>:-(tjrcman'>r J . 'J'J7; B ' J C I J C e: aL3000: ;chn» jn ct a". 'ÎtOS, VVh:>CPh3i|-»*r- u%«i M jJcfwcTtcjcrffitauicafCJii

• it-'.' Xr\t\ i::7-'cri '-f c i T ' ' •i-^ assay* u rn;? emfa-y-i-Iarvas ;.f if.o -an=o.jr.rclid: r.V!*'.. '.n'- JumrnJir ;ha ct j l„ ia9C>.(iitle futiniv •iaMi-A.*; hi iftc;—pi,- >fih:s phinnai'euncjlonmanrt .•"Tjanis-ni-

2. MJi f rubandmetbotb

ftlkl-er.icalj andirar!:nrj*.\-c:co£a.-.aJ>ii.-aIc.-adcandchla.red ItJ-n Lrii*mical C' -^ r< •'.»•, IJ ir jseL I K . i:nlrts ua'ted a the : -v.:it.Tciicrcrn.JiJs,tPcthjlrT:?:h.»nctulffyia;c'V..W.C\5nnTnber V C 0 7 - i pwnivW'tlarJi^-.i p h .spharnidemonshydrjicICPjCAS num'rvc* COSr-I'j-2'. puMv-'.i7? M-trc procured fr&m iKc same s r u : c i \ * ( per ihc publt>rft; ,-t"..im.«non .c.c.i-''ir.5cn ct aL'200S: ht :p;'i J .Iks'-jfs .vl.^lh.^:c*•', EX in .\'.MS anc L T a:e readily cduble in v .a ic* - iT . i - i t : u*inFeij:afc M.VS 2."illgl, ':CPupu»SOgl ' ) .

Kiv'iiis i^-iL-mm ir.f tcn.-iht i.vere c^.'êclecCLnng Match 20001^. mfVti Quirt. C;uw.l!! ' . ( f l .snJ rerercr.cc. SWg72 a refcTcr . - rs .u- J^JÎowin^Vt'I-Citi, Ji rniry.wc;r :-|eanedcfeptbioar.t& and placed.mic *0l giai* ^l;.a:.a t> aeeljnaiise under the lab-i.'ai^n: ;';nd.tii:»ns until rr^-rrrd. They were held at I 5 i 2 C. 34 i I ptu. tn .l ^ lm ti!lc:t'i;,M.-a.vatcr. with a lifjhtidark c>'dc cf >h I j h'.inii-'nsi'.y 3Vjiv. . h i v r >-u^*\[hcyv.T.'e fed UvlcriweHv

c:J Kt'ir.iiK w;:h a wand.Kd a ; _ A n i . m invrfirbrjic tf^cii miMiufe •\ViTrii:V :r.ixry F-jcd"^. LîiJfrv Marine™. Rcn.Ridi™ and J i icJa l ja r A 'L6£Ww.'[rc..Lotiecdurin5\1a;chatidApnl20lMi fr->niPyini-::ibS3i:nd.l>ci-•- iJK-i:^d reference:SX4735]CXuS4ng iiJUOACivera j r J irjv.în3:.i'.-rv ia :ac i . five-armed, sea stars were iei;_T.t'dii inc'jb.iratcr. ' .Ai.s;i:r:sfi wc:ci.rOG- ' .Mmmindnni-c i r : . Fhry rtoie hf lJ in ". JI c j is i J i ;uaT:j eq-aippeJ w-iih air-driven liliv-:; iin:iT rc<;..:ed lunJ-.-r '"v i j m e ciwdu'cni as the m j s s t l r ' . Pner :t- - i c :hey Arre red niu'.jr.'i ad .'itlium,

Thc<êr—.'cr{atcex:;rni - noV JsiatiLCxp<->ure>issh«v\-nin Tig I IJroc^pi luif i . o a * - t«: ?he two species differed slightly: nETi in^- ifxvrs' • ix A, ru^;':? - A C T Cercrryncd fti lowrg 3. 5 and 7dc.«i.p WTc pcrciand-r; , - i-.vxTiipy a r j genou.Aiaiy wcri: drtc:-mined a:ir: ^ anc 7 d . e » T " a * i .fjr M. fdvhs 'clearance raic?

M.i:v ard cer3ti«ikily'.^cie d e : c r r r j r ^ fallcwiinj; 3. S and 'dexp-'ûrj- Ih r ^^^!S^T»•vcnl^'ft^^s w r e selected based o n pre-v;na-s -v v'r i-n ihe emhr>« ..J:-.,1V- i,( P. itunnrtlii artd Ai cdullS L'ha ri al. lis-j<j, 2C00i.by. k i i*jlK:a::^n pOrpau». iUnV and Jensen, ..'U(-3 jK. ,;fcJV\UiCt,r^cii:fJnorMnthtTar.Eciif I-33mgL ' i n mJi-r^v^J-_i•^! x'lJullL l.tLni-'.vihra;lirrit-dici Lir.embryo-ljrk-jc y'r Ut.Pi.Tli:: an J A) rtufs. i - c inc rcTul:i caiained hy Rark and )cricn _'v'«.a* usirg adult 1! r'-Jiii'^. :hr »ni;ccn;-aiinr range u'ied

205.

A / I d 7 J i l , > I ^ j h a n g lur te - J e i K i i T W ^ d ^

T . -

C-eaiaflce taw ammryr^^

assays carneo cu IComei

r^'.'iiirJ :Hr scni-U,giflift.T.ic sr j lc :cu:!rc]v,cicd tn c.'cmi-'irv tcvm^. This riHfc w i s ff(TJ.^'jiciJ fdf ihi' CP.e'jrKtr.UW^nt ir, lr,c; •M 'h. 'Kj ci'ii.OlKhJL "hpcxpoiuics ivewîrncdcui 4; 15 : 0.2 C iT'J t '*» Jnimali vvcfi- >,i fcJ duiins-hi* cNr-*j'^ff pcrttd. T^.c cVpi-tuic v«>ch. wiVri- jcr j icd ir j iviaujl lv usirjr'Pasicu,- pipcitrs ^ - J scaled with rj-'irrp-.'^' VesiclS d^vrc chrcicd daily mcr-(Jt.'.sor JtvJ ihr icTTijJcijruri* . sj:(riry./ai%*:4i-i:d V;xvgcp Ard pli

liiv v.'Jiei w e l l - i ) t i i - . . T . ^ j i iric i jh ic l i n u . In mriK-n viiih; rt-jTjtj^^ry matnrnîliin rcrKid«ir TV c i i ; c r )A , ' r^ i . \ tmur r . ' i Jc i3 i« i J j t r . ^ i \;r'nccn(rj:i'Ji sXlTi or MZC, i i n . v A ' t sprr * ' ; tal.cn

.' -J. .•Je.'rrmuianu'; oj npir^^ nmnt • ffr •

SlgKcina [tmrj. ;KTi wcfr dciesmrcd: j y di:vf;:b*.'d by uhcf w i t c i > ! c . 2 . T e r T U : i f ; 31 (Mb; p r i j r ;f^«bmc>c>ic lair^plr l<rcia.n; whidi were uifd :?r cyitifcAi.MV and £CfK>icxiatv JS»»r 'ritfRT vvttr dv'ivrrrj i i f Ji'i^drtidbjUfv in j l ss * . -q i i i r ia 'd imehs>jrA iOin»> i O c m s j-l-zmj :urij:-\ir,(; i^-^ f c k j r . j f j .v i ;ct , t.i:n ^ca

W i s in-.-cned u-J STJri pri »:<Tnccr':r3r.i.nl And p l j ; c J ccn-:»ahy s - i : h c biSt d rhc unic, Nii iJf ;l^.-fv"Fi*her?irir^*^"vvrrc Viitn prc\Tnt iny ji'«-:'..tcrttii>* fiTL: jmiiu:irni7jrin<-HJ. f'JvJt ^fif* rr-.c liken U!>• jn i -nal 10 :Vrv--tYarri i-ni.i r^ abir j l s j ic w j i rr.ciiurcd w::h ihi- imhiJaci j • f jU I'he Jrnn in ; v r i x ; wiin iKcSJicpf ihecant, Tr. ipirx-cdurewaNrcFCJtecihrcc timcs.'OOOf v \ i i rilvft; forej:fc j:v.'n:ri >T)fi :f-CTTKjnKT JcifrrTyncd;

=^01 iohjtni'«.ymph-«Tiplcc;(Tciir:n c l c j r j i K c : ' * ; c wa

S.'3ik-.: rt al- 20()a HnL'ilj: indi\'rj-al l iL«^tls i'J ir.-isicU per ti-rc'eiiii j i i jo" V.\'TV pT:n,od ;ri Vcpjrii^ -1lJJml,bcjVcTs rcn t jx i r^ 3;'Gr-.: •w^wite j / l i l ic jcC i. JUttT. .A v^^'^.g h j r i 12 mrr. • V-mmi w::i idjcd.ro carh bco-Tt jnd p J i ^ d .n [v.i; rS-p^Jiiit nu^cciic i ' j f t i iv VtissdswcrcÎbwcdliac^lirtwin^ctr-r iVmtnf.rt^ i t l h c idJ.li<n->iâa>i.l ^ f J S r f t n j U j l L ' ^ J a J F a . i - * p c r . 5 ! c n r>:iFpl.cd Iry Vjr;( o A q u j ^ -hi-n-n lld_M*^-er--; JK' . t t 7ivc e.'20CCc£EaI ^rlhml • IT. each bcikc:. imrrcdtJ^cly jfi'c: ihc 'iicM tTcf :f:e i l c i l iiu^pcnsicn j iMqur* •af wjic.- was'rcr't'^Td u'.fiE ^

vv.'.rigc. Tt-iv pi - i c J u - c vf3\ lefco'cd ii 30 anciGOmin i i K c

liibe fined jnd scl 10 rt;uhi;pjn:vlf i bciwccn -Jin jnd laOjwn in dumeiei. The clffirince ta;v ihe muncii A I S vaL-uIacd us in£ ihe fo!l;>wtn£ rquiMonfCauphtar. (0G!>*;

wTic f f CS,-flfarjr,fe r j i c . tcF;«rnicd i s Lh * . '."•wJumr ct wj ict . C: yni C/r algji o r c c n n i ^ i j r i : h<sxnm5 anJ end at i i i rc

Cfciomic fluid w » Cftttoc:cd irtm between ihi ; irmi («'ihe tea sur^ j n d hjcntdfymph wa^ c x : r j n c d frotr. the p:>stcnor addurisr nîiiclr r>rthrniinfcls,inb>:th j a t n u i i r ^ a t fnl Ayringr Hlicdwiih 9^1 B J i i j c nccdic- The coc'omi: riutdJhiCTtwi»ympS a n g l e s wore uJ^)fcr^rdlruolIultvlduJI9iUc^^^^drpfTClld<>^fIu2>rs'J^

th^rcuabV miied by vartcxin^ pr ior to iist.

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The jsu>' pTMccoi *va? j j j j M c d fr\:ni.'HP •r:*'rKr-d of !iabi.-h. •and 8f-isnfrp3ad 11993J. w,v^h « faaicd cn | h e ahilttv m'"ccl-i t:< r e u n ^ a p w i l j l ncutni! r e d J>T: I'txij^ M\-3S i^. u^rd rauii:«I> in c u : )jbcrji4f>- » J screeninj fwf the ccionrinj i fcn cf t:isde.e{tcci%Qf<tKKOCils w.-J ( u r . t J m i n j m s xzh under and

ilirO ondinors lCin iy c : j l :2007:'R»revdi i-i 3. 200at BrteCy. sampl«,7iTociomxfiurd>hjcfr.u^rnFHi5titj .L i-vereirtaibaicd. m dupjicjie tn a.flai-botto-neil rr.tci»iliireplaic pre-iTMicd witK lilt. \'iv;pclya^ly»inc-»oluiionf!C»a]t':vv am ivtayei .<fcells t_cadhrie ;3 ihf After.4Srnln rcji-adhered ic iU were d i ^ J r d c c and ihr plaics vi-asFcd w l i h pJ>>-'i"^gi'ra; âlint-. A 5*'.uct;ip i f 0.034 V ;v.*.'vinccirjl red dye I'ndui l|ed wa'.cr .SCCuLvwJs added to carh • w e J a n d the CzWs incidiated J>r j b Jt -» C in a [.'injefaor TVr wc^'i Were waihed jiuJan jr .d)l;rd t«h.in->! » J j r i i i n . .A ac r ix arid. ,drd e;hanal m diKiHrd n-ateriadJcJ tJ:'ri:îiwblUsoihcd>-T. The plate was shaken fcr :Smin hef.ire ri-iîng (rfeâbsorhance at 5«J0nm'in an Optimax u n i t l r rrjcrophie readet 'Moleriilw Dcv:ccsS. ut iPg Softpro Mj.»^V.2,^,i I tijf'.WAre Protein canceniia--tl;«r was deieimined ft4t.«ftirg a :ec*-pn'icd me:hjc -'aradhrd, lITfOi .=\rjulw were prei'erted'a* 'ipitcjj Jemiiy p « cf pr:.^ teia

206

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tluiit'hjhemdlvmpt; v»a? gMily, iprtfjj qrat^Ja fr.uiwsitiw sl:iii'

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I'cmsvcd by linnng-.iVb-trc in d i i n j l ^ w j tc : i rd owe the Jlidi'i.

:<jdedjnd randomi«d :hcn cjiamr.ctl iitidtrthc micit^c^ tcf

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KOKH y^jlito f t r ^ h the tpccini and tHr;iricrmle.jp!»ni^^ :hit:

l ionvcju j n«Vrww.T{. For ihcClMnevisMVvtricnv.riMicd i l i i c i

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tfr"ji Myntoi-cdup. In riimn\'>rtiirtjh.^.h(ncjs.whLr A C T

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jFrri «tidy hcwevi:n>-J.'Stf-Ce?e-ndi«^ ifkî wa j fcb ic rvTd in the' m ,:c wnji iBTipc;iei . inin: iC£;. 'ue«:. i i %j»ieen?h?«r.rha:cJf7\' •paleJ•tnr^yis«K^ftj,.^^.î;^-M:iur^.^ni.*,"»r^^'l&a^^3 ,'w.'i.VilivVf5 ihe indi«h.'r «"{l3NA<ir3n'.i hii'-aJ.tftMlawxg.SKp'ja-.ire' Wfrudii ui! r o b i c «. / . 3t>P4'.fji-vp: f.-'r fpe ?*i'itrnnar>- ifud> by Tabai c'l aCf2C!iL-!i 'iH Jd.Ji;WA,urj:tir> flr»ii'[;ro'vet, ar^.sev « i ( «i;dlr>irttaij cTtkr>' «-\a5 Mage* <ei u-chin.-and mussetf;

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jested mori jcorrvâti Jr< 3«vtttrtind_<i jfDNrtaiTjh and i h c cviOfcncrrt c^naifct^'heih'«pac»riiHJiiudd^naid 2£K» ; 'Higg,eret i i . JUtili ' VifleVa ei " ffOrV livîcme. itudiet-ti-jw-evc; inlng crV!^^• *^':ci •C'fiih..v,ti -n fca> ivmt'iiihwerjtJ iuticm^ Biifh.correbl:rni ravi' ivi' .'>t<-' iBelpjiemf ct aL. ilMO ciiing et. V J-Vr*! ;ihrcTC?i'in?ly :hc,pri:f«iar,respnnscs, •3l5,-* »hcwed pKsd c^rrt-ijtiîr.s w:tfi;r«?h^^^^ rt- pnysiolssical retprrudinboih ihf t i ,*jrti)nji.- It-Wi'c-.-cr l necc.ic-texic rctpiintci. w i c delerinjiied i:!. ..'e^<nv>.7:c* i n J hjtir.aciu'% ci'ihc.iVgan-isiri i i i .cellula: It-.ii. Whereii i w nkj'ifi r u r j . mponMS Wif^-ircaturcd ar the %vh.cle.»S[?ihistr:liW'I 'ciivoTt-i.-w.appn Vatfulir syiiem .-rf tfie IWP le*; tpr\2C9..M jrciiM Jsiu^wd.iHaj CIKCT 'JiEci yeUa.couid yJvi Sim

indnrnly") leadinîi'.i iV^piîset is ii-w i.sh(rfc^',":ffani»m jcvf L !frirespeaivfly;uw.i? l^ti^CrrnîiiejandltacTri^ i idercd iana iagpuEi r IIK-kituaiian in c:::iirira£iatt'a turhan healih arrt*3 wlkre uitb:V.;trt»j:a:c\'rIl5li\jr; lv'mph:W>«i.-biinc rn j r row t r l ls lareMt ldsi" iKedt - iv '^ p i i i ienbi JTciivirflTWi talapi'ntswiih^jirierii.i' ^ t > r ^ hej'l;>N^'n«c<;Meiw« iSrgarij/mldfei.tnVeA^^ ^'Vx? jIi>»V»L^^^*'^'"*y-^^ cariier. wecqufanji rihca, O'li*)-*'; ciic-r.'. a rcnelaiion btnveen e!ll?>»'"'ic.efrcn and gen ifttrt effcc. I'js.ceiermlp.ed Mt i^vyf iriOffetcnt.ceatypcj.ii' Ji^cMP-vjî.-vjandmllceiliôfmiHfsHj [mvh<en'foi:t»dl'Cj'p«A.'e»AUiCt4i t i f i i r tCta ie i i^ ikatdi f fcTeni ccHr>TMn(iSwldhf «:rrriliirt^ uilyifiVVii-i^^ fi:llJwrjei('iH)Su;t-ii^ ^ -r! jn-paritTi,

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(ha: necrnicracityi-is h'!fi,.n^ i'!:riiîe-i:cfrc^: i f ECtwaih: j w l chenwihtiapcutik i ? m n :Rf eiki c:'j"! r20M Klsb>-el aL '206c i in *dJiiiijn' ta bcin? iv i r li'aii.-^'mmi.-v Vh- a:iire:nid(^'*'<^'d»s-ruFtin^jgrai U i a s ^ * ' * ' ' s ^ - . ' h a . Ûtla J'^nhL-rniDrc. nrjiriy ScKilibirm tncluJt.VA.Phj'.c:btffn ElH/v:r '.c fcc p'jsm nituiiiftOiiihi h-'th chder fn vKTi- j n J ' '.r.V,;f .-.-nUiff i i inirnjn-nial:i!n:sy?iem5 •nai«mrrrtaL li*4'.-X/p»Mc- iOW^ .S, a. -.T j i . •/pujv Inhuman FttjJrh arena. CTpni^fi »r >*ns.r''nriicnia*,4ffcp'i mil'Jdinjt'ecnSGjVl it;v:riivt>?en-corteij(rt: :( dtvcl jprttr-a. Ji i j rdcri le.^: lej rrtitiie; (lialiihneiiirtenui.a Jcfi. i.dnjriirf^.[r whiA: i r i i«p l jy between eriituSmehtal (ictLr, ' j r j jj/iietK" s-^wt-pr sihiy^^tail^ ab:i ;rfay ar. impi>rtan(jiÎe l?f j i iu-n t: at. ;2f9>i.-l,j'Klr.f_aTi e:~at, 20041; ;r i h i l l.i;nirw.;li«c: .:-an:rraij efh,>î: .rTni; hive zHr^Wi1il> l.» i>ri:het3e.,««e ar'JfclyaArncur'Titar*:!*.^^^ (c.g,,a>c;ykhjlihc^ * e ; iicnin. dapairjov vr^ l in^ p- fcTtK fh t-l^Cui pvpuIal'(^nlQf rev;'pr'jrA:ar,d ihei- ;.viA.::jtr?am wf.^J:i^•:F ja5czliKt3u7.n:i[0V;ffi

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fp e^tfidiiif^v, vibjff vjlj^^ anj crnpIqying jVargc cf r,-i;»' dfnruV;uTi>:ii^m3>Uri aid^ wc hj>t'<âlw!rd the rd^^^ >sni5ii^;ni' o f E w c «f;r;-irn.jthr.rcici

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213

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