Aquatic Toxicology: Mechanisms and Consequences

SYMPOSIUM PROCEEDINGS

Chris Kennedy

Alan Kolok

Don MacKinlay

International Congress on the Biology of Fish University of British Columbia, Vancouver. CANADA

Aquatic Toxicology:

Mechanisms and

Consequences

SYMPOSIUM PROCEEDINGS

Chris Kennedy

Alan Kolok

Don MacKinlay

International Congress on the Biology of Fish University of British Columbia, Vancouver. CANADA

Copyright © 2002 Physiology Section,

American Fisheries Society All rights reserved

International Standard Book Number (ISBN) 1-894337-22-0

Notice This publication is made up of a combination of extended abstracts and full papers, submitted by the authors without peer review. The papers in this volume should not be cited as primary literature. The Physiology Section of the American Fisheries Society offers this compilation of papers in the interests of information exchange only, and makes no claim as to the validity of the conclusions or recommendations presented in the papers. For copies of these Symposium Proceedings, or the other 50 Proceedings in the Congress series, contact: Don MacKinlay, SEP DFO, 555 West Hastings St.,

Vancouver BC V6B 5G3 Canada Phone: 604-666-3520 Fax 604-666-6894

E-mail: mackinlayd@pac.dfo-mpo.gc.ca

Website: www.fishbiologycongress.org

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PREFACE This symposium provides a forum for the presentation and discussion of relevant toxicological investigations involving fish. Like the previous 2 symposia on piscine toxicology in the Fish Biology Congress series, this one was held because there is an immediate and dire need to understand the various direct and indirect impacts of both natural and anthropogenic contaminants on fish and fisheries. The talks in this series range from the molecular level, through biochemistry and physiology, to effects on populations. These papers highlight the multidisciplinary nature of aquatic toxicology, and the diversity of approaches used to understand the mechanisms of toxicity and the relevance of those toxicities to individuals, populations and ecologies. In this series, papers discuss toxic mechanisms, comparative approaches to toxicology, various toxicants including organics and metals, effects on different biochemical, physiological and organ systems, as well as endeavors with applicability to population-level impacts. The cumulative knowledge communicated in this symposium will enhance our understanding of piscine toxicology, an area often underrepresented in the general world of basic and applied fish biology. Symposium Organizers:

Chris Kennedy, Simon Fraser University, Burnaby B.C., Canada

Alan Kolok, University of Nebraska, Omaha NE, USA

Don MacKinlay, Fisheries & Oceans Canada, Vancouver, B.C., Canada

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CONGRESS ACKNOWLEDGEMENTS This Symposium is part of the International Congress on the Biology of Fish, held at the University of British Columbia in Vancouver B.C., Canada on July 22-25, 2002. The sponsors included: - Fisheries and Oceans Canada (DFO) - US Department of Agriculture - US Geological Service - University of British Columbia Fisheries Centre - National Research Council Institute for Marine Biosciences - Vancouver Aquarium Marine Science Centre The main organizers of the Congress, on behalf of the Physiology Section of the American Fisheries Society, were Don MacKinlay of DFO (overall chair, local arrangements, program and proceedings) and Rosemary Pura of UBC Conferences and Accommodation (facility arrangements, registration and housing). Thanks to Karin Howard for assistance with Proceedings editing and word-processing; to Anne Martin for assistance with the web pages; and to Cammi MacKinlay for assistance with social events. I would like to extend a sincere ‘thank you’ to the many organizers and contributors who took the time to prepare a written submission for these proceedings. Your efforts are very much appreciated.

Don MacKinlay Congress Chair

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TABLE OF CONTENTS The impact of metal ions on some components of glutathione cycle in carp gill

cells Arabi, M. and M.S. Heydarnejad.. ......................................................................................1 Biochemical effects of environmental nitrite in matrinxã Moraes, Gilberto, et al......................................................................................................15 DNA synthesis after exposure to heavy metals in the testis of the spiny dogfish Redding, Michael.. ............................................................................................................27 Quantitative PCR of Atlantic salmon CYP1A in gills Rees, C, S McCormick, and W Li......................................................................................31 Expression of multi-drug resistance (MDR) p-glycoprotein and associated

proteins (MRPs) in the flounder; effects of exposure to the polyaromatic hydrocarbon, benzo(a)pyrene

Cramb,G, Cutler,et al. .. ...................................................................................................33 Sediment-associated pollutants in the marine environment: a multi-biomarker

approach for assessing sediment toxicity in turbot Kilemade, M., Hartl, M. et, al. .........................................................................................39 Responses of Atlantic salmon and bivalve molluscs to paralytic shellfish Eddy, FB, et al. .. ..............................................................................................................43 The effect of environmental levels of freshwater contaminants on juvenile

Atlantic salmon: Implications for marine survival Lower, Nicola and Andy Moore........................................................................................47 Chloride cell changes induced by nitrite exposure in an amazonian fish species Da Costa O.T.F. and M.N. Fernandes..............................................................................51 Behavioral and Neurophysiological Effects of Carbamate Pesticides on

Olfactory Capabilities in Pacific Salmon Jarrard, H and C Kennedy.. .............................................................................................63

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Effects of exposure to sub-lethal concentrations of ammonia and hypoxia on the swimming performance of brown trout

Shingles, A. et al. .. ...........................................................................................................69 Effects of chlorpyrifos (Lorsban)on reproductive performance of guppy De Silva, M.. .....................................................................................................................75 Maternal transfer of copper resistance in fathead minnows Peake, EB, et al. ...............................................................................................................87 Reduced gonadal somatic index and external coloration following exposure to

P,P’-dde in adult male Fundulus heteroclitus Monosson, Emily et al. .....................................................................................................93 Blood cell responses of the tropical fish Prochilodus scrofa to acute copper

exposure and subsequent recovery Cerqueira, Carla and M.N. Fernandes.............................................................................99 Toxicity of cadmium: a comparative study in the airbreathing fish, Clarias

batrachus and in non-airbreathing Ctenopharyngodon idellus Joshi, P.K. & M. Bose.....................................................................................................109 Effects of vanadate oligomers on lipid peroxidation and antioxidant enzymes in

the Lusitanian toadfish kidney and liver: short term-exposure Figueiredo, Inês, et al. ...................................................................................................119 Incubation and pH-dependent effects of vanadate oligomers and cadmium with

Halobactrachus didactylus sarcoplasmic reticulum calcium pump Leonardo, Lília, et al. .. ..................................................................................................123 Histological analysis of vanadate oligomers effects on heart, kidney and liver of

the Lusitanian toadfish: an acute exposure study Borges, Gisela, et al. ......................................................................................................129 Cadmium and vanadate oligomers comparative effects on the toadfish

erythrocyte Soares, Sandra, et al. .. ...................................................................................................133

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Toxicity of Lake Enrichment nutrients to aquatic life MacKinlay, Don and Craig Buday .................................................................................139

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THE IMPACT OF METAL IONS ON SOME COMPONENTS OF

GLUTATHIONE CYCLE IN CARP GILL CELLS

Mehran Arabi and Mohammad Saeed Heydarnejad

Department of Biology, Shahrekord University, Shahrekord- 88186, POB 115 , Iran .

mehranarabi@yahoo.com Abstract The extent of cellular damage was investigated after in vitro addition of two metal ion componds Viz. CuSo4 and HgCl2 in various concentrations 300, 500, 700, 1000 and 3000 mM to gill cells preparation of freshwater Fish Carp (Cyprinus carpio L.). The objectives of current study were to determine the influence of these metal ions on the levels of TBARS/Lipid peroxidation (LPO) and GSH, also to evaluate the role of BSA (0.5 & 1.0%) and DMSO (0.5%) as ROS scavangers to encounter the relative processes. The outcomes of this report are: (a) Copper and mercury increased the rate of LPO dose-dependently ( r = + 0.995 and r = + 0.993 respectively, P< 0.001) but the GSH Content was only marginally affected ( r = -0.787 and r = -0.844 respectively, P <0.05). (b) CuSO4 played a more potent role in oxidative damage to fatty acid chains than HgCl2 in gill cells preparation. (c) Depleting of GSH molecules by copper had a wider range than mercury. (d) In the highest concentration of metal ions (3000 mM) both DMSO and 1.0% BSA showed a pro-oxidative potential to elevate the levels of TBARS (P<0.001) but for other concentrations when supplemented with three scavanges , it was found a fall in the levels of the latter. (e) Addition of 1.0% BSA to medium containing 3000 mM of metal ions caused a significant decline in GSH content (P<0.01). Collectively, these findings indicate that copper and mercury have deleterious influence on membrane integrity and GSH content as well in a relative dose-dependent manner. The complexes of metal ions and thiol residues of cell proteins could also act as a more potential cell toxicant leading to disturbances in cell functions towards cell death.

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Introduction Common Carp (Cyprinus carpio L.) is an important commercial species around the world to feed populations and is as an economic rather than an ornamental fish. As a group, carp provide 4 million metric tonnes of fish annually - over a quarter of all fish culture worldwide. Reducing the number of these precious animals for water contamination (e.g. metal ion toxicity) is as a paradox to heavy demands. Teleosts, functionally, have four pairs of gill arches furnished with tiny structures called gill lamellae. The latter, are rich in capillary networks and covered with a simple squamous epithelial cells which are responsible for gas exchanges in aquatic media. Due to direct exposure of gills in the water medium, it has been dominantly accepted that they are the main site to water contamination and toxicity (Karan et al. 1998; Dalzell and Macfarlane 1999). Coping with to fish is very crucial. The water companies are routinely consuming metal ions in aquatic media e.g. Iron sultate is used to combat algal blooms in reservoirs (Hayes et al. 1984); Copper sulfate is one of the most widely used algicides for the control of phytoplanktons in lakes, reserviors, and ponds, it is also used for aquatic weed control (Karan et al. 1998); Mercuric chloride is also mainly used to control the mass of other fish partners as mollusicide (Radhakrishnaiah et al. 1993). It has been shown that some metal ions in fish express deleterious impacts on some organs such as liver, gill, gonads and components of blood as well (Radi and Matkovics 1998; Mukherjee et al. 1994; Karan et al. 1998 and Akahori et al. 1999). A number of researchers have observed that the reactive oxygen species (ROS) such as superoxide anion (.O2

-) and hydroxyl radical (.OH) through their own generating systems can stimulate the peroxidation of polyunsaturated fatty acids (PUFAs) of the biological membranes which is called lipid peroxidation, LPO (Halliwell and Gutteridge 1985; Radi and Matkovics 1988). Continued fragmentation of fatty acid side chains to produce further more aldehyde and hydrocarbons will eventually leads to LPO propagation and complete loss of membrane integrity and cell functions. Therefore, LPO is an useful index to measure the membrane integrity and relative lesions. To overcome this degenerative process cells are well-equipped with a powerfull battery of defence systems (Antioxidants) to mop up and then neutralize the ROS (Sharma and Agarwal 1996). The ubiquitous tripeptide, reduced glutathione (L-gamma-glutamyl-L-Cysteinylglycine, GSH), most prevalent intracellular thiol ( - SH) functions as a

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quick and vibrant antioxidant in cells by donating one hydrogen atom from two molecules to a toxic substance (Meister and Anderson 1983). Glutathione is present in the oxidized (GSSG) form, which is readily converted to the GSH form by the enzyme glutathione reductase ( GRD). It has been reported that glutathione is present mainly in its reduced form in biological tissues, at concentrations as high as 2180 mg/g of tissue and in form of GSSG to be present in much smaller concentrations, ranging from 0 to 288 mg/g of tissue (Hissin and Hilf 1976). A shift to a more oxidative state or any imbalance between production and degradation of ROS in animal tissues may causes LPO, plasma membrane alternations, inactivation of enzymes, ect. (Radi and Matkovics 1988; Matta et al. 1999; Akahori et al. 1999; Anand et al. 2000). The current study aimed to investigate the influence of different concentrations of two metal ion compounds viz. copper sulfate (CuSo4) and mercuric chloride (HgCl2) on some properties of carp gill cells; membrane integrity and GSH content. We also estimated the scavanging activity of 0.5 & 1.0% small molecule protein, bovine serum albumin (BSA) and 0.5% dimethylsulfoxide (DMSO) to possibly generated ROS in different media. Material and Methods Reagents Thiobarbituric acid (TBA), and DTNB (5,5'-dithio-bis [ 2-nitrobenzoic acid]) were obtained from Merk, Darmstadt ( Germany). All other chemicals were from Himedia company (India) and were of analytical grade. All solutions were made with triple-distilled water. Collection of Samples Six fresh carps were collected from local fish farm. As soon as possible, branchial arches were gently taken out and kept in chillled sorenson's buffer (pH 7.4) and transfered to lab and kept at 4°C for a short period of time till final processing. Protocol for the estimation of Lipid peroxidation and GSH Content a) Preparation of tissue homogenate After careful removal of the cartilageous portions supporting gill arches, the gill filaments gently cleared of other parts and washed several times with chilled 0.154 M NaCl solution (0.308 Osmolar) to remove blood. After that, the bulk of gill fillaments grinded resulting in bared ones and free bands of lamellae. The latter, separated from bony parts by means of ordinary tea-strainer.

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Consequently, the resulting filtrate homogenized in 0.05 M phosphate buffer (pH 7.4) using polytron homogenizer at a speed of 13000 rpm in ice bath. The homogenate was then centrifuged (3000 rpm, 10 min). The supernatant was used for assays throughout the work. All the above steps commencing from dissection of the tissue till the preparation of homogenate were performed at 0-4°C. b) Estimation of Proteins Cell protein content was estimated by the modified sodium dodecyl sulphate - Lowry method of Lees and Paxman (1972) using BSA as standards. Activities of measured parameters were calculated and expressed per mg protein. c) Assay of Lipid Peroxidation by Spectrophotometric method The LPO was estimated in terms of Thiobarbituric acid reactive substances (TBARS), particularly Malondialdehyde (MDA) by the method of Ohkawa et al. (1979) with slight modifications. The reaction mixture (3ml) contained 0.1 ml of homogenate supernatant, 0.2 ml of 8.1% sodium dodecyl sulphate, 1.25 ml of 20% glacial acetic acid (pH 3.5), 1.25 ml of 1.2% aqueous solution of TBA and 0.2 ml distilled water in control group instead of adding 0.1 ml of metal ions or antioxidants in treated one. Finally, after heating, adding 3 ml of n-butanol-pyridine mixture and centrifuging at 2200 xg, the amount of MDA formed was measured by the absorbance of the upper organic layer at 532 nm which is the λmax of MDA (Extinction Coeff. of 1.56 x 105 / M/cm) using appropriate controls. d) GSH content measurement Reduced glutathione residue after treatments as well as in control group was assayed according to the method of Sedlak and Lindsay (1968) using DTNB with minor modification. Statistical Analysis Data reported in the paper are means of two or more assays. All measurements were performed in triplicate. The data are given as means + standard deviation (S.D.). The comparison of the control and treated series was statistically analysed by Student's t-test and validity of investigation was expressed as probablity (p) values. Values of P>0.05 and P<0.05 were considered to be non-significant and almost significant; values of P<0.01 and P<0.001 were considered significant and very significant respectively. Correlations were

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evaluated using linear regression (LR) analysis between parameters assayed in the present study. Results Influence on TBARS levels As showed in table 1 production of MDA/TBARS in carp gill cells preparation was induced by the presence of various concentration of copper in a dose-dependent and linear manner (p<0.001, t-test). There was a positive and strong correlation (r = +0.995, see fig.1A) between the elevated Cuso4 concentrations and TBARS levels. When ROS scavangers (DMSO & BSA) supplemented to the media a sharp significant decrease was observed in TBARS levels (P<0.001) for all concentration of Cuso4 as compared to treated ones with same concentration of metal ion (alone), but here at the concentration of 3000 µM for DMSO and 1.0% BSA instead of decreasing there was a dramatic increase in TBARS production (P<0.001) which is due to their pro-oxidant activities. Also all three scavangers exerted a significant fall in TBARS levels as compared to control group (no metal ion treatment and supplementation) which expressed a spontaneously LPO production in medium, Table 1 (P<0.01 & P<0.001, cf.different data). The data demonstrated in Table 2 show that the addition of upgrade concentrations of HgCl2 caused highly significant elevation in TBRAS production (P<0.001). The strength of associatioin between metal ion effect and LPO generation was found positive and also considerable (r=+0.993, see fig. 1B). As soon as supplementation it was clearly observed that BSA (0.5 & 1.0%) can inhibit the extent of MDA/TBARS production (P<0.001) when compared to the same samples treated with metal ions (alone) but DMSO affected the LPO rate only marginally (P<0.05 & P <0.01, cf. data). Here again we found out that when 0.5% DMSO and 1.0% BSA added to media with 3000 �M metal ion a drastic increase in TBARS levels was occured (P<0.01 & P<0.001, respectively). Addition of the all three scavangers to control group resulted in a fall in TBARS levels (P<0.01 & P<0.001, see Table 2). The protective effect of 0.5% BSA for holding the membrane integrity was found much more better than DMSO & 1.0% BSA (P<0.001). Influence on reduced glutathione (GSH) content Table 3 summarizes the effect of CuSO4 concentrations on the GSH Pool in Carp gill cells preparation. This evaluation showed only 3 concentration of

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copper (700-3000 mM ) could change the GSH content which was a slightly significant decrease (P<0.05). A negative correlation represented between metal ion influeuce and GSH levels in preparation (r = -0.787, see fig. 1A). No significant change was observed in the levels of GSH when DMSO & 0.5% BSA were added (P>0.05). There was an interesting finding when 1.0% BSA added to assay mixture that was a significant diminution for GSH content as compared to control value (P<0.05 & P<0.01, cf. different data), Moreover, supplementation of the ROS scavanger 0.5% BSA to control group (alone) led to an elevation in GSH levels (P<0.05) and no significant alteration following addition of two others (P>0.05). The data mentioned in Table 4 reveal that addition of various concentration of mercury (HgCl2) to gill cell preparations of carp could substract the levels of GSH dose-dependently but slightly significant (P<0.05) for two higher concentrations namely 1000 & 3000 µM. A negative significant correlation existed between this parameter and metal ion concentrations which is given in fig. 1B (r = -0.844). Supplementation of DMSO to assay mixture containing metal ion resulted in a slight diminution but non-significant (P>0.05) in GSH content. 0.5% BSA caused an almost significant elevation (P<0.05) to GSH content when compared to respective samples treated with same metal ion concentration (alone) (1000 & 3000 µM). Addition of 1.0% BSA could not change the GSH levels significantly in series treated with mercury except at the highest concentration which led to a sharp significant (p<0.01) diminution as compared to control group (cf. Table 4). Among the ROS scavengers used in this experiment only 0.5% BSA could change and elevate the levels of GSH when compared to the control data (P<0.05). Discussion As cited in the introduction there is a growing body of evidence to indicate that metal ions can exert some deleterious impacts on a number of tissues of fish Carp (Cyprinus Carpio L.). For instances, mercury concentration effect in vivo model is believed to be associated with lowered succinate dehydrogenase activity and O2 consumption (Radhakrishnaiah et al 1993) and also an induction of severe proteolysis in carp gill cell metabolism (Suresh et al 1991). Radi et al (1988) reported a sharp fall in antioxidant activity of some enzymes e.g. glutathione peroxidase (GPx) and an elevation in the rate of LPO and also protein contents after mercury treatments in liver, white muscles and gill of carp. On the other hand, copper increases the activity of some key metabolism enzymes such as alkaline phosphotase, alanine and aspartate aminoteransferases

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and showed induction of some lesions in carp gill e.g. epithelial hyperplasia and chloride cells dysfunction (Karan et al 1998). In addition Kurant et al (1997) reported that copper treatment in carp liver cells induces a diminution in the content of low molecular weight protein thiols. The reports on the effect of metal ions toxicity as an in vitro model are not too many, this kind of work is able to open new lines of intracellular compartments when exposed to pollutants. Owing to the role of metal ions as important catalysts in living organisms finding the knowledge regarding to other facets of these compounds sounds much vital. Under the present experimental conditions, the results presented here have clearly demonstrated that the elevated metal ion concentrations for copper and mercury is associated with high production of the TBARS (p < 0.001, Table 1 & 2). We observed a strong link between concentration effect of metal ions used and extent of TBARS. The correlation coefficients for copper and mercury amounted to + 0.995 and + 0.993 at 370C respectively (p < 0.001, Fig. 1. A & B). Here, copper and mercury acted as potent oxidants to biological membranes, both plasma and intracellular membranes. Therefore, these results indicate that in spite of physiological role of metal ions in maintenance of cell functions, could also function as toxic agents when present in excess. The metal ions as transition metals cause cellular damages via formation of highly reactive oxygen free radical viz .OH. LPO initiation phase results in the formation of lipid hydroperoxides (Lipid – OOH) in the presence of .OH (Sharma and Agarwal 1996) which is derived from .O2

- and hydrogen peroxide (H2O2), generally metabolically generated in medium. Neither .O2- nor H2O2 is energetic enough to initiate LPO directly, but in presence of catalytic amounts of metal ions, they can react and form .OH radicals under a net equation, Harber – Weiss reaction (Halliwel and Gutteridge 1984): metal ion .O2

- + H2O2 . O2 + .OH + OH- catalyst Here, the reduction of membrane integrity and fluidity is as a consequence of enhancement of LPO process. On the other hand, with commencing LPO cascade, the release of significant quantities of lipid – OOH in medium will be occurred, which is presumably a consequence of the activity of membraneous enzyme phospholipase A2 (PLA2) by cleaving oxidized PUFAs from the 2-position of phospholipid glycerol backbone so that they can then be

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metabolized by GPx to the corresponding alcohols (Halliwell and Gutteridge 1985; Alvarez and storey 1995). It has been determined that metal ions like mercury can alleviate the rate of GPx activity in carp gill and liver resulting in accumulation of lipid – OOH in cell and LPO propagation throughout the cell membranes which injures the membrane integrity (Radi et al 1988). Furthermore, as Huang et al (1993) and Halliwel (1995) suggested it is possible that lipid-OOH formed in the biological membrane systems by the effect of metal ions convert to the form of peroxyl radical (Lipid – O2.) which is able to reattack the membrane. It is of great interest to know that MDA formed during LPO could conceivably reacts as bifunctional reagent to cross-link through schiff’s base formation between some classes of phospholipids (Eichenberger et al 1982). Lack of the uniformity to these cross-links in membranes will be led to physical forces which may disturb the membrane lipid distributions (Yuli et al 1981). We found that GSH was marginally affected by the elevated metal ion concentration effect (p < 0.05, Table 3 & 4 and fig. 1 A & B) to a lesser content. GSH can react with peroxyl radicals to achieve a steady state for themselves and itself converts to thiyl radical (GS.) which is not rapid as oxidative attack at polyunsaturated membrane lipids: GSH + Lipid – O2. lipid – OOH + GS. It seems that the metal ion extrusion from cells presumably involves movement of diffusible complexes such as Hg – GSH, it also gives rise to alleviating in the level of thiol groups pool including GSH in medium. Due to more reactivity of copper with GSH molecules than mercury, it can be assumed that copper could present more oxidative stress on living organism making its cells being more susceptible to damages. Despite the extensive evidence implicating the depletion and / or oxidation of GSH in a wide variety of experimental toxicities (Smith et al 1996), here it has been shown that GSH content was not much altered under the influence of metal ion stress (p < 0.05). An explanation here is that relative inhibition following metal ion impact on some enzymatic defence systems such as GPx (Radi et al 1988), converts lipid – OOH to respective alcohol through oxidation of GSH to GSSG, causes no remarkable alteration in GSH content due to less GSH consumption by inhibited enzyme. Supplementations were carried out to check the protective role of BSA (0.5 & 1.0%) and DMSO (0.5%) in media. DMSO as a lipophilic OH. scavanger

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normally is used to mop up ROS and protect biological membranes against oxidation. On the other hand, BSA also as a OH. scavanger might play a far more important role in protection of plasma membrane from peroxidative injuries by binding to and neutralizing the lipid hydroperoxides liberated by PLA2 (Alvarez and storey 1995). BSA can impair the LPO cascade either by acting as a sacrificial antioxidant or as a chelator of the metal ions that promote this process (Halliwel and Gutteridge 1989). Another intriguing possibility from cha and Kim (1996) is that BSA is able to exhibit peroxidase activity, providing that thiol-reducing equivallents are available. Addition of 1.0% BSA and DMSO caused a significant fall in the level of TBARS upto concentration of 1000 �M of metal ions and after that led an increase in MDA production (Table 1 & 2), this is due to of the fact that thiol groups which exist in these antioxidants could play a multifaceted role as pro-oxidants by autoxidizing to produce .O2

- in the presence of metal ions (Tien et al 1982). R – SH + metal ion (oxidized) RS + metal ion (Reduced). Metal ion (Reduced) + O2 metal ion (oxidized) + .O2

- Collectively, our results showed that 0.5% BSA exerts a positive influence in protection against fatty acid peroxidation than 1.0% BSA and DMSO as well (Table 1 & 2) in control and treated ones. The intracellular compartmentations of GSH molecules including the nucleus, mitochondrial matrix and endoplasmic reticulum have many important implications for cells that are exposed to toxic compounds or to other stresses (Smith et al 1996). It is noteworthy that there is a kinetically distinct pool of GSH in the nucleus, estimated to be 5 to 10% of the total GSH and may be concentrated at the nuclear membrane surface (Loh et al 1990; Britten et al 1991) when an adequate stimulation is made, it might be released and added to cytoplasmic pool. We found that among the antioxidants tested only 0.5% BSA yielded an alteration in the GSH content of the control group of preparations (p < 0.05, see Table 3 & 4). It seems that BSA acted as a stimulation either for releasing the stored GSH molecules from their cellular compartments or to detach the masked GSH molecules from trapping points in the preparation. However the exact mechanism of this phenomenon remains to be elucidated. As mentioned earlier, thiol groups in cases of DMSO and 1.0% BSA could imposed a severe peroxidation on cell membranes, however 1.0% BSA has been

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found to be more toxic than DMSO to substract the GSH content when accompanied with metal ions particularly in higher concentrations. More analysis needs to be conducted. Taken together, these findings revealed that used metal ions namely copper and mercury are very harmful to gill cells of fish carp when are expressed in excess. These metal ions not only act as anti – fluidity agents for membrane lipids but also in making a complex with cell protein thiols will be able to express more toxic effects to push the cell towards dysfunction. Acknowledgement We are extremely grateful to Dr. (Mrs.) Noha Eftekhari for conducting supplemental statistical analyses to check those data reported here. References Akahori, A. Gabryelak, T. Jozwiak, Z. and Gondko, R. 1999. Zinc – induced

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and cadmium chloride on gonadal function and its regulation in sexually mature common carp cyprinus carpio. Biomed. Environ. Sci. 7(1): 13 – 24.

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tissues by thiobarbituric acid reaction. Anal. Biochem. 95: 351 – 358. Radhakrishnaiah, K. Suresh, A. and Sivaramkrishna, B. 1993. Effect of

sublethal concentration of mercury and zinc on the energetics of a freshwater fish cyprinus carpio (Linnaeus). Acta. Biol. Hung. 44 (4): 375 – 385.

Radi, A.A., Matkovics, B 1988. Effects of metal ions on the antioxidant enzyme

activities, protein content and lipid peroxidation of carp tissues. Comp. Biochem. Physiol (C). 90(1): 69 – 72.

Sedlak, J. and Lindsay, R.H. 1968. Estimation of total, protein – bound and non

– protein Sulfhydryl groups in tissue with Ellman’s reagent. Anal. Biochem. 25: 192 – 205.

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Sharma, R.K. and Agarwal, A. 1996. Role of reactive Oxygen species in male

infertility. Urology. 98(6): 835 – 850. Smith, C.V. Jones, D.P., Guentiner, T.M. and Lauterbury, B.H. 1996.

Compartmentation of Glutathione: implications for the study of toxicity and disease. Toxicol Appl. Pharmaco. 140: 1 – 12.

Suresh, A. Sivaramakrishna, B. Victoriamma, P.C. and Radhakrishnaiah, K.

1991. Shifts in protein metabolism in some organs of freshwater fish, Cyprinus carpio under mercury stress. Biochem Int. 24 (2): 379 – 89.

Tien, M. Bucher, J.R. and Aust, S.D. 1982. Thiol – dependent lipid

peroxidation. Biochem Biophys Res. Commun. 107 (1): 279 – 285. Yuli, I. Wilbrandt, W. and Shinitzky, M. 1981. Glucose transport through cell

membrane of modified lipid fluidity. Biochem. 20 : 4250 – 4258.

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14

BIOCHEMICAL EFFECTS OF ENVIRONMENTAL NITRITE

IN MATRINXÃ (Brycon cephalus)

Gilberto Moraes Department of Genetics and Evolution, Federal University of São Carlos. Rod.

Washington Luiz, Km 235. São Carlos, SP, Brazil. CEP 13565-905.

E-mail <gil@power.ufscar.br>

Ive Marchioni Avilez, Alexandre Eneas Altran and Lucia Helena de Aguiar

Department of Genetics and Evolution – Federal University of São Carlos.Rod. Washington Luiz, Km 235.

São Carlos, SP, Brazil. CEP 13565-905. Introduction

Nitrite usually occurs in aquatic environments as a product of bacterial activity. It normally comes from oxidation of ammonia and this (nitrification) depends on the water aeration. Another source of water nitrite is the industrial wastes (Nikinmaa, 1992; Heckman et al., 1997). The fish-culture systems also increase ammonia and nitrite, followed by undesirable consequences (Hargreaves, 1998; Hagopian & Riley, 1998). The organismal disorder arisen from such environmental disturbances are particularly observed in fishes.

The prompt nitrite effect in fishes is observed at the blood level. Plasma can accumulate it (Shechter, 1972), working as a vehicle to spread it over the tissues. Within the red blood cells it oxidizes the hemoglobin-Fe2+ to Fe3+ yielding methemoglobin, unable to transport oxygen. This effect is supposed to result in tissue hypoxia (Cameron, 1971; Bath & Eddy, 1980; Doblander, et al., 1996, Vedel, et al., 1998) even in the presence of oxygen (functional hypoxia). The intensity of methemoglobin formation is dependent on the non-oxygenated level of hemoglobin (Jensen, 1990; Jensen, 1992; Willians, et al., 1993). Methemoglobin content varies among the species and depends of the nitrite levels and the exposure time. The increase of nitrite concentration increases lead to raise methemoglobin concentration (Schoore, et al., 1995). Usually, the fish plasma concentration of nitrite is higher than environmental one.

15

Environmental nitrite can cause physiological disturbs as decreases in total hemoglobin, hematocrit and RBC. This phenomena can be explained by erythrocyte hemolysis (Kundsen & Jensen, 1997).

In freshwater teleosts, nitrite exposure is followed by several osmoregulation responses as hyponatremia, hypochloremia (Jensen et al, 1987), branchial chloride cell failure (Gaino et al, 1984), and inhibition of chloride uptake (Willians & Eddy, 1986).

Nitrite exposed fish is completely recovered in clean water (Huey & Beitinger, 1981), and a couple of mechanisms involved in such process are proposed. The first is NADH-Methemoglobin reductase system (Diaforase I) reduces hemoglobin-Fe3+ to Fe2+ and it role in nitrite detoxification has been studied in fish (Scott and Harrington, 1985; Woo & Chiu, 1997). The nitrate synthesis, the more oxidized form of nitrite, is a second way of detoxification (Doblander & Lackner, 1997) and catalase and citocrome oxidase-aa3 is proposed to take a share in such process (Doblander & Lackner, 1996). However, both mechanisms are still unclear and further studies should establish the role of those enzymes in fish detoxification of nitrite.

In this study the environmental nitrite effects hematological and osmoregulator response, the methemoglobin formation and NADH-methemoglobin reductase system will be investigated in the neotropical teleost Brycon cephalus (matrinxã).

Material and methods

Juveniles of B. cephalus (matrinxã) ranging 90 ± 5g (means ± SD) were obtained from the fish farm Aguas Claras, Mococa, SP, Brazil. The fish were brought to the aquaculture facilities of the Comparative Biochemistry Laboratory. Before the experiments, fish were equally divided in four tanks of 250L, covered with black plastic sheets, provided with well-aerated water. Quality of water was measure by APHA (1980) (pO2 = 7.5 ppm, pH = 6.8 ± 0.2, temperature = 23 ± 1oC, conductivity = 74.3 ± 4.8 µS. cm-3, alkalinity = 37 mg/L of CO3

-, hardness = 28 mg/L of CO3-, ammonia concentration = 0,01 p.p.m,

nitrite concentration = 0 ppm). The indoor-experimental tests were performed in August.

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Experimental design

Kept starved for 1 day, one tank nitrite free was the control. In a semi-static system (with water replace each twenty four hours), six fish per tank were exposed to 1, 2, and 3 p.p.m of NO2

– by 48 h. After this, the fish were collected and anesthetized with MS 222. A blood sample was drawn from the caudal vein and divided into aliquots for further analysis.

Blood analysis

Microhematocrit was done with blood samples centrifuged at 12.000 g for 3 min in capillary tubes. Total hemoglobin was colorimetricaly determined at 540 nm in samples containing 10 µL of blood into 2.0 mL of Drabkin solution. Methemoglobin was optically quantified at 563 nm as Matsuoka (1997). Red blood cells were counted under light microscope with a Neubauer chamber and the mean corpuscular volume (MCV), the mean corpuscular hemoglobin (MCH) and the mean corpuscular hemoglobin concentration (MCHC) were calculated.

One blood aliquot was centrifuged at 12.000 g for 3 min and the plasma was used to flame photometric determinations of Na+ and K+, and optical determination of Cl– at 480 nm (APHA, 1980) and NO2

– at 520 nm (Shechter, et al., 1972).

NADH-Methemoglobin reductase system

One aliquot of blood was re-suspended into 0.9% saline solution and centrifuged at 1.000 g for 10 min. This procedure was repeated thrice and the cells were re-suspended into 0,04 mL of mercaptoethanol-EDTA solution and the erythrocytes were lysed by termal shock (with liquid nitrogen). This hemolysis was used as enzyme source. The enzyme assay was performed into a buffer solution 0.2M Tris-HCl pH 7.5, 1,2 mM 2.6-dichlorophenol indophenol, 6 mM NADH, and a suitable enzyme aliquot. The substrate consume was optically followed at 600nm and one unit equals a decrease in absorbancy of 1.0 per minute, in 25°C as Beutler (1984). The specific activity is expressed U (units) per mg of hemoglobin (U/mg total Hb).

Statistics

For comparison of the dates was used STATISTICA 5.5 software. Normality of the data set was evaluated by the SHAPIO-WILL W test with 95% of

17

confidence limit. The parametric test ANOVA was used to compare the groups and the Post-Test of multiple comparisons DUNCAN was applied considering p < 0.05.

Results

Increasing concentrations of environmental nitrite affected the blood parameters of matrinxã (Table I). The methemoglobin and the plasma nitrite increased very sharply keeping high values (figure 1).

Hematocrit decreased in the all the nitrite exposures, however total hemoglobin and the red blood cell number did not change. The MCV, MCH and MCHC did not change too.

The NADH- methemoglobin reductase enzyme system was detected in the red blood cells of B. cephalus. . That enzyme activity was found in all the fish and it was not affected by nitrite exposure (Table1). The figure 1 shows the trend of this enzyme during the nitrite exposure for 48 hours.

No significant effects were found in plasma protein, K+, Na+ and Cl– under nitrite exposure.

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Table 1.Hematological parameters of Brycon cephalus exposed to environmental nitrite for 48h.

PARAMETER NITRITE (ppm)

Total blood 0 1 2 3 MetHb 0.17± 0 56.1* ± 6 84.3* ± 7 78.4* ± 8

Ht 35.8 ± 2 26.0* ± 3 24.0* ± 4 23.9* ± 2 Total Hb 9.06 ± 2 8.50 ± 2 7.35 ± 1 8.10 ± 1

RBC 2.40 ± 0.2 1.86 ± 0.5 2.00 ± 0.6 1.98 ± 0.3 MCH 37.96 ± 4.9 47.30 ± 4.8 39.02 ± 2.6 41.72 ± 4.1 MCV 133.6 ± 13 145.6 ± 14 127.4 ± 12 122.4 ± 6

CMCH 2.85±0.39

3.22±0.40 3.05±0.41 3.41±0.42 NADH-MetHb reductase 0.33±0.14 0.52±0.20 0.47±0.06 0.43±0.12

Plasma NO2

- 0.01 ± 0.00 0.23* ± 0.02 0.34* ± 0.04 0.73* ± 0.05 Na+ 144.7 ± 12 130.5 ± 10 128.5 ± 12 103.0 ± 9 K+ 3.1 ± 0.5 2.7 ± 0.3 3.4 ± 0.3 3.2 ± 0.4 Cl- 130.4 ± 6 134.2 ± 7 123.1 ± 12 110.6 ± 12

Protein 0.45 ± 0.04 0.49 ± 0.05 0.43 ±0.04 0.41 ± 0.01 The values are expressed as: Ht (%), Total Hb (g.dL–1), Red Blood Cells-RBC (106.mm–3), Mean Corpuscular Hemoglobin-MCH (pg.cell–1), Mean Corpuscular Volume-MCV (µmm3), mean corpuscular hemoglobin concentration –MCHC (%), NADH-methemoglobin reductase (U.mg total Hb-1) Methemoglobin-MetHb (%), Na+ and K+ (mEq.L–1), Cl– (nmol.mL–1 ), Protein (mg.mL–1), NO2

– (nmol.ml–1). The mark (*) means significantly different (p< 0.05) as compared to the control.

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0 1 2 3Nitrite ppm

0

25

50

75

100

Met

aHb

Redu

ctas

e %

0

25

50

75

100

Met

Hb

%

0

25

50

75

100

Plas

ma

NO

%

Figure 1. Relative concentration (%) of blood methemoglobin reductase, methemoglobin and plasma nitrate of Brycon cephalus (matrinxã) exposed to NO2

- for 48 h, considered as 100% the maximum value of the parameter.

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Discussion

Nitrite exposure in fish is supposed to result in tissue hypoxia that usually causes significant stress (Huey et al, 1980, Arillo et al, 1984, Hilmy et al, 1987). As a common classical response to stress from hypoxia, the increase of hematocrit should be expected. This strategy increases the number of red cells and the content of hemoglobin to keep the oxygen availability (Swift, 1981; Peterson, 1990). However, these responses were not detected in several fishes (Eddy, et al., 1983; Hilmy, at al 1987; Tucker, at al 1989; Jensen, 1990; Knudsen & Jensen, 1997; Woo & Chiu, 1997), or are thinly discussed. In this particular, decrease of the hematocrit of matrinxã can be attributed to the blood cell lyses (Jensen, 1990; Knudsen & Jensen, 1997), for the reduction of number of cells without changes of MCV, MCH and MCHC. This response should remove the blood methemoglobin but it will reduce the hemoglobin availability. Decrease of hematocrit in matrinxã without changes of the red blood cell number and total hemoglobin is suggestive of hemolytic anaemia, which is a posterior response to functional anemia (Scarano & Saraglia, 1984). Those authors propose the increase of methemoglobin as an early functional anaemia.

The NADH-methemoglobin reductase system has been detected in the most animals in nature, as well as in .B. cephalus. This enzyme recovers the hemoglobin from methemoglobin keeping the equilibrium between both forms. Among the fishes, it was reported in the channel catfish Ictalurus punctatus (Huey & Beitinger, 1981), the rainbow trout Salmo gairdneri, Oncorhynchus kisutch, Oncorhynchus nerka, Salmo malma (Scott and Harrington, 1985) and Lates calcarifer (Woo and Chiu, 1997) and others.

Several studies attribute to the nitrite exposure the increase of methemoglobin concentration. Also, the recovery of the methemoglobin levels to normal values is observed as the fish return to nitrite free water. Schoore and col (1995) attribute this fact to the activity of NADH-methemoglobin reductase system, in spite of it was not assayed. One study on enzyme changes in fish exposed to nitrite is reported by Woo and Chiu, (1997) but no significant change was observed. The same occurred in matrinxã exposed to nitrite, as the level of NADH-methemoglobin reductase system did not change under any level of environmental nitrite. However, the presence of that system is very important for the species because it prevents the hemoglobin oxidation. The fact of NADH-methemoglobin reductase system be unchanged in matrinxã, does not

21

mean that this enzyme is not working, but more likely it was not induced as proposed to Lates calcarifer (Woo & Chiu, 1997).

As well as the nitrite attacks the heme group of hemoglobin it is possible that other heme proteins are affected, like the NADH-methemoglobin reductase system. This enzyme also presents a heme group in the molecular structure (citocromo b5). This attack could camouflage an increase of this enzyme production. In rainbow trout, Arillo and col. (1984) suggest that nitrite could attack hemoproteins as citocrome P450 in liver.

The main characteristic of fish nitrite exposure is the increase of methemoglobin and plasma nitrite concentration (Cameron, 1971; Shechter et al, 1972). These levels are usually very low and the increase depicts different trends. The increase of plasma nitrite concentration reveals an exponential tendency and the methemoglobin concentration reached a plateau. This characteristic suggests an equilibrium of methemoglobin formation by the reductase system activity (Huey et al, 1980).

The exposure of matrinxã to nitrite did not change the ion concentrations. In the marine teleost Lates calcarifer the enhancement of plasma sodium and chloride has been reported (Woo and Chui, 1997) and such fact should be associated to environmental seawater. Plasma potassium concentration is proposed to be associated to nitrite uptake. The increase of plasma K+ in Cyprinus carpio exposed to NO2

- is reported, and the direct correlation for both ions leads to such assumption (Jensen, 1990). In matrinxã, the plasma concentration of pattern nitrite followed the environmental one but the plasma level of potassium remained constant indicating that hipercalemia did not occur in matrinxã. The Cl- concentration did not change in the most of the freshwater teleosts fish. However, nitrite uptake by chloride cell in gills occurs by competitive interaction between Cl- and nitrite with the uptake sites. Gaino and col. (1984) suggest that there is no decrease of Cl- concentration because the hipertrophia of some gill chloride cells. Other exchange mechanism may be an hyperactive response working jointly to chloride cells to maintain the physiological Cl- levels, despite the nitrite competition or decrease of HCO3

-production. This process could result in degeneration in these cells. That author showed that nitrite inhibits the carbonic anhydrase of the gills (Cl- exchange with HCO3

- in gills) in vitro.

The present data call attention to the fact that the anti-oxidative mechanism to prevent the hemoglobin conversion to methemoglobin in the freshwater teleost

22

matrinxã exposed to nitrite is not enough efficient. No other mechanism to prevent nitrite deleterious effects seems to work in matrinxã, since the external and the plasma concentration of nitrite was practically the same. Those fact plus the osmotic disturbs in matrinxã, are cumulative and certainly responsible for the great sensibility of the species to the nitrite poisoning.

References

APHA 1980 Standard methods for examination of water and wastes. 12. ed., Washington, DC: Join Editorial board.

Arillo, A., Gaino, E., Margiocco, C., Mensi, P. Schenome, G. 1984 Biochemical

And Ultrastructural Effects Of Nitrite In Rainbow Trout: Liver Hypoxia As Of The Acute Toxicity Mechanism. Environ. Res. 34:135-154.

Bath, R.N., Eddy, F.B. 1980 Transport Of Nitrite Across Fish Gills. J. Exp.

Zool. 214:119-121. Beuteler, E. 1984 Red Cell Metabolism: Manual Of Biochemical Methods. 3.

Ed., Grune & Stratton, Inc, 187p. Cameron, J.N. 1971 Methemoglobin In Erythrcytes Of Rainbow Trout. Comp.

Biochem. Physiol., 40a:743-749. Doblander, C., Lackner, R. 1996 Metabolism And Detoxification Of Nitrite By

Trout Hepatocytes. Biochimica Et Biophysica Acta, 1289:270-274. Doblander, C., Lackner, R. 1997 Oxidation Of Nitrite To Nitrate In Isolated

Erytrocytes: A Possible Mechanism For Adaptation To Environmental Nitrite. An. J. Fish. Aquat. Sci., 54:157-161.

Eddy, F.B., Kunzilik, P.A., Bath, R.N. 1983 Uptake And Loss Of Nitrite From

The Blood Of Rainbow Trout, Salmo Gairneri Richardson, And Atlantic Salmon, Salmo Salar L. In Fresh Water And Dilute Sea Water. J. Fish Biol., 23:105-116.

Gaino, E., Arillo, A., Mensi, P. 1984 Involvement Of The Gill Chloride Cells Of

Trout Under Acute Nitrite Intoxication. Comp, Biochem. Physiol. A., 77(4):611-617.

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Hagopian, D.S., Riley, J.G. 1998 A Closer Look At The Bacteriology Of

Nitrification. Aquacultural Engineering, 18:223-244. Hargreaves, J. A. 1998 Nitrogen Biogeochemistry Of Aquaculture Ponds.

Aquaculture, 166:181-212. Heckman, C.W., Campos, J.L.E., Hardoim, E.L. 1997 Nitrite Concentration In

Well Water From Poconé, Mato Grosso, And Its Relationship To Public Health In Rural Brazil. Bull. Environ. Contam. Toxicol., 58:8-15.

Hilmy, A.M., El-Domiaty, N.A., Weshana, K. 1987 Acute And Chronic

Toxicity Of Nitrite To Clarias Lazera. Comp. Biochem. Physiol. C., 86(2):247-253.

Huey, D.W., Simco, B.A., Criswell, D.W. 1980 Nitrite –Induced

Methemoglobin Formation In Channel Catfish Transaction Of The American Fisheries Society 109:558-562.

Huey, D. W., Beitinger, T.L. 1981 A Methemoglobin Reductase System In

Channel Catfish Ictalurus Punctatus. Can. J. Zool., 60:1511-1513. Jensen, F.B., Andersen, N. A., Heisler, N. 1987 Effects Of Nitrite Exposure On

Blood Respiratory Proprieties, Acid-Base And Electrolyte Regulation In Carp (Cyprinus Carpio). J. Comp.Physiol. 157b:533-542.

Jensen, F.B. 1990 Nitrite And Red Cell Function I Carp: Control Factors For

Nitrite Entry, Menbrane Potassium Ion Permeation, Oxygen Affinity And Methaemoglobin Formtion. J. Exp. Biol., 152:149-166.

Jensen, F.B. 1992 Influence Of Haemoglobin Conformation, Nitrite And

Eicosanoids On K+ Transport Acroos The Carp Red Blood Cell Membrane. J. Exp. Biol., 171:349-371.

Knudsen,P.K., Jensen, F.B. 1997 Recovery From Nitrite-Induced

Methaemoglobinaemia And Potassium Balance Disturbances In Carp. Fish Physiol. Biochem., 16(1):1-10.

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Matsuoka, T. 1997 Determination Of Methemoglobin And Carboxyhemoglobin In Blood By Rapid Colorimetry. Biol. Pharm. Bull., 20(11):1208-1211.

Nikinmaa, M. 1992 How Does Environmental Pollution Affect Red Cell

Function In Fish? Aquatic Toxicology, 22:227-238. Peterson, M.S. 1990 Hypoxia –Induced Physiological Changes In Two

Mangrove Swamp Fishes: Sheepshead Minnow, Cyprinodon Variegatus Lacepede And Sailfin Molly, Poecilia Latipinna (Lesuer). Comp. Biochem. Physiol. A, 97(1):17-21.

Scarano, G., Saglia,M.G. 1984 Recovery Of Fish From Functional And

Haemolytic Anaemia After Brief Exposure To A Lethal Concentration Of Nitrite. Aquaculture, 43:421-426.

Schoore, J.E., Simco, B.A., Davis, K.B. 1995 Responses Of Blue Catfish And

Channel Catfish To Environmental Nitrite. J. Aquatic Animal Health, 7:304-311.

Scott, E.M., Harrington, J.P. 1985 Methemoglobin Reductase Activity In Fish Erytrocytes. Comp. Biochem. Physiol. B, .82(3):511-513.

Shechter, H. Gruener, N. Shuval, I. 1972 A Micromethod For The

Determination Of Nitrite In Blood. Anal. Chim. Acta, 60:93-99. Swift, D.J. 1981 Changes In Selected Blood Component Concentratios Of

Raibow Trout, Salmo Gardneri, Exposed To Hypoxia Or Sulethal Concentration Of Phenol Or Ammonia. J. Fish Biol. 19:45-61.

Tucker, C.S., Francis-Floyd, R., Beleau, M.H. 1989 Nitrite – Induced

Anemia In Channel Catfish, Ictalurus Punctatus Rafinesque. Bull. Environ. Contam. Toxicol., 43:295-301.

Vedel, N.E., Korsgaard, B., Jensen, F.B. 1998 Isolated And Combined Exposure

To Ammonia And Nitrite In Rainbow Trout (Oncorrhynchus Mykiss): Effects On Electrolyte Status, Blood Respiratory Properties And Brain Glutamine/Glutamate Concentrations. Aquatic Toxicology, 41:325-342.

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Williams, E.M., Eddy, F.B. 1986 Cloride Uptake In Freshwater Teleosts And Its Relationship To Nitrite Uptake And Toxicity. J.Comp. Physiol. B.156:867-872.

Willians, E.M., Glass, M.L., Heisler, N. 1993 Effects Of Nitrite-Induced

Methaemoglobinaemia On Oxygen Affinity Of Carp Blood. Enviorn. Biol. Fishes.37:407-413.

Woo, N.Y.S. Chiu, S.F. 1997 Metabolic And Osmoregulatory Responses Of The

Sea Bass Lates Calcarifer To Nitrite Exposure. Environ. Toxicol. And Water Qual., 12(3):257-264.

DNA SYNTHESIS AFTER EXPOSURE TO HEAVY METALS IN THE

TESTIS OF THE SPINY DOGFISH (SQUALUS ACANTHIAS)

J. Michael Redding Department of Biology, Tennessee Technological University

Cookeville, TN 38501 USA 931-372-3135/931-372-6257/mredding@tntech.edu

EXTENDED ABSTRACT ONLY – DO NOT CITE

Contamination of aquatic environments with metal compounds poses a serious risk to the health of aquatic species and terrestial species that rely on food from aquatic environments. Widespread metal contamination of both marine and freshwater systems has been reported, and a large body of literature documents deleterious effects of such pollution on various species. Of particular concern is the tendency of animals, especially carnivores, to accumulate metals from dietary sources, thereby increasing their risk for dose-dependent toxic effects. The spiny dogfish, Squalus acanthias, and other elasmobranchs are especially susceptible to such cumulative effects because they are long-lived, carnivorous animals whose home range includes coastal marine habitats which tend to have the highest concentrations of metals. Several studies have documented high metal concentrations in tissues of the spiny dogfish and other members of the genus Squalus (e.g., Taguchi, 1979). Metal intoxication may be directly detrimental to the health of the dogfish populations by decreasing survival or reproductive fitness. The known toxic effects of metals are diverse. Effects are thought to be exerted via the formation of stable complexes with many different biological molecules including proteins, DNA, RNA, and phosphorylated compounds. Specific mechanisms of metal toxicity have been characterized extensively for mammalian systems, but much less is known about non-mammalian vertebrates. Information on the specific cellular effects of metals on male reproductive systems in vertebrates is sparse but suggestive of profound disturbances (e.g., Clarkson, et al., 1983). The dogfish testis has proven to be an excellent model for studying the regulation of vertebrate spermatogenesis (DuBois and Callard, 1991). Distinct developmental stages of spermatocysts (germ cell:Sertoli cell units) can be

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isolated and cultured in vitro for at least two weeks. Moreover, mitotic activity, as indicated by DNA synthesis, is maintained quantitatively during this period and is responsive to stimulatory and inhibitory factors (Piferrer, et al., 1993). Thus, this model system would seem suitable for toxicological studies of vertebrate spermatogenesis. The purpose of this study was to determine the effects of various metals on DNA synthesis in the testis of the spiny dogfish. For each experiment, spermatocysts were isolated from zone I tissue from testes of 2-4 sharks and maintained in culture with Leibovitz L-15 medium, modified for use with elasmobranch tissue (DuBois and Callard, 1991). After various periods of treatment with metals, 5.0 uCi/ml of 3H-thymidine was added to the cyst cultures for 6-24 hr before harvesting the cysts. Harvested cysts were washed twice with saline solution augmented with excess unlabelled thymidine. Then, cysts were treated with ice-cold 10% trichloroacetic acid for 1-24 hr. Cysts were then washed again before solubilizing overnight in 0.2 M NaOH. Aliquots of the solubilized cysts were analyzed for radioactivity (cpm). Data were not standardized by sample protein concentration as previously reported; in these experiments such standardization would not significantly change the results. Results from cysts treated with metals were standardized as a percentage of the mean of untreated controls. The standardized means of treatment groups were compared to that of untreated control groups by an unpaired t-test with a pooled variance estimate. Preliminary experiments showed that mercuric chloride (HgCl2) at concentrations greater than 100 uM inhibited synthesis of DNA. Subsequently, the effects of equivalent concentrations (100, 500, 1000 uM) of HgCl2, the organic mercurial parachloro-mercuric-phenol-sulfonic acid (PCMBS), sodium vanadate (NaVO3), zinc chloride (ZnCl2), and cadmium chloride (CdCl2) were evaluated. Effects of these metals were compared to untreated controls, positive controls treated with bovine insulin (10 ug/ml), and negative controls treated with isobutyl-methylxanthine (1 mM, IBMX). Results of this experiment are shown in Table 1. Mercurial compounds showed dose-dependent inhibition of DNA synthesis. Of these HgCl2 was most potent, virtually negating DNA synthesis between 100 and 500 uM. Cadmium stimulated DNA synthesis at 100 uM but markedly inhibited it at 500 and 1000 uM. Vanadate was relatively ineffective, reducing DNA synthesis only to 71% of controls at 1000 uM. In contrast, zinc slightly stimulated DNA synthesis at all concentrations. Beyond simply identifying their effective concentrations, these results demonstrate the specificity of

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various metals with respect to their effects on DNA synthesis in shark testis. This specificity may reflect differences in the capacity and affinity of metal binding proteins, such as metallothionine, that effectively sequester metals and prevent them from affecting critical cellular processes. It is evident from these results that DNA synthesis in the Squalus testis is sensitive to metal intoxication, generally supporting previous results from mammalian models (Clarkson, et al., 1983). These results support the use of Squalus testis as a model system for toxicological studies of vertebrate spermatogenesis. [This research was supported by a fellowship from the Lucille P. Markey Charitable Trust via the Mount Desert Island Biological Laboratory. I thank G.V. Callard, D.S. Miller, D. Barnes and A. Kleinzeller for valuable advice and material support.] Table 1. DNA synthesis rates of Squalus zone I spermatocysts cultured in vitro with metals, insulin (10 ug/ml, positive control), or IBMX (1 mM, negative control) for 24 hr and exposed for the last 12 hr to radiolabelled thymidine. Results are shown as a mean (SE) percentage of control cultures. Sample size was four for each treatment and eight for the untreated control. All means except those noted by "ns" were significantly (P< 0.01) different from the control. Metal Concentration (uM) Treatment 0 100 500 1000 Control 100 (2) --- --- --- Insulin 192 (7) --- --- --- IBMX 34 (3) --- --- --- HgCl --- 84 (2) 3 (1) 0 (0) 2PCMBS --- 95 (3)ns 62 (3) 8 (3) CdCl2 --- 128 (6) 69 (2) 21 (3) NaVO3 --- 98 (3)ns --- 71 (4) ZnCl 2 --- 117 (6) 117 (6) 127(12)

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References Clarkson, T.W., Nordberg, G.F., and Sager, P.R. 1983. Reproductive and developmental toxicity of metals, Plenum Press, New York. DuBois, W. and Callard, G.V. 1991. Culture of intact Sertoli/germ cell units

and isolated Sertoli cells from Squalus testis: I Evidence of stage-relate functions in vitro. J. Exp. Zool. 258:359-372.

Piferrer, F., Redding, M., DuBois, W., and Callard, G. 1993. Stimulatory and inhibitory regulation of DNA synthesis during spermatogenesis:

studies in Squalus acanthias. Taguchi, M., Yasuda, K., Toda, S., and Shimizu, M. 1979. Study of metal contents of elasmobranch fishes: Part I – Metal concentration in the

muscle tissues of a dogfish, Squalus mitsukurii. Mar. Environ. Res. 2: 239-249.

QUANTITATIVE PCR ANALYSIS OF ATLANTIC SALMON

CYP1A IN GILLS

Christopher B. Rees Department of Fisheries and Wildlife

Michigan State University 13 Natural Resources Building

East Lansing, MI 48823 Phone: 517-432-1141

Fax: 517-432-1699 Email: reeschri@msu.edu

Stephen D. McCormick

USGS, Leetown Science Center, Conte Anadromous Fish Research Center

One Migratory Way, PO Box 796 Turners Falls, MA 01376

Phone: 413 863-3804 Fax: 413 863-9810

Email: mccormick@umext.umass.edu

Weiming Li Department of Fisheries and Wildlife

Michigan State University 13 Natural Resources Building

East Lansing, MI 48823 Phone: 517-432-1141

Fax: 517-432-1699 Email: liweim@msu.edu

EXTENDED ABSTRACT ONLY - DO NOT CITE

Environmental pollutants are hypothesized to be one of the causes of recent declines in wild populations of Atlantic salmon (Salmo salar) across Eastern Canada and the United States. Some of these pollutants, such as polychlorinated biphenyls and dioxins, are known to induce expression of the CYP1A subfamily of genes. We applied a highly sensitive technique, quantitative reverse transcription-polymerase chain reaction (Q-RT-PCR), to measure CYP1A induction and expression patterns in Atlantic salmon gills. This assay was used to detect patterns

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of CYP1A mRNA levels, a direct measure of CYP1A expression, in Atlantic salmon exposed to pollutants under both laboratory and field conditions. Two groups of Atlantic salmon juveniles (48-76 g) received an intraperitoneal injection of 50 µg g-1 β-naphthoflavone (BNF) or vehicle only (n=10 for both groups). Non-lethal gill biopsies were taken for each treatment group prior to injection and 1,2, and 7 days post-injection. The same fish were used at each sampling point. After RNA extraction and Q-RT-PCR analysis, control fish showed static levels of CYP1A over the course of sampling. Induced salmon demonstrated similar levels of CYP1A to control fish at time zero and a significant induction over the course of each additional sampling period. The quantitative RT-PCR was also applied to salmon sampled from two streams (n=10 for each stream) in Massachusetts, USA. Salmon gill biopsies sampled from Millers River (South Royalston, Worcester County), known to contain polychlorinated biphenyls (PCBs), showed a significant induction over those collected from a stream with no known contamination (Fourmile Brook, Northfield, Franklin County). Gill biopsies coupled with Q-RT-PCR analysis is a novel, sensitive, and accurate method to estimate CYP1A expression dynamics in gill tissues of Atlantic salmon. This method has the potential to be a valuable tool in environmental assessment of not only wild Atlantic salmon populations, but many other populations of salmonids as well. Acknowledgements This work is supported by the National Marine Fisheries Service and the Great Lakes Fishery Commission.

EXPRESSION OF MULTI-DRUG RESISTANCE (MDR)

P-GLYCOPROTEIN AND ASSOCIATED PROTEINS (MRPs) IN

THE FLOUNDER; EFFECTS OF EXPOSURE TO THE

POLYAROMATIC HYDROCARBON, BENZO(a)PYRENE.

Gordon Cramb School of Biology, University of St Andrews, Bute Medical Buildings,

St Andrews, Fife, Scotland, KY16 9TS. Email gc@st-and.ac.uk; Tel. 044/0 1334 3530; Fax 044/0 1334 3600.

(1)Christopher P. Cutler, (1)Jean-H. Lignot, (1)Anne-Sophie Martinez, (2)Belinda S. Chesman, (2)Susan C. Frankling and (2)J.Anne Brown,

(1) School of Biology, University of St Andrews, Bute Medical Buildings, St Andrews, Fife, Scotland, KY16 9TS

(2) School of Biological Sciences, University of Exeter, Hatherley Laboratories, Exeter, EX4 4PS

EXTENDED ABSTRACT ONLY - DO NOT CITE Introduction Aquatic organisms are frequently faced with periods of exposure to various environmental pollutants, often as the result of the release of chemicals from agricultural and/or industrial activities. At present, a number of potential biomarkers have been characterised in terms of their sensitivity to a wide range of natural and anthropogenic xenobiotics. In addition to these classic biomarkers, there is increasing evidence to suggest that certain members of the ATP Binding Cassette (ABC) transporter superfamily, including the multi-drug resistance (MDR)/p-glycoprotein and multi-drug resistance associated protein (MRP) sub-gene families, are also involved in the survival of marine organisms exposed to various environmental pollutants. The products of these genes are active transporters that couple the hydrolysis of ATP with the epithelial transport of organic cations and anions. The normal physiological role for these proteins is in the transport of bile salts and organic ions within the liver and intestinal parenchyma

33

(Meier and Stieger, 2002), however as a result of their wide substrate specificity these complex membrane proteins can also transport, and therefore aid in the excretion of, a wide range of environmental toxins. As a result, the products of these gene families have been implicated in the development of cellular resistance to various environmental xenobiotics. To date, very little information is available about the regulation of expression and function of the MDR/p-glycoprotein/MRP families in marine teleosts and how these genes may confer protective resistance following exposure to various environmental pollutants. In this study we report the cloning, sequencing and expression of several members of the MDR and MRP gene families from the European flounder (Pleuronectes flesus). The levels of expression of these gene products were monitored after exposure of fish to the polyaromatic hydrocarbon, benzo(a)pyrene (B(a)P). Materials and methods Juvenile flounder were maintained in normal seawater or in seawater containing 0.5ppm B(a)P for up to 6 days before removal of tissues for RNA extraction (Cutler et al., 2000), plasma membrane preparation (McCartney and Cramb, 1993) and sampling of bile for analysis of B(a)P metabolites by fluorescence spectroscopy. Cloning of the MDR and MRP cDNAs was conducted using Invitrogen's TA Cloning Kit (Leek, The Netherlands) following amplification of fragments using degenerate primers in RT-PCR and then 3'- and 5'-RACE specific primers with Marathon cDNA (Clontech, Basingstoke, UK). Fragments were sequenced using a Big Dye Terminator Cycle Sequencing kit (Perkin Elmer Biosystems, Warrington, UK) and radiolabelled (Amersham Pharmacia Biotech, Megaprime labelling kit; Bucks, UK) for use in Northern blots. Purified plasma membrane fractions were run on denaturing SDS PAGE gels, blotted onto PVDF membranes and processed for Western blotting using specific antisera raised to two isoforms of flounder MDR. Liver and gut tissues were fixed in 4% paraformaldehyde and processed for immunohistochemistry. Sections were probed with MDR isoform-specific antisera and then incubated with FITC-conjugated donkey anti-rabbit IgG before visualisation on a fluoresecent microscope. Results and Discussion RT-PCR and Marathon RACE techniques amplified full-length cDNA

34

clones for two members of the MDR family (termed MDR-A and MDR-B) from the European flounder. Comparison of sequence homologies between the teleost MDRs and the genebank data bases indicated that MDR-A is a member of the bile transporter sub-family that is known as “sister of p-glycoprotein” or abbreviated to sp-gp. MDR-B amino acid sequences are more consistent with the multi-drug resistance/p-glycoprotein family, however due to heterogeneity within the sequences it is not possible to assign MDR-B as the homologue of any one member of the mammalian MDR family. In addition to these two members of the flounder MDR gene family, cDNA fragments of five members of the related multi-drug resistance associated protein (MRP) family were amplified and cloned. These cDNA fragments have subsequently been characterised, based on sequence comparisons) as homologues of mammalian MPRs 1 to 5. Although expressed in most tissues at low levels MDR and MRP isoforms exhibited differential tissue expression with the liver being the major site of expression of MDR-A, the intestine and brain for MDR-B and mixed expression profiles for the MRP isoforms (Fig. 1). Expression profiles were not affected to any major degree by exposure of fish to B(a)P. Highly purified plasma membrane fractions were isolated from both the liver and the intestine and samples run on denaturing SDS-PAGE for Western blotting. Multiple immunoreactive bands ranging in size from 100kDa to >250 kDa were apparent in both tissues and with antisera raised to both MDR-A and MDR-B with no marked differences between control and B(a)P-treated fish (see Fig 2). Immunohistochemistry revealed that, as expected, the bile canniculae were the main site of expression of both MDR-

35

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A and MDR-B isoforms in the liver. In the intestine both proteins were expressed within in the apical membrane of the intestinal epithelia although MDR-A was also found in duodenal gland-like structures which lie under the absorptive cell layer. Acknowledgements This study was funded by a project grant from the Natural Environment Research Council - GR3/12467.

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References Cutler, C.P., Brezillon, S., Bekir, S., Hazon, N. and Cramb, G 2000.

Expression of a duplicate Na, K-ATPase β1 isoform in the European eel (Anguilla anguilla) Am.J.Physiol. 279: R222-R229.

McCartney, S. and Cramb, G. 1993. Effects of a high salt diet on hepatic

atrial natriuretic peptide receptor expression in Dahl salt-sensitive and salt-resistant rats. J.Hypertens. 11: 253-262.

Meier, P.J. and Stieger, B 2002. Bile salt transporters. Ann. Rev. Physiol.

64: 635-661

SEDIMENT-ASSOCIATED POLLUTANTS IN THE MARINE

ENVIRONMENT: A MULTI-BIOMARKER APPROACH FOR

ASSESSING SEDIMENT TOXICITY IN TURBOT

Kilemade, M., Hartl, M. G. J., O’Halloran, J. Department of Zoology and Animal Ecology

University College Cork, Ireland Telephone: +353214904594, fax: +353214904593; e-mail: m.hartl@ucc.ie

Sheehan, D.

Department of Biochemistry University College Cork, Ireland

van Pelt, F. N. A. M.

Department of Pharmacology and Therapeutics University College Cork

O’Brien, N. M.

Department of Food Science, Food Technology and Nutrition University College Cork, Ireland

EXTENDED ABSTRACT ONLY- DO NOT CITE

Sediments in the aquatic environment have become an area of concern due to their potential for accumulating toxic compounds and acting as a secondary pollutant source to benthic fauna. Owing to their predominantly benthic life style, fish of the order Pleuronectiformes (flatfish) are particularly vulnerable to sediment-associated pollutants. This and the relative ease of obtaining specimens, from either commercial hatcheries or local estuaries, makes them the preferred choice for studying sediment-water-organism interactions in benthic fish (Courtney et al., 1980; Hartl et al., 2001). Here we report on the first phase of an ongoing project applying a multi-biomarker approach to the toxicity of field-collected sediments from a polluted estuary to juvenile turbot (Scrophthalmus maximus, L.). Laboratory experiments using fish exposed to spiked sediments have been instrumental in establishing

39

biomarkers for single compounds. The aim of the present study was to determine a suite of biomarkers, in combination with chemical and statistical analysis, capable of establishing cause and effect relationships of exposure to sediments containing complex mixtures of pollutants. Sediments were sampled at low tide from two sites, Whitegate and Aghada, in Cork Harbour, Co. Cork, Ireland, where sediments have previously been shown to contain elevated levels of organic pollutants, particularly polychlorinated biphenyls and organotin compounds (Boelans et al., 1999) and from a clean reference site at Ballymacoda, Co. Cork, Ireland (Byrne and O'Halloran., 2000). The top layer of the sediments were collected, thoroughly mixed and stored at 4°C over night. Subsamples were frozen (-20ºC) prior to chemical analysis. A layer (approx. 5 cm) of sediment was applied to 500 litre aquaculture tanks filled with aerated seawater (S = 35; 15ºC). Following 7 days acclimation, 60 turbot (30g ± 5) were added to each tank and exposed for 21 days. Individuals were sampled at 0, 7, 14 and 21 days and sacrificed. Blood samples were taken from the caudal vein for the analysis of blood osmolality, haematocrit, differential cell counts, serum protein and DNA single strand breaks. A Comet assay, for the analysis of DNA single strand breaks, was also performed on skin, gill, spleen and head kidney cell preparations. The liver and two gill arches from both the upper and the lower gill pouch were removed, shock-frozen in liquid nitrogen and stored at –70ºC until further analysis of P450 induction (measured as EROD activity in hepatic S9 post-mitochondrial fractions) and membrane-bound Na+/K+-ATPase activity, respectively. Results from preliminary experiments showed that turbot exposed to contaminated sediments displayed an increase in DNA single strand breaks in gill cells, haemocytes and haemopoietic tissues when compared to those exposed to sediments from the clean site. There was also an increase in blood osmolality in fish exposed to the polluted sediments, indicating an increase in membrane permeability, due to the possible interaction of lipophillic organic compounds with gill epithelia, and the resulting osmotic loss of water across the membrane. The blood parameters, haematocrit, differential cell counts and serum protein, remained unchanged during the exposure period in all treatments. Although sediments spiked with a single compound have aided the understanding of toxicological mechanisms, they generally have limited environmental relevance, in particular by disregarding potential synergistic

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and/or antagonistic pollutant effects. By using an array of relevant biomarkers combined with chemical and statistical (e.g. Principal Component Analysis) analysis, we are currently assessing the toxicological effects of field-collected sediments from Cork Harbour and the principle pollutants involved. References Boelans, R. G. V., Maloney, D. M., Parsons, A. P. & Walsh, A. R. (1999).

Ireland's Marine and Coastal Areas and Adjacent Seas - An Environmental Assessment Marine Istitute, Dublin: 388.

Byrne, P. A. & O'Halloran, J. (2000). Acute and sublethal toxicity of estuarine

sediments to the manila clam, Tapes semidecussatus. Environ. Toxcol. 15: 456-468.

Courtney, W. A. M. & Langston, W. J. (1980). Accumulation of polychlorinated

biphenyls in turbot (Scrophthalmus maximus) from sea water sediments and food. Helgol. Meeresunters. 33: 333-339.

Hartl, M. G. J., Hutchinson, S. & Hawkins, L. E. (2001). Sediment-associated

tri-n-butyltin chloride and its effects on osmoregulation of freshwater-adapted 0-group European flounder, Platichthys flesus (L.). Aquat. Toxicol. 55: 125-136.

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RESPONSES OF ATLANTIC SALMON

AND BIVALVE MOLLUSCS TO

PARALYTIC SHELLFISH

F B Eddy Department of Biological Sciences

University of Dundee, Dundee, DD1 4HN, Scotland, UK. f.b.eddy@dundee.ac.uk

Matt J Gubbins and Susan Gallacher

FRS Marine Laboratory, PO Box 101, Victoria Road Aberdeen AB11 9DB, Scotland, UK

EXTENDED ABSTRACT ONLY – DO NOT CITE Paralytic shellfish poisoning (PSP) toxins are a group of potent neurotoxins produced by certain strains of marine dinoflagellates. Blooms of these algal species can result in the passage of PSPs through marine food webs, with detrimental effects on the marine environment and human health. These toxins have been implicated as the causative agent in some of the many fish kills that have occurred during blooms of PSP producing dinoflagellates. As such, PSPs represent a potential threat to fisheries resources and aquaculture. Little is known of the fate of these compounds in fish, but they have been shown to accumulate in the liver of mackerel sampled after bloom events. Analysis of fish tissues also suggests that transformation of these toxins does occur following absorption since the PSP analogues differ from those to which the fish were exposed. The induction of xenobiotic metabolising enzyme activities has also been noted in Atlantic salmon exposed to PSPs and may represent a detoxification mechanism. Better understood is the fate of PSPs in shellfish. Bivalve molluscs accumulate significant of PSPs during toxic blooms and pose a serious heath risk to vertebrate consumers. Many species of marine invertebrates are reported to be resistant to the effects of neurotoxins, a trait that is innate in some species and acquired in others. There are reports of PSP biotransformations in shellfish,

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namely epimerisations, decarbamoylations and reductive elimination’s that alter the profile of the toxins as they are transferred from causative dinoflagellate to shellfish tissue. There are also reports of protiens, inducible on exposure to PSPs that can bind to saxitoxin. Such a mechanism may be responsible for inactivating PSPS and could confer resistance to invertebrates that express these proteins. Little is known about the sub-lethal effects of exposure from this group of toxins on marine organisms. Laboratory based exposure experiments on Atlantic salmon (Salmo salar) indicate that intra-peritoneal exposure to low doses (2-4 µg/kg) of saxitoxin causes an induction of hepatic glutathione S-transferase (GST) activity within four days. Doses approximating the LD50 for this compound (4 µg/kg) had little effect on blood plasma ionic concentration of surviving fish. Mussels (Mytilus edulis), like other invertebrates, appear insensitive to the paralytic effects of PSTs. Exposure to high doses (intra-muscular, >100 µg/100g soft tissue) of saxitoxin, however, causes an induction of digestive gland GST activity. This is in contrast to scallops (Pecten maximus) which showed no induction of GST activity after acquiring high digestive gland toxicities from feeding on cultures of toxic dinoflagellates. After toxic events, scallops retain PSTs considerably longer than mussels. It is suggested that the induction response of GST in mussels may be partly responsible for this discrepancy in toxicokinetics between the two species. In conclusion further work is required to define the metabolic pathways leading to the detoxification and excretion of saxitoxins and related compounds. References MJ Gubbins, FB Eddy, S Gallacher and RM Stagg (2000). Paralytic shellfish

poisoning toxins induce xenobiotic metabolising enzymes in Atlantic salmon (Salmo salar). Marine Environmental Research 50, 469-483.

M J Gubbins, EA Guezennec, F B Eddy, S Gallacher & RM Stagg,. (2001).

Paralytic shellfish toxins and glutathione S-transferases in artificially intoxicated marine organisms. GM Hallegraeff, SI Blackburn, CJ Bolch & RJ Lewis (eds), Harmful Algal Blooms 2000. Intergovernmental Oceanographic Commission of UNESCO. Paris, 2001. 387 - 390

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Fig 1. Hepatic glutathione S-transferase (GST) activities of Atlantic salmon

administered intra-peritoneal injections of physiological saline, saxitoxin and extraxts of a toxic (Alexandrium fundyense) and a non-toxic (Scrippsiella trochoidea) strain of dinoflagellate over 21 days. Error bars represent + or – SEM, n=10 (except for saxitoxin group where n=9). Ab, groups with different notation are significantly different (P<0.05)

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THE EFFECT OF ENVIRONMENTAL LEVELS

OF FRESHWATER CONTAMINANTS

ON JUVENILE ATLANTIC SALMON (Salmo salar L.):

IMPLICATIONS FOR MARINE SURVIVAL

Nicola Lower Centre for Environment, Fisheries and Aquaculture Science,

Lowestoft Laboratory, Pakefield Road, Suffolk, NR33 0HT, UK Tel: +44 (0) 1502 562244 Fax: +44 (0) 1502 513865

E-mail: N.J.Lower@cefas.co.uk

Andy Moore Centre for Environment, Fisheries and Aquaculture Science,

Lowestoft Laboratory, Pakefield,Road, Suffolk, NR33 0HT, UK Tel: +44 (0) 1502 562244 Fax: +44 (0) 1502 513865

E-mail: A.Moore@cefas.co.uk

EXTENDED ABSTRACT ONLY – DO NOT CITE There is increasing concern over the continuing decline of wild stocks of Atlantic salmon, Salmo salar, throughout the North Atlantic, and the impact on commercial and recreational fisheries. Recent research has demonstrated that the freshwater and marine environments cannot be considered in isolation and that conditions within the freshwater zone experienced by Atlantic salmon may be critical to their subsequent survival in the sea. In particular, exposure of juvenile salmon to a range of sub-lethal concentrations of freshwater contaminants, such as pesticides, may operate to reduce survival in fish once they have migrated to sea. Environmental levels of a wide range of pesticides have previously been shown to deleteriously effect Atlantic salmon reproduction and fecundity, by disrupting pheromone-mediated spawning and reducing fertilisation rates (Moore and Lower, 2001; Moore and Waring, 2001). However, the effects of such exposure on other critical life history stages, for example developing embryos and the parr-smolt transformation, and the subsequent survival of salmon in the sea is less clear. The aim of the present studies was therefore to determine the impacts of freshwater contaminants on smolt physiology, behaviour and marine survival and to further investigate the sub-lethal effects of contaminants on reproduction and the emergence of juvenile salmonids.

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Two types of persistent freshwater contaminants were studied: flame retardants and pesticides derived from intensive agricultural. Flame retardants are substances used in the manufacture of a wide range of materials such as plastics and textiles. These chemicals enter the aquatic environment as point sources either through leaching from landfills or from effluents derived from the manufacturing process. The majority of flame retardants contain brominated organic compounds, making them persistent and lipophilic with the ability to bioaccumulate. Pentabromodiphenyl ether (PeBDE), tetrabromobisphenyl-A (TBBPA) and hexabromocyclododecane (HBCD) are the most frequently used brominated flame retardants, and they are all found at low levels in many European rivers (de Wit, 2002). The widespread usage of pesticides in agriculture has resulted in the extensive contamination of many rivers and tributaries supporting salmon populations. Three such pesticides are cypermethrin, a synthetic pyrethroid insecticide primarily used in crop sprays and in sheep dips; diazinon, an organophosphate insecticide; and atrazine, a pre- and post-emergence herbicide used in the control of annual and perennial grass and broad-leaved weeds. In a series of separate studies, pre-smolts (Salmo salar) were continuously exposed to either environmental levels of flame retardants (PeBDE, TBBPA and HBCD) or the pesticides (cypermethrin, diazinon and atrazine). Fish were maintained in a freshwater flow-through system and contaminants were introduced using a peristaltic pump. Dosing periods ranged from 10 to 14 days, or in the case of the atrazine study, a period of 2 months. A number of fish from each group were then sampled for some of the physiological parameters associated with the parr-smolt transformation (plasma ion levels [Na+, K+, Cl-], gill Na+ K+ ATPase activity and plasma thyroid levels [triiodothyronine, T3; thyroxine, T4]. The remaining fish in each group were subsequently exposed to a seawater challenge test (SWC) for 24 hours to assess osmotic capabilities and adaptation of the fish to saltwater. At the end of this test, all fish were again sampled and the above physiological parameters measured. In the atrazine study, fish were first PIT-tagged, exposed to the pesticide and subsequently released into an experimental stream channel in order to determine any effect of exposure on downstream migratory behaviour. Exposure of the smolts to the flame retardants (0.5, 5, 10, 50, 100ngl –1) had little or no effect on gill Na+ K+ ATPase activity or plasma Na+, K+, or Cl- levels. Thyroid hormone levels significantly increased in the control group following exposure to seawater, but exposure to 0.5ngl-1 PeBDE abolished this increase and both T3 and T4 plasma levels were significantly lower compared to the SWC control. All fish in the groups dosed with higher

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levels of PeBDE died during the SWC. Exposure to atrazine (0.5, 5µgl-1) also had no effect on the plasma ion levels in smolts, but the gill Na+ K+ ATPase activity was significantly lower in the fish dosed with 0.5µgl-1 atrazine compared to the SWC control. Current research is focused on modelling the impacts of environmental levels of both individual pesticides and mixtures on smolts, reproductively mature parr, and developing embryos to determine possible synergistic or additive effects. In one study, salmon eggs and milt were briefly exposed during fertilisation to pesticide-dosed water, before being incubated in artificial redds. Preliminary results indicate that exposure even at this stage of the salmonid life cycle can have implications for the production of quality juveniles, as both timing of emergence and mortality of fry were effected. In conclusion, exposure of environmental levels of waterborne contaminants within freshwater has been shown to disrupt a number of sensitive stages in the life cycle of the salmon. This has implications for the number of juvenile salmon recruited into the population, their subsequent survival in the marine environment and the numbers of returning adults. Acknowledgements This work was funded by DEFRA. The authors wish to thank Lorraine Greenwood for assistance with the gill ATPase assay. References De Wit, C.A. 2002. An overview of brominated flame retardants in the

environment. Chemosphere. 46: 583-624. Moore, A. and Lower, N. 2001. The impact of two pesticides on olfactory-

mediated endocrine function in mature male Atlantic salmon (Salmo salar L.) parr. Comparative Biochemistry and Physiology Part B. 29:269-276.

Moore, A. and Waring, C.P. 2001. The effects of a synthetic pyrethroid

pesticide on some aspects of reproduction in Atlantic salmon (Salmo salar L.). Aquatic Toxicology. 52 (1):1-12.

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50

CHLORIDE CELL CHANGES INDUCED BY NITRITE EXPOSURE

IN AN AMAZONIAN FISH SPECIES

Oscar Tadeu Ferreira da Costa Federal University of São Carlos

Post-Graduate Program in Ecology and Natural Resources C. Postal 676, Phone: 55 16 260-8314

Department of Morphology University of Amazonas

Marisa N. Fernandes

Federal University of São Carlos Department of Physiological Sciences C. Postal 676, Phone: 55 16 260-8314

email: dmnf@power.ufscar.br

Abstract The gill chloride cells of the juvenile Amazonian fish Colossoma macropomum were analyzed using light and scanning and transmission electron microscopy after 96h exposure to 0.04 and 0.2 mM nitrite (NO2

-). Although the number of chloride cells decreased significantly in the lamellar epithelium, no decrease was found in the interlamellar region of the gill filament. A positive dose-effect was evidenced by ultrastructural changes in chloride cells. NO2

- exposure caused significant reduction on the apical surface area of individual chloride cells (p < 0.05), with a resulting reduction of the fractional area of these cells in both the lamellar and filament epithelium. Swelling of endoplasmic reticulum cisternae, nuclear envelope and mitochondria were the main changes found in the chloride cells. Cristae lysis and matrix vacuolization characterized the mitochondrial changes. The overall ultrastructural changes in the chloride cell indicated cellular functional disruption caused by exposure to nitrite. Introduction Nitrite (NO2

-) is toxic to aquatic organisms. High NO2- levels may develop as a

result of an imbalance between the bacterial nitrification and denitrification processes of ammonia (Lewis and Morris, 1986). Some environmental conditions such as temperature and high levels of organic matter favor the

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occasional development of NO2

- in the Amazonian environment and in fish culture systems. The active mechanism for NO2

- uptake by the gill chloride cell and its possible toxic effect on these cells were discussed by Williams and Eddy (1986). However, most studies have investigated the toxic effects of NO2

- on hemoglobin, the P450 enzymatic system and mitochondrial cytochrome oxidase. Few studies have focused on the ultrastructural changes in gill chloride cells. The main purpose of the study reported on herein was to evaluate the effect of NO2

- on the number and ultrastructure of the chloride cells of an Amazonian serrasalmid species, Colossoma macropomum (Cuvier, 1818), popularly known as tambaqui. Materials and Methods Experimental animals Juvenile specimens of tambaqui, Colossoma macropomum, weighing 66 ± 3 g, were obtained from the Aquaculture Research and Training Center–CEPTA/IBAMA (Pirassununga, SP). The fish were kept for at least 4 weeks in 1000-L holding tanks supplied with running aerated water (at a temperature of 21oC; pH 7.4; water PO2 > 130 mmHg; Ca++, 1.71 mg/L, Na+, 0.78 mg/L, Cl-, 0.5 mg/L, NO2

- < 0.004 mM). The animals were fed daily to satiation with commercial fish food pellets. Feeding was suspended two days before the experiments. The photoperiod adopted was 12 h daylight and 12 h night during acclimation and throughout the experimental period. Nitrite exposure Six fish per treatment were placed in an experimental aquarium (63 L) supplied with continuous aerated water from the same source as the holding water, and acclimatized for 24 h prior to adding reagent grade sodium nitrite to provide the selected concentrations of 0.04 and 0.2 mM NO2

- (respectively 31% and 154% of 96 h LC50 NO2

- estimated for tambaqui (96h LC50 NO2- = 0.13 mM; Costa

and Fernandes, 2000). The flow through the experimental aquarium was adjusted to 1.0 L/min and the NO2

- exposure was maintained for 96 hours. The entire volume of water in each tank was completely renewed twice a day. The nitrite concentration was measured and adjusted according to Strickland and Parsons (1972). Similar tests, although without the addition of NO2

-, were used for control purposes. Following the 96 h exposure to NO2

-, the fish were removed from the test aquarium, anesthetized (benzocaine, 1:10,000), and sacrificed with a blow to the head.

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Sampling and tissue processing The second gill arch on the right-hand side was immediately excised and washed in a 0.9% saline-saccharose solution. For light microscopy (LM), filaments (n = 6) still attached to the septum of the arch were immersed in 4% paraformaldehyde, 1% glutaraldehyde in 0.1 M of sodium phosphate-buffer (pH 7.2) fixative solution for 24 h. The samples were dehydrated and embedded in Historesin LKB (Reichert-Jung). Serial sections (1 µm thick) from the trailing to the leading edge of the filament (longitudinal sections perpendicular to the surface of the secondary lamella) were stained with 0.4% toluidine blue and subsequent PAS (periodic acid-Schiff reaction, Kiernan, 1990) and analyzed under an Olympus CBA-K compound microscope equipped with a JVC video camera. For transmission electron microscopy, small pieces of gills (n = 4) were fixed in 2.5% glutaraldehyde solution buffered with 0.1 M sodium phosphate at pH 7.3 for 2 h. The samples were post-fixed in a sodium phosphate-buffered 1% OsO4 solution, dehydrated with an acetone series, and embedded in Araldite 6005 (Ladd Research). Ultra thin sections (60-70 nm) were contrasted with uranyl acetate and lead citrate and examined under a Philips CM 120 electron microscope at 100-kV accelerating voltage. For the analysis by scanning electron microscopy, pairs of filaments attached to the septum (n = 4) were fixed in 2.5% glutaraldehyde solution buffered with 0.1 M sodium phosphate at pH 7.3 for 2 h. The samples were dehydrated with a graded ethanol series up to pure ethanol, followed by ethanol/acetone up to pure acetone and CO2 critical point dried. Filament pairs were glued with silver paint onto the specimen stub, coated with gold in a vacuum sputter and examined under a DSM 940 ZEISS Scanning Microscope at 25 kV. Morphometry and histopathological analyses Morphometric analyses were performed on 1 µm serial sections stained with toluidine blue-PAS. Digitized images from each section were analyzed using SigmaScan 3.0 Image Analyser software (Jandel Scientific) and a Merz grid (Merz, 1967) superimposed over the video monitor to ensure random orientation of the measurements on each section with 1000 times magnification. Chloride cells were counted in five consecutive lamella and interlamellar regions (filament) of three different filaments per fish (30 measurements per fish). Each cell count and measurement was made using a randomized blind method in

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which the counter did not know whether the tissue sections were from nitrite-treated experimental or untreated control animals. A semi-quantitative analysis of the ultrastructural components of chloride cells (10,000x) and mitochondria (35,000x) was made with high magnification. Statistics All data were expressed as mean ± standard error of the mean. Statistical significance between data sets was determined by one-way ANOVA, followed by Bonferroni's multiple comparison tests. Regression analysis and correlation coefficients between the variables were also calculated. A statistical significance of p < 0.05 was adopted. Results The secondary lamellae of C. macropomum were regularly spaced on both sides of gill filament and consisted of two epithelial layers of cells kept apart by rows of pillar cells interspersed with blood vessels. The control fish of this species presented numerous chloride cells on the secondary lamella. NO2

- exposure produced drastic morphological changes in the gill. There was a gradual increase in the interlamellar distance from 18.42 ± 0.85 µm in control conditions to 21.94 ± 0.95 µm after exposure to 0.2 mM NO2

-. The several layers of undifferentiated filament cells were significantly reduced after exposure to NO2

-. Nitrite exposure reduced the chloride cell number in the lamellar epithelium by 51-57% (p < 0.05), but no significant changes were found in the number of these cells in the filament (interlamellar region) (Fig. 1). The total number of chloride cells (lamellar plus filament) was reduced by 49% and 45% (to 0.04 and 0.2 mM NO2

-, respectively) (Fig. 1).

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Control 0.04 mM 0.2 mM0

1

2

3

Control 0.04 mM 0.2 mM0

4

8

12

16

Control 0.04 mM 0.2 mM0

3

6

9Chloride cells from lamella Chloride cells from filament Total chloride cells

* * * *

A CB

Number of cells per five interlamellar spaces

Figure 1. Chloride cell count at the lamellae (A), filament (B) and total number ofchloride cells (lamellae plus filament). The number of chloride cells was only reduced inthe lamellae. * Significantly different (p < 0.05) from control. Data presented as mean ±standard error of the mean (n = 6).

Table 1 summarizes the cytological analysis of chloride cells of control C, macropomum and fish exposed to NO2

- for 96 h. The chloride cells were rich in mitochondria and presented a well-developed tubular network system similar to other freshwater teleosts (Laurent and Dunel, 1980). The lamella in control fish was lined with chloride cells (Fig. 2A). Chloride cells of the control group were round, with a basally located nucleus, which normally presented heterochromatin in the internal membrane of the nuclear envelope (Fig. 2A-C). Ellipsoid-shaped mitochondria were arranged predominantly near the apical region. Some cisternae of the tubular network were visible close to mitochondria (Fig. 2C). The smooth endoplasmic reticulum (SER) was extensive, through not dilated. Spherical vesicles were dispersed in the cytoplasm. The apical surface of chloride cells, in some cases, displayed a few short microvilli and, in others, was buried by pavement cells. The exposure to 0.04 mM NO2

- for 96 h of NO2- caused a decrease in the number

of chloride cells, resulting in the increased distance of the interlamellar spaces (Fig. 3A). Hypertrophied chloride cells were visible in the lamellae. Chloride cells exhibiting signs of degeneration in cellular compartments and nuclei displaying a slight swelling of the perinuclear space were common (Fig. 3C).

55

Table 1. Ultrastructural changes in the gill chloride cells of tambaqui, C.

macropomum, after 96 h exposure to NO2-. The data are the mean

values of semi-quantitative evaluations from four specimens per group. [NO2

-] (mM) Cytological Parameters 0.00 0.04 0.20 Nuclei Deformation of nuclear envelope - ++ ++++ Dilation of nuclear envelope cisternae - ++ ++++ Smooth endoplasmic reticulum (SER) Overall amount ++ +++ ++++ Dilation of cisternae - +++ ++++ Branching aspect - +++ ++++ Fragmentation of cisternae - +++ +++ Formation of SER whorls - ++ ++ Lysosomal elements Overall amount + ++ +++ Myelinated bodies +/- ++ +++ Cytoplasmic vacuoles ++ ++ +++ Mitochondria Overall amount ++++ +++ ++ Morphological heterogeneity - ++ +++ Formation of mitochondria clusters - + ++ Formation of myelin-like whorls - ++ +++ Vacuolization - ++ +++ Outer membrane rupture - - ++ Association with SER cisternae - ++ ++++ Association with lysosomal elements - +/- + - absent; +/- very little developed; + little developed; ++ moderately developed; +++ strongly developed; ++++ very strongly developed. The mitochondria showed great morphological heterogeneity. These organelles were partly associated with myelin-like membrane whorls, displaying vacuolization in the matrix region and showing a tendency to

56

1.4 µm

D

25 µm

A

0.3 µm

CA CB

M

T

N CC

PVCL

Figure 2. Control fish. A. Gill filament showing lamellae (L) and chloride cells

(arrowheads); B. Chloride cell cytoplasm showing mitochondria (M), tubular system (T) and nucleus (N); C. Photomicrograph showing chloride cells (CC) and short microvilli (arrowhead) on the apical surface. PVC, pavement cell.

aggregate in small mitochondrial clusters. The total amount of SER increased and the lysosomal elements, particularly small myelinated bodies, proliferated in the cytoplasm. Furthermore, chloride cells showed a large number of swollen mitochondrial profiles, which were often intimately associated with dilated SER cisternae and lysosomal elements. After exposure to 0.2 mM NO2, the lamellae were thinner, with few chloride cells on their epithelium (Fig. 3A). The chloride cells displayed a proliferation of SER (Fig. 3A-C). In addition, the SER cisternae showed signs of fragmentation and formation of SER whorls. Exposure to this concentration of NO2

- resulted in a high density of cytoplasmic vacuoles, which were very similar to lysosomal elements (Fig. 3C). The mitochondria displayed pronounced matrix vacuolization, partial lysis of cristae, and destruction of mitochondrial membranes (Fig. 3C). At this stage, vacuolated chloride cells were overlapped by epithelial cell extensions and had lost contact with the environment’s water. Despite this generalized pathological condition, a few chloride cells were hypertrophied, with their apical surface in contact with the environment’s water.

57

10 µm

A

2 µm

C

CC

EP

V

SER

M

M

M

SER

SER

MWL

A B

2,5 µm

CCC1

CC2

SER

W

V

C D

M

LCC

PVC

Figure 3. C. macropomum exposed to 0.04 (A-B) and 0.2 (C-D) mM NO2-. A.

Cross section of lamellae (L) showing the chloride cells (arrows) on lamellar epithelium; B. Chloride cell (CC) partially covered by pavement cell (arrowhead). Note the heterogeneity of mitochondria (M); C. A few chloride cells (arrow) on lamellar epithelium; D. Chloride cell (CC1) presenting apical pit (arrowhead) and chloride cell (CC2) buried in lamellar epithelium. PVC: pavement cell; SER: Smooth endoplasmic reticulum; V: vacuole

Discussion

The morphology of tambaqui, Colossoma macropomum, gills is similar to that described for other teleost species (Laurent and Dunel, 1980). The Importance of the role of the gills on gas exchange (Hughes, 1972), ion- and acid-base equilibrium (Evans, 1980), and nitrogen excretion (Haswell et al., 1980) is proportional to their susceptibility to changes in the environment. The

58

gills are the main sites for nitrite uptake; hence, the environmental Cl- concentration is a determining factor influencing NO2

- toxicity (Eddy et al., 1983). High ambient Cl- concentrations inhibit the uptake of NO2

- and protect against its toxic effects (Perrone and Meade, 1977; Bath and Eddy, 1980), for Cl- is a competitive inhibitor of NO2

- uptake (Bath and Eddy, 1980). The high density of chloride cells in the lamellar epithelium of C. macropomum may be due to the low ion concentration in the environmental water, which characterizes Brazil’s continental waters. Several studies have shown high chloride cell proliferation in fish living in ion-poor water (Perry, 1997), although Moron (2000) and Fernandes and Perna-Martins (2002) noted the significantly variable number of chloride cells in Brazil’s freshwater fish. Nitrite exposure caused a sharp reduction of chloride cells in lamellar epithelium, which may be related to increased cell death and low cellular differentiation in this epithelium. Indeed, most of the ultrastructural changes in chloride cells displayed a morphological pattern of necrosis. Previous studies have suggested that nitrite increases chloride cell activity to maintain ion homeostasis (Gaino et al., 1984), which, incombination with the direct toxic effects of nitrite, contributes to reduce the cell’s cycle. Our results are, in general, congruent with those found in the ultrastructure of chloride cells of Oncorhynchus mykiss exposed to 450 µgN-NO2/L (0.03 mM NO2

-) for 72h (Gaino et al., 1984). Furthermore, we also found signs of severe damage in chloride cell mitochondria after NO2

- exposure, evidencing clear irreversible cellular damage. NO2

- binds to heme moieties of mitochondrial cytochrome components of the respiratory chain, inhibiting respiration, and stopping or reducing ATP production (Cotran et al., 1994). It may cause mitochondrial swelling and rupture of mitochondrial cristae. Mitochondrial swelling has been related to the toxic action of compounds deriving from nitrite in rat liver hepathocytes (Rusu et al., 1980). In conclusion, the changes observed in the gill chloride cells investigated in this study indicate that the cellular structures involved in the process of energy production become severely damaged by exposure to nitrite. Acknowledgments This research work was supported by Fundação de Amparo à Pesquisa do Estado de São Paulo (FAPESP) and Conselho Nacional de Desenvolvimento

59

Científico e Tecnológico (CNPq), Brazil. O.T.F.Costa acknowledges CAPES for the award of a scholarship. References Bath RN and Eddy FB 1980 Transport of nitrite across fish gills. J. Exp. Zool. 214:

119-121. Costa OTF, Ferreira DJS, Mendonça FLP and Fernandes MN 2000 Acute toxicity

of nitrite to freshwater teleost tambaqui, Colossoma macropomum (Teleostei, Serrasalmidae). Comp. Biochem. Physiol.(Suppl) 126B: 26.

Cotran RS, Kumar V and Robbins SL 1994 Robbins Pathologic Basis of Disease.

(Schoen, FJ, ed.). W. B. Saunders Company: Philadelphia. Evans DH 1980 Kinetic studies of ion transport by gill epithelium. Am. J. Physiol.

238: R224-230. Eddy FB, Kunzlik PA and Bath RN 1983 Uptake and loss of nitrite from the

blood of rainbow trout, Salmo gairdneri Richardson, and Atlantic salmon, Salmo salar L. in freshwater and in dilute sea water. J. Fish Biol. 23: 105-116.

Fernandes MN and Perna-Martins SA 2002 Chloride cell responses to long-term

exposure to distilled and hard water in the gill of the armored catfish, Hypostomus tietensis (Loricariidae). Acta Zool (in press).

Gaino E, Arillo A and Mensi P 1984 Involvement of the gill chloride cells of trout

under acute nitrite intoxication. Comp. Biochem. Physiol. 77A: 611-617. Haswell MS, Randall DJ and Perry SF 1980 Fish gill carbonic anhydrase: acid-

base regulation or salt transport. Am. J. Physiol. 238: R240-245. Hughes GM 1972 Morphometrics of fish gills. Respir. Physiol. 14: 1–25. Kiernan JA 1990 Histological and histochemical methods: theory and practice.

Pergamon Press, London

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Laurent P and Dunel S 1980. Morphology of gill epithelia in fish. Am. J.

Physiol. 238: R147–R159. Lewis WM and Morris DP 1986 Toxicity of nitrite to fish: A review. Trans. Am.

Fish. Soc. 115: 183-195. Merz WA 1967 Die Streckenmessungen an gerichteten Strukturen im

Mikroskop und ihr Anwendung zur Bestimmung von Oberflachen-Volumen-Relationen im Knochengewebe. Mikroskopie 22: 132–142.

Moron SE 2000 Efeito das concentrações de íons Na+, Ca2+ e Cl- na morfologia

branquial e nos parâmetros fisiológicos de Hoplias malabaricus e Hoplerythrinus unitaeniatus (Teleostei Erythrinidae). Ph D thesis. Universidade Federal de São Carlos, 81 p.

Perrone SJ and Meade TL 1977 Protective effects of chloride on nitrite toxicity

to coho salmon (Oncorhynchus kisutch). J. Fish. Res. Board Can. 34: 486-492.

Perry SF 1997 The chloride cell: structure and function in the gills of freshwater

fishes. Ann. Rev. Physiol. 59: 325–347. Rusu MA, Preda N, Graciun C, Gadaleanu V, and Bucur N 1980

Histoenzymological and ultrastructural changes in rat following the administration of aminopyrine and nitrite (nitrosoaminopyrine). In: N-nitroso-compounds: Analysis formation and occurrence (Walker EA, Castegnaro M, Gricinte L, Borzsonyi M, eds.), Lyon IARC Scientific Publ. no. 31.

Strickland JDH and Parsons TR 1972 A Practical Handbook of Seawater Analysis.

Fish. Res. Bd. Canada Bull 167. Williams EM and Eddy FB 1986 Chloride uptake in freshwater teleosts and its

relationship to nitrite uptake and toxicity. J. Comp. Physiol. 156B: 867-872

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62

BEHAVIORAL AND NEUROPHYSIOLOGICAL EFFECTS OF

CARBAMATE PESTICIDES ON OLFACTORY CAPABILITIES

IN PACIFIC SALMON

Hugh E. Jarrard Department of Biological Sciences, Simon Fraser University

Burnaby, BC V5A1S6 Tel (604) 291-5634, Fax: (604) 291-3496, Email: hej@sfu.ca

and

C.K. Kennedy

Department of Biological Sciences, Simon Fraser University Email: ckennedy@sfu.ca

EXTENDED ABSTRACT ONLY – DO NOT CITE Introduction The presence of pesticides within aquatic environments can induce physiological and behavioral changes in teleosts that, although sublethal, impair the survivability and ecological fitness of the organism. We are interested in carbamate pesticide effects on olfactory behaviors and physiology in the Pacific salmon, a teleost that relies heavily on olfaction for successful completion of its life history. Specific impairment of olfactory-based behaviors by pesticides has been little studied in teleosts. Among salmonids, sublethal exposure to diazinon (organophosphate) reduces the ability of Chinook (O. tshawytscha) to react appropriately to alarm pheromones and to home [Scholz, 2000 #109]. An important issue is the locus and mechanism by which pesticide exposure may impair olfactory behaviors. Work at the neurophysiological level in Atlantic salmon (Salmo salar) suggests that pesticides can act on olfactory receptor neurons (ORNs), impairing their ability to respond to odorants of biological importance (e.g. Moore 2001).

63

We have begun to examine more thoroughly the mechanisms involved in pesticide impairment of salmonid olfaction by establishing a baseline of carbamate effects on Coho (O. kisutch) at behavioral and neurophysiological levels utilizing Y-maze avoidance-behavior assays and electro-olfactrograms (EOG), a sensitive measure of multi-unit voltage potentials created by ORNs in response to odorant stimulation (e.g. Hara 1973). The carbamates chosen are compounds of current concern with moderate to high aquatic toxicity and a limited refereed database regarding their olfactory toxicity. Methods Experimental Animals Juvenile Coho salmon were obtained locally at ages 3–5 months for Y-Maze experiments (mean weight= 0.97 g), and 10-15 months (mean weight= 16.0 g) for EOG experiments. Fish were maintained at natural photoperiod/temperature in flow-through tanks supplied with dechlorinated `background’ water (BKD; pH 6.2–6.8, hardness 3.49 -6.19 mg/L CaCO3). Pesticide Exposures For Y-maze experiments, fish were tank exposed for 21 days prior to behavioral testing to either carbofuran (0.075 ppm), 3-iodopropynylbutylcarbamate (IPBC, 0.048 ppm), mancozeb (1.1 ppm), or control condition in chilled (10-11°C) static 18L aquaria. Fish were fed 2% b.w. daily, and tanks were cleaned, refilled, and exposures refreshed every other day. For EOG experiments, exposures were acute (30 mins) and applied locally to the olfactory epithelium (OE). All pesticides dissolved in BKD immediately prior to use (exception: IPBC dissolved in polyethyelene glycol at 10 mg/ml; same vehicle also given to controls). Behavioral Testing Effects of sublethal pesticide exposure on olfactory behaviors were examined using salmonid avoidance of L-Serine (10-8 M; SER) in a two choice Y-trough maze (Fig. 1A; after Rehnberg et al. [1985]). Behavioral assays began with a single naïve fish in the start box with all gates closed. After 5 min. acclimation, SER was introduced into one arm of Y-maze for 5 min., then all gates were opened simultaneously. Fish were given 10 min. in maze to explore and choose an arm, after which all gates were dropped and the choice noted. The arm receiving SER was changed every second run, and the maze cleaned, in order to eliminate any bias.

64

Neurophysiological testing EOGs were recorded from coho parr after Evans and Hara (1985). Individual fish were anaesthetized (2-phenoxyethanol, 0.4 ml/L), and paralyzed with intra-muscular injections of Flaxedil (2.4 mg/kg b.wt.). Fish were then wrapped in gauze, secured in a Plexiglas trough, and maintained under anesthesia by gill perfusion. After removal of skin overlying the OE, a gravity-fed stream of BKD was passed over the exposed OE into which brief pulses of odorant (10-5 M SER) were introduced. EOG responses to SER delivery (Fig. 2A) were recorded from the OE using an Ag/Ag-Cl electrode, then filtered, amplified, and displayed on a computer. Experimental trials consisted of the collection of pre-exposure EOG responses (PRE; n=5), followed by acute OE pesticide exposure, then post-exposure EOG responses (POST; n=5). Results Y-Maze Behavior While statistical analysis has not yet been conducted on this data, it appears:

• Non-exposed control fish consistently avoid SER-scented arms • In contrast, carbofuran- and mancozeb-exposed groups choose

scented/unscented arms at a 50:50 ratio (suggesting anosmia) • IPBC-exposed fish avoid the SER-scented arm more frequently than

expected based on chance alone.

0

2

4

6

8

10

12

Contr ol Car bof ur an IPBC Mancozeb

Exposur e Gr oup

Scent ed

No Choice

Unscent ed

gate inflowdrain pumpSER

Figure 1. Y-Maze Behavioral Assay. A. Flow-through Y-maze apparatus. Flow: 6.4 L/min., maze length: 2.3 m. B. Effects of Pesticide Exposure on SER Avoidance. Plot displays number of fish choosing each arm per exposure group (carbofuran 0.075ppm, IPBC 0.048ppm, mancozeb 1.1ppm).

A B

65

EOG Results In preliminary results, significant reduction in EOG amplitude after pesticide exposure (in comparison with H2O-exposed controls, one-tailed t test, p < 0.05) occurred after acute 30 min. exposure:

• for carbofuran at 0.0001, 0.002, 0.01, and 2ppm • for IPBC at 0.00048, 0.0048, and 0.048 ppm • and, for mancozeb at 0.022, 0.22, and 2.2 ppm.

B

10-5 SER

peak amplitude

Fig. 2. Pesticide EEOG A. EOG resSER delivery. B.dependant carbamaton EOG amplitude.POST EOG amplitudPRE EOG amplitude

0.2 mV

5 secs

ffects on ponse to Dose-e effects Data as e as % of .

66

Conclusions • Carbamate pesticides interfere with salmon olfactory capabilities at

both behavioral and neurophysiological levels • This interference appears as loss of ability to avoid SER at low concen-

trations in Y-maze tests and as a significant decrease in the population of ORNs responding to SER in EOG responses.

• These effects appear at 50% of the 96 hr. LC50 for these compounds in behavioral tests (LC50s: carbofuran=200ug/l, IPBC=100ug/l, mancozeb=2.2mg/L), and at several orders of magnitude below that in neurophysiological studies.

• Ongoing work is focused on establishing both a clear correlation between these different classes of effects and the mechanism by which carbamates exert their effects in the olfactory periphery.

Acknowledgements

We wish to acknowledge the generous assistance of Dr. Kerry Delaney in establishing a functional EOG apparatus, the Capilano Hatchery for generous salmon supply, and Doug Wilson for equipment loans. HEJ supported by NIH NRSA Postdoctoral Training Fellowship F32-NS10973. References Evans, R. E. and T.J. Hara (1985). The characteristics of the electro-olfactogram

(EOG): its loss and recovery following olfactory nerve section in rainbow trout (Salmo gairdneri). Brain Res 330(1): 65-75.

Hara, T. J. (1973). Olfactory responses to amino acids in rainbow trout, Salmo

gairdneri. Comp Biochem Physiol A 44(2): 407-16. Moore, A. and C. Waring (2001). The impact of two pesticides on olfactory-

mediated endocrine function in mature male Atlantic salmon (Salmo salar L.) parr. Comp Biochem Physiol B Biochem Mol Biol 129(2-3): 269-76.

Rehnberg, BG and CB Schreck. (1987) Chemosensory detection of predators by

coho salmon (Oncorhynchus kisutch): behavioral reaction and physiological stress response. Can. J. Zool. 65:481-485.

67

68

Scholz, N. (2000). Diazinon dirsupts predator. Can J of Fish Aquat Sci 57: 1911-1918.

EFFECTS OF EXPOSURE TO SUB-LETHAL CONCENTRATIONS OF

AMMONIA AND HYPOXIA ON THE SWIMMING PERFORMANCE

OF BROWN TROUT (SALMO TRUTTA)

A. Shingles,

School of Biosciences, University of Birmingham, Birmingham, B15 2TT, United Kingdom.

Tel +44-(0)121-4145472; Fax +44-(0)121-4145925; a.shingles@bham.ac.uk

D.J. McKenzie1, S. Ceradini2, A.Z. Dalla Valle3, A. Moretti2 & E.W. Taylor1

1Biosciences, University of Birmingham, Birmingham, UK; 2Environment Unit, CESI SpA, Milan, Italy.

3Pharmacological Sciences, University of Milan, Milan, Italy.

EXTENDED ABSTRACT ONLY – DO NOT CITE Introduction Ammonia is toxic to all vertebrates. It has become a pervasive pollutant of aquatic habitats but is also, for the majority of aquatic organisms, an end-product of protein metabolism. In teleost fish ammonia can, therefore, accumulate to toxic levels either as a consequence of exposure to elevated water ammonia concentrations or as a consequence of impaired excretion of the endogenous metabolite. Ammonia accumulates in fish during exposure to sub-lethal concentrations of heavy metals such as copper and, in brown trout, this has been linked statistically to a decline in the ability to perform exercise (Beaumont et al., 1995). Beaumont et al. (1995) found a negative linear relationship between plasma ammonia concentrations and maximum sustainable swimming speed (Ucrit) in brown trout exposed to a sublethal combination of copper and acidic water, and went on to demonstrate that the ammonia accumulation caused a partial depolarisation of muscle membrane potential (Beaumont et al., 2000). Shingles et al. (2001) demonstrated that exposure to elevated water ammonia alone was sufficient to reduce Ucrit in rainbow trout (Oncorhynchus mykiss), with evidence that this was linked to a partial depolarisation of muscle.

69

Objectives of the current study included to investigate whether a linear relationship between plasma ammonia concentration and Ucrit could be elicited in brown trout by ammonia exposure alone, and to investigate the mechanisms for impaired performance. Hypoxia is a growing problem in many aquatic habitats and, therefore, can be expected to occur concurrently to elevated ammonia. Hypoxia impairs swimming performance in salmonids by limiting aerobic scope (Bushnell et al., 1984). Thus, a further objective of the current study was to determine how concurrent exposure to ammonia and hypoxia influenced exercise performance in brown trout. Methods To investigate the relationship between plasma ammonia concentration and swimming performance, adult brown trout (mean mass approx. 500g) were exposed to two different sub-lethal concentrations of NH4Cl in the water, nominally 100 µmol l-1 and 200 µmol l-1, for 24h in hard water at 15 °C and pH 8.2. Their swimming performance, and associated respirometry, were then investigated in a Brett-type swimming respirometer (Shingles et al., 2001), in comparison to controls in normal water. In parallel, the effects of the ammonia exposure regimes on plasma ammonia and on the membrane potentials (EM) of red muscle, white muscle, heart and brain were investigated on trout with an indwelling catheter in the dorsal aorta. To investigate the effects of ammonia and hypoxia, the experimental series were repeated, but trout were exposed to mild hypoxia at a nominal water O2 partial pressure of 80 mmHg for 1h prior to, and then during, the various measurements. Results Exposure to either 100 µmol l-1 or 200 µmol l-1 NH4Cl caused an increase in plasma total ammonia to 386±42 µmol l-1 or 771±92 µmol l-1, respectively, compared with 133±29 µmol l-1 in control fish (mean ± SE, n = 6). This ammonia accumulation was associated with a significant decline in Ucrit from 2.24±0.15 bodylengths s-1 (BL s-1) in control trout to 1.46±0.09 BL s-1 or 1.08±0.16 BL s-1 in trout exposed to 100 µmol l-1 or 200 µmol l-1 NH4Cl, revealing a direct negative relationship between plasma ammonia concentration and Ucrit (Figure 1). The linear relationship was surprisingly similar to that

70

observed by Beaumont et al. (1995) in brown trout exposed to sublethal copper at low pH, which is shown on the figure for comparison.

00.0

0.5

1.0

1.5

2.0

2.5

3.0

200 400 600 800 1000plasma [ammonia] (µmol l-1)

Ucr

it(B

L s-1

)

00.0

0.5

1.0

1.5

2.0

2.5

3.0

200 400 600 800 1000plasma [ammonia] (µmol l-1)

Ucr

it(B

L s-1

)

Figure 1. The least squares linear regression relationship between plasma

ammonia concentration and maximum sustainable swimming speed (Ucrit) in brown trout exposed to three water concentrations of ammonia (this study, solid symbols, solid line); as reported for brown trout that accumulate ammonia following exposure to a sub-lethal combination of copper and low pH (Beaumont et al., 1995, dashed line), and in rainbow trout exposed to two water concentrations of ammonia (plotted from data reported in Shingles et al. 2001, open symbols, dotted line). The error bars on the symbols are ± 1 SEM of the mean values for plasma ammonia and Ucrit, n = 6 or 7. For brown trout exposed to ammonia (this study), the linear regression was described by the equation: mean Ucrit = -0.0018(mean[ammonia]) + 2.347 (R2 = 0.903, n = 3). Beaumont et al. (1995) reported an equation based on individual values rather than means: Ucrit = -0.0020[ammonia] + 2.089 (R2 = 0.670, n = 30). For rainbow trout (Shingles et al., 2001), the equation was: mean Ucrit = -0.0024(mean[ammonia] + 2.674 (n = 2).

71

Also, the figure shows the relationship between plasma ammonia and Ucrit in rainbow trout, replotted from the data reported in Shingles et al. (2001). The relationship between the plasma ammonia concentration and Ucrit was very similar in the two species. Interestingly, however, a significantly higher water ammonia concentration was required in rainbow trout to elicit the same plasma ammonia accumulation (Shingles et al., 2001). The respirometry measurements revealed that the impaired performance in the brown trout exposed to ammonia was associated with reduced swimming efficiency, and with a partial depolarisation of EM in white muscle and the brain, although there were no significant effects on EM of the red muscle and heart. Exposure to hypoxia caused a 45% decline in Ucrit in control animals, down to

1.23±0.09 BL s-1, as a consequence of the expected limitation to aerobic scope. Hypoxia did not, however, cause the same proportional decline in performance in the ammonia-exposed fish; both groups had a Ucrit of approximately 1 BL s-1. Therefore, hypoxia had no further effect on the reduced performance of animals exposed to 200 µmol-1 NH4Cl. Conclusions The results provide further evidence that the impaired swimming performance of trout following exposure to sub-lethal concentrations of copper in acid water (Beaumont et al., 1995) can be attributed to the accumulation of ammonia. Ammonia accumulation has similar effects on performance in both rainbow and brown trout, but rainbow trout appear better able to limit plasma ammonia accumulation during exposure to elevated water ammonia. The fact that hypoxia did not elicit any further decline in Ucrit in trout exposed to NH4Cl indicates that ammonia impairs performance by a mechanism unrelated to oxygen supply in brown trout, perhaps through effects on nerve and white muscle function. References Beaumont, M.W., P.J. Butler and E.W. Taylor. 1995. Plasma ammonia

concentration in brown trout (Salmo trutta) exposed to acidic water and sublethal copper concentrations and its relationship to decreased swimming performance. J. Exp. Biol. 198: 2213-2220.

72

Beaumont, M.W., E.W. Taylor and P.J. Butler. 2000. The resting membrane potential of white muscle from brown trout (Salmo trutta) exposed to copper in soft, acidic water. J. Exp. Biol. 203: 2229-2236.

Bushnell, P.G., J.F. Steffensen and K. Johansen. (1984). Oxygen consumption

and swimming performance in hypoxia-acclimated rainbow trout Salmo gairdneri. J. Exp. Biol. 113: 225-235.

Shingles A., D.J. McKenzie, E.W. Taylor, A. Moretti, P.J. Butler and S.

Ceradini. (2001). Effects of sub-lethal ammonia exposure on swimming performance in rainbow trout (Oncorhynchus mykiss). J. Exp. Biol. 204: 2699-2707.

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74

EFFECTS OF CHLORPYRIFOS (LORSBAN) ON REPRODUCTIVE

PERFORMANCES OF GUPPY (POECILIA RETICULATA)

De Silva, P. M C. S Department of Zoology, University of Ruhuna,

Matara, Sri Lanka. Telephone: 0094 41 27025

Fax: 0094 4122683 Email: chathura@zoo.ruh.ac.lk

Samayawardhena, L. A,

Department of Zoology, University of Ruhuna, Matara, Sri Lanka.

Telephone; 0094 41 27025 Fax: 0094 4122683

Email: samaya@zoo.ruh.ac.lk

Abstract Toxicity of Chlorpyrifos on reproductive disorders including altered fertility, reduced viability of offspring, impaired hormone secretion and modified reproductive anatomy were little concerned. In the present study, we selected Guppy (Poecilia reticulata) to investigate the reproductive effects of Lorsban, a common insecticide in Sri Lanka. Male and female guppy were selected with proven fertility from our own colony and the groups of fish (n=12x6/group) were exposed to pre-determined 2µg / l, 0.002 µg /l Chlorpyrifos based on the 96 hrs LC50 for guppy. Mating behavior of pairs was recorded on the 2nd day of exposure. Offspring were counted and survival recorded on the 14th day. Gonopodial thrusts (4 /15 min, in 2 µg/l and 8/15 min in 0.002 µg/l) were significantly different from the control (11/15min, in the control). Similarly, live birth reduced significantly to 8/female in 2 µg/l compared to 27/female in the control group. Survival of offspring at the 14 days was reduced to 66% in 2 µg/l group. Our findings show that low soluble concentrations of Chlorpyrifos can impair reproductive behavior and capabilities of Guppy to a significant extent. F1 generation of treated fish showed reduced survival suggested importance of mating behavior. Pesticide exposure throughout embryonic development could result weak offspring and lesser their survival. Our study confirmed

75

Chlorpyrifos could potentially alter mating behavior, live birth and F1 survival of Guppy. Introduction Chlorpyrifos (O, O–diethyl–O (3,5,6–trichlor–2–pyridyl) phosphorothioate) is a broad-spectrum organophosphate insecticide with growing concern due to its aquatic toxicity (Foe, 1998; Bailey et al 1997). The toxicity effects may include neurological, behavioral and possibly reproductive effects (Mueller-Beilschmidt, 1990; Hill, 1995). Recent studies have revealed that Chloropyrifos together with Diazinon are responsible for most of the toxicity to aquatic organisms (de Vlaming et al., 1993; Moor et al., 1998; Foe et al., 1998). Especially it is very highly toxic to fresh water fish and aquatic invertebrates. The agricultural, residential and commercial use of Chloropyrifos on pest control leads to presence of Chloropyrifos in sufficient concentrations in agricultural runoff and as well as in urban storm water runoff resulting high toxic effects on Ceridodaphnia and Mysidopis, two zooplankton species (Corner et al., 1998; Larson et al., 1998). Similar toxicity assessment subjected by Lee and Jones- Lee (1997). Toxicity of Chloropyrifos on zooplankton is well documented, but less work carried on fish. Johnson and Finley (1980), Odenkirchen et al, (1988) documented its acute toxicity. Reproductive disorders including altered fertility, reduced viability of offspring, impaired hormone secretion and modified reproductive anatomy were little concerned. Few early studies give some viable information on this aspect. Dursban exposure on fathead minnows for 200 days resulted a decline of offspring in first generation survival (USPHS, 1989). Growth of early life stage of California grunion (Leuresthes tenuis) was reduced 20% and 26% in two low concentrations of Chloropyrifos (Odenkirchen et al., 1988). But the validity of reproduction as a key parameter to evaluate the impacts of chemical pollutants Kime (1999) tends us to study the Chloropyrifos toxicity with some aspects of reproduction. Guppy (Poecilia reticulata, Peters) is selected as the model organism, a livebearer fish species that was introduced to Sri Lanka as a biological tool in mosquito control. The experimental organism Guppy is a viviparous fish with a short reproductive period (Houde, 1997). Male guppies approach two methods of mating behaviour, sigmoid displays and gonopodial thrusts (Evens et al.,

76

1999). They perform a sigmoid display in which the body is held in S- shape while fins are extended and quivered. Alternatively they may attempt sneaky mating, in which the female is approached sideways or from behind and the modified anal fin the gonopodium is thrust toward the genital pore. Successful mating produced a litter size range from 12 – 46 in monthly intervals depending on circumstances (Hutchins, 1996). In this study we observed their mating behaviour representing an organism level parameter, brood size and the survival of the offspring as a population level parameter with two more or less similar concentrations of Chloropyrifos. 2 Materials and Methods 2.1 Study population and their maintenance We collected wild guppies (Poecilia reticulata) from the urban cannel systems around the Nilwala river basin in the southern region of Sri Lanka. Fishes were returned to the laboratory and stocked in (200 l) tanks. This stock aquarium received fully aerated water from a header tank. The water temperature was kept at 26 ± 2 C0.These fishes were fed with special aquaria food purchased through local market. By maintaining this colony we select female guppy that belongs to the highest length class (3.5±1.0 cm) was selected as the test female animals assuming that they were well suited to give birth to offspring. They were separated into (18 l) tanks and monitored for 4 – 6 weeks, until they had given birth to offspring due to early fertilization in stock aquaria. Adult males representing the length class (2.0 ± 1.0 cm) were selected based on their early sexual behaviour. 2.2 Stress exposure Pre determined 2 µg / l, 0.002 µg / l were used as the exposure concentrations based on the 96 Hrs LC 50 for guppy (our own study). Each pair of guppy was transferred to the tanks (23x23x35cm), that contained 10 l of Lorsban EC 40% (Chloropyrifos) solutions for consecutive three days together with controls. Chlorinated free tap water was used in the process of dilution of the pesticide. Test solutions were changed every 24 hours followed by the addition of fresh Chloropyrifos solution. After three day exposure time male guppies were removed and the females were kept in the test solutions until they produced their brood.

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2.3 Mating behaviour Sigmoid displays and gonopodial thrusts characterize male mating behaviour of guppy. In sigmoid displays the cooperation of the female is necessary hence we observed the number of gonopodial thrusts as our observation because it is an alternative mating tactic, which does not require female reception. When the modified anal fin, gonopodium made contact with the female genital region it was referred as a successful attempt of mating. A day after exposure to the concentrations the male mating behaviour was observed for consecutive two days. The number of gonopodial thrusts performed by each male in the tanks was calculated over 15 minutes using a counting deviser. 2.4 Reproductive rate and Offspring survival The reproductive capabilities of the pair were estimated by counting the total number of offspring born to the female. The reproductive rate was calculated as the average number of offspring born per female for each concentration and the control. The number of offspring born to each pair was separated into glass tanks (23 x 23 x 35 mm) and monitored for consecutive two weeks. The mortality in each set of offspring was observed in every 24 hrs and the total number of dead offspring was determined after a period of two weeks. Statistical analysis The 96 hrs LC50 was determined through probit analysis, using the software SPSS for windows 98.The relationship between the observations in exposed concentrations and control experiments were investigated by ANOVA modeling and comparisons by Student-Newman-Keuls test. All analysis was subjected at < 0.001 probability level. 3.Results During the overall study period General appearance was prime in condition and no pathological effects were observed. Also there were no changes in feed intake in both control and treatment groups. The 96 hrs LC 50 was 7.17 µg / l. Mating behavior observed as number of gonopodial thrusts performed by male in 8.00 hr, 12.00 hr, and 6.00 hr respectively, with control and two treatments as given in Figure 1.

78

Fi gur e 1 : M e a n numbe r of gonopodi a l t hr ust s pe r f or me d by ma l e i n di f f e r e nt pe r i ods of t he da y .

0

2

4

6

8

1 0

1 2

1 4

8.00am 1 2.00pm 6.00pm

T i me

control

T 1

T 2

Fi gur e 2 : M ean number of gonopodi al t hr ust s per f or med by mal e i n exposed concent r at i ons ( ** *P < 0. 001 ) .

0

2

4

6

8

1 0

1 2

1 4

con t r ol T 1 T 2

E x p o s e d c o n c e n t r a t i o n

It is quite evident that the time and the number of attempts were not significantly different. But number of gonopodial thrusts by male was considerably reducing with increased concentration of LORSBAN. In control mean number of gonopodial thrusts were recorded as 11 while in males in 2 µg / l (T 1) it was 4 and 8 in the lowest concentration 0.002 µg / l (T 2) (Figure 2).

Each of these values were significantly different at P<0.001. The litter size calculated as number of offspring per female varies to minimum number of 18 to maximum of 36 in the control, which made it to average of 27 offspring per female. Female guppy in exposed concentration 2 µg / l recorded least number of offspring ranging 5 – 11 and mean of 8 offspring per female. But in the lowest concentration the mean number of offspring per female has risen to 24.

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Although the mean number of attempts were reduced at 0.002µg / l, still exposed females were able to produce a closer number of offspring as the control (Figure 3).

Figure 3: Mean,maximum,and minimum number of offspring per female with different concentration (Mean,***P< 0.001).

0

5

10

15

20

25

30

35

40

control T 1 T 2

Exposed concentration

Num

ber o

f offs

prin

g Max

Min

Mean

MinMean

Produced offspring in control for additional 14 days were recorded a highest percentage survival, over 86% while in the concentration of 2µg / l it was less than 66. Quite contrast to the observations related to number of offspring in the control and exposed concentration, 0.002µg / l, the percentage survival of this concentration significantly lesser than control (Figure 4).

Figure 4: Recorded maximum,minimum and mean percentage survival with exposed concentrations( ***P< 0.001)

0

20

40

60

80

100

120

control T 1 T 2

Exposed concentrations

Perc

enta

ge S

urvi

val

Max***

***

80

4.Discussion Fish exposed to xenobiotic pollutants have manifested a range of reproductive defects including behavioral, anatomical, physiological levels from reduced fertility to alternation of sexual behaviour and one of the key parameters to evaluate the impact of xenobiotics (Jones et al, 1997; Kime, 1999). Reproductive behavior it self is a key phase of the reproduction cycle of Guppy, which ensures the successful mating. Endler (1987) and Houde (1997) have illustrated the two approaches as sigmoid displays and alternative tactic gonopodial thrusts difference in change of temperature, in presence of predators. But the effects of pesticides the sexual behaviour was either poorly highlighted or neglected. Although the degradation of Chloropyrifos is rapid still it can cause serious consequences on fish as it coincides with this key phase of their reproduction cycle, mating behaviour. Our study suggests that even low concentrations of Chloropyrifos well below LC50 value heavily ceased male mating behaviour. Similarly Matthiessen and Logan (1984) suggested that 0.002 – 0.0015 mg/l exposure of Endosulfan inhibited male reproductive behavior of Sarotherodon mossambicus. Moore et al., (1997) focused on interesting phenomenon stating as successful mating behavior sometimes depend on the secretion of chemicals by female fish to elicit a response that both triggers production of sperm and male mating behaviour. Pesticides such as carbofuran and diazinon both could disrupt male fish to detect such chemicals. Similar mechanism could be suggested for Chloropyrifos since its effects are more similar to diazinon. Further research on this aspect might be able to confirm this mechanism. In our early study using guppy juveniles and Chloropyrifos suggested that exposed animals had signs of paralysis even in low concentrations of this pesticide. These signs of paralysis which caused them, incapable of moving might lead to the reduction of their mating behaviour. Male guppy reduces their mating tactics in the presence of predators (Endler, 1987). In natural environment impacts of Chloropyrifos alone with predators might cause the situation worst. Guppy, which can produce the brood to the external as fries and its short reproductive cycle, provides an excellent model to examine effects of Chloropyrifos on female fertility. The sperm storage of female guppy ensures that female could fertilized new embryos even if she was unable to remate (Constantz, 1989). Since we used females, which confirmed as no early mating, prior to exposure of Chloropyrifos we strongly suggest that reduction of average

81

number of offspring per female associated with pesticide exposure. It is further highlighted as the highest concentration was used have been recorded the least number of offspring. It is true that male vertebrates produce more sperm than eggs. But studies on fish and mammals confirmed that small change of sperm quality and quantity could reduce female fertility which might resulted fewer offspring (Kime, 1999). Change of mating behaviour in exposed males could lead to change of sperm quality and quantity resulting few offspring in exposed groups. The validity of this observation needs further research on Chloropyrifos effects on sperm quality and quantity. Decrease production of yolk protein resulting from inhibition of ovarian or liver function may lead to small number of eggs (Tyler et al, 1990). In this study we did not focus on this aspect but this relationship should be revealed with this experiment as well. Several authors suggested decreasing number of offspring with various pesticides. Yasuno (1980) used the same species Guppy as model organism to evaluate the effects of Fenitrothion and he suggested that reduced number of juveniles in pesticide contaminated female population. Exposure of Dursban on Fathead minnows for 200 days resulted the reduction of first generation of offspring. Hose et al, (1989) and Thomas et al, (1989) suggested similar results on fish as the test species. Another aspect that could have a possible relationship was percentage survival of the juveniles born. Although females in the lowest concentration were able to produce quite similar number of offspring with control their survival was significantly lesser than control. Xenobiotics could pass on from mother to offspring when it was developing. Pesticide exposure throughout this embryonic development could result week offspring that makes them struggle to survival. Less than 66 % of survival in exposed concentrations of this study revealed that fries affected by Chloropyrifos, although they were not entirely exposed. Contaminated yolk passed on from mothers who have accumulated high pesticide burden and liver alterations itself might lead to the nutrition content and quality of the eggs, which were the possible causes for weak offspring production. Also we suggest that Chloropyrifos could decrease the responses to stresses and further decrease in growth and metabolism could possibly affect these juveniles, ability to survival. In wild they might highly struggling to survival since they are vulnerable to combination of stress conditions including effects of pesticides. There is increasing evidence that some of the problems found in fish now being applied to human population as well (Colborn, et al 1996). So suggested effects

82

of chlorpyrifos on guppy might insight into a much more general view as well. We can conclude that LORSBAN (Chlorpyrifos) could potentially impair mating behavior and the effects could be extended to survival of F1 up to 14 days after birth. References

Bailey, H. C., Miller, J.L., Miller, M.J., Wiborg, L.C., Deanovic, L., Shed, T. 1997. Joint acute toxicity of diazinon and chlorpyrifos to Ceriodaphnia dubia. Environmental Toxicology and Chemistry.16 (11). 2304-2308.

Colborn.T, Dumasanoski and Myres, J.P. 1996. Our Stolen future. Little brown. London.

Connor, V., 1995. Status of urban storm runoff projects. Central Valley Regional Water Quality Control Board, Sacramento, CA.

Constanz, G.D. 1989. Reproductive biology of poeciliid fishes. Ecology and evolution of livebearing fishes (Poeciliiadae)(Ed.by G.Kmeffe & F.F Snelson, Jr) pp33-50. EnglewoodCliffs, New Jersy, Prentice Hall.

deVlaming, V. DiGiorgio, C. and Deonovic, L., "Insecticide-Caused Toxicity in the Alamo River," Presented at the NorCal SETAC annual meeting, Reno., NV, June. (1998).

Endler, J. A. 1987. Predation, light intensity and courtship behaviour in Poecilia reticulata (Pisces: Poeciliidae). Anim.Behav.35. 1376-1385.

Evens.J.P, Magurran.A.E. 1999. Geographic variation in sperm production by Trinidadian guppies. Proc.R.Soc.Lond. B266, 2083-2087.

Foe, C., Deanovic, L., Hinton, D. 1998. Toxicity Identification Evaluations of Orchard Dormant Spray Storm Runoff. California Regional Water Quality Control Board, Central Valley Region, Sacramento, CA.

Hill, E.F. 1995. Organophosphorus and Carbamate Pesticides. Pp. 243-274 in D. J. Hoffman, B.A. Rattner, G. A. Burton, Jr., and J. Cairns, Jr. (eds.). Handbook of Ecotoxicology. Lewis Publishers, Boca Raton, FL.

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Hose, J.E., Cross., Smith, S.G ., Diehl.D.1989. Reproductive impairment in a fish inhabiting a contaminated coastal environment off southern California. Env. Poll. 57, 139-148.

Houde, A. E. 1997.Sex, colour, and mate choice in Guppies. Princeton University Press. New Jercy.USA.

Hutchins,L.1996.Freshwaterresources.OnlinePublication.http://www.Aqualink.com/Fresh/Z1-Guppy.

Jones, J .C., Reynolds J.D. 1997.Effects of pollution on reproductive behaviour of fishes. Rev. Fish.Biol. & Fisheries 7. 463-491.

Johnson, W.W., Finley, M.T. 1980. Handbook of acute toxicity of chemicals to fish and aquatic invertebrates. U.S.Fish.Wild.Serv.Resour.Pub.137, 98-99.

Kime.D.E. 1999. Endocrine disrupting chemicals.,ed R.F Hester and R.M Harrison, Issues in Environ.Sci.and Tech.12. 27-48.

Landis, W.G., Yu, M.H. 1995. Introduction to Environmental Toxicology: Impacts of Chemicals Upon Ecological Systems. Lewis Publishers, Boca Raton, FL.

Larsen, K. L, Connor, V.M and Hinton, D.E."Sacramento River watershed Program Toxicity Monitoring Results 1996 - 1997,SETEC annual meeting Reno, NV.June.(1998).

Lee, G.F., Jones-Lee.1998. A Development of a regulatory approach for OP

pesticide toxicity to aquatic life in receiving waters for urban storm water runoff. SETAC Meeting, Reno, NV.

Matthiessen, P.Logazn, J.W.M.1984.The effects of low concentrations of

endosulfan insecticide on reproductive behavior in the tropical cichild fish, Sarotherdon mossambicus. Bull. Contam. Toxicol.33.575 - 583.

Mueller-Beilschmidt, D. 1990. Toxicology and environmental fate of synthetic pyrethroids. J. Pest. Reform .10(3). 33-34.

Moore, M. T., Huggett, D.B., Gillespie, W.B., Rodgers, J.H.J., Cooper, C.M. 1998. Comparative toxicity of chlordane, chlorpyrifos, and aldicarb to four aquatic testing organisms. Archives Environmental Contamination and Toxicology. 34. 152-157.

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Odenkirchen, E.W., Eisler, R. 1988.Chlorpyrifos hazards to fish, wild life, and invertebrates synoptic review. U.S fish wild. Serv Biol. Rep.85, 9-11.

Thomas, P. 1990.Teleost model for studying the effects of chemicals on female reproductive endocrine function. J. Exp. Zool.Suppl.4. 126-128.

Tyler.C.R, Sumpter.J.P.1996.Oocyte growth and development in teleosts. Rev.Fish.Biol.Fisheries.6, 287.

U.S. Environmental Protection Agency. Registration Standard (Second Round Review) for the Registration of Pesticide Products Containing Chlorpyrifos. Washington, DC, 1989.5-44.

Yasuno.M, Hatakeyama.S, Miyashita.M., 1990. Effects on reproduction in the

guppy (Poeciliia reticulata) under chronic exposure to Temphos and Fenitrothion. Bull. Env.Contam.Toxicol. 25, 29-33.

Acknowledgements

Authors acknowledge University of Ruhuna research grant (RU/SF/RP/99/05) for providing financial assistance.

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86

MATERNAL TRANSFER OF COPPER RESISTANCE IN FATHEAD

MINNOWS

Elizabeth B. Peake Biology Department, University of Nebraska at Omaha

6001 Dodge Street, Omaha, Nebraska 68182-0040, USA phone (402) 221-4474, fax (402) 221-4886

e-mail: peake@cox.net

Laura L. Tierney, Jessica C. Locke, and Alan S. Kolok Biology Department, University of Nebraska at Omaha

EXTENDED ABSTRACT ONLY – DO NOT CITE

The overall objective of this research was to determine if differential resistance to copper (Cu) could be transmitted from adult fish to larval offspring. Differential resistance can be transferred from adults to offspring in two different ways. The first is genetic inheritance. This is supported in the literature by several studies that found that differences in Cu resistance were significantly correlated with differences in allozyme genotypes (e.g. Schlueter et al., 1995). The second way is by maternal transfer of non-genetic material (e.g. Lin et al., 2000). We conducted two experiments with fathead minnows (Pimephales promelas) to investigate these two potential methods for transfer of Cu resistance from parents to larval offspring. In our first experiment, we tested the hypothesis that genetic differences in adults were a major influence in determining the resistance of larvae. We used the same methods as detailed in Kolok (1998) to classify 48 male and 48 female adult minnows as being either Cu-susceptible or Cu-resistant. Critical swimming speeds (Ucrits) were measured for the 96 minnows before and after an 8-9 d exposure to 150 µg/L Cu. Cu resistance or susceptibility was based on the percent decrease in Ucrit after Cu exposure. After the minnows were classified, the 12 most Cu-susceptible pairs and 10 most Cu-resistant pairs were bred together. A 7-d survival test (U.S. EPA, 1994) was conducted on the larvae produced by Cu-resistant, Cu-susceptible, and naïve (previously unexposed) parents. Eighteen 600-ml beakers, each with 250 ml of Cu solution, served as test chambers. Larvae <24 h old from five to seven breeding pairs were distributed

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evenly among the test chambers, with about 15 larvae per test chamber. There were triplicates at each of six Cu concentrations (0, 100, 200, 400, 800, and 1600 µg/L), with static renewal of ≥80% daily with freshly made test solutions. Actual Cu concentrations were analyzed by flame atomic absorption spectrometry. Mortality was recorded daily. Figure 1 shows mortality for the four groups, with replicates pooled.

0

0.2

0.4

0.6

0.8

1

MO

RTA

LITY

0 500 1000 1500 Cu CONCENTRATION (ug/L)

SUSCEPTIBLE

RESISTANT

CONTROL 2

CONTROL 1

Figure 1. Mortality after 7 days for four larval groups at six actual Cu

concentrations, 0-1400 µg/L Cu. The LC50s of the larvae from Cu-susceptible and Cu-resistant parents were surprisingly similar (922 and 924 µg/L, respectively) and significantly greater (t-test, p=0.02) than those of controls (412 and 210 µg/L). Cu-susceptible parents did not produce Cu-susceptible larvae; therefore the data did not support the hypothesis that the relative Cu tolerance of the larvae would be similar to that of their parents.

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In the second experiment, we tested the hypothesis that larvae would be more Cu resistant if their female parent had been previously exposed to Cu 1 week prior to breeding. We bred 12 pairs of minnows and conducted 96-h time-to-death tests on their larvae. About 20 larvae <24 h old from each breeding pair were placed in each of three 600-ml beakers. Each of these test chambers had 350 ml of Cu solution at 0, 400, or 800 µg/L Cu, with daily static renewal of ≥80% with freshly made test solutions. Mortality was recorded 3, 6, 12, 24, 48, and 96 h after test initiation. Six females were exposed to 100 µg/L Cu for 5 d, while the remaining females were sham-exposed. Females were then bred with their original partners, and 96-h larval time-to-death tests were again conducted as above. Figure 2 shows cumulative survival for the larval groups.

0.6

0.7

0.8

0.9

1

1.1

0 24 48 72 96

POST-EXPOSUREPRE-EXPOSURE

0.7

0.8

0.9

1

CU

MU

LATI

VE

SU

RV

IVA

L

0 24 48 72 96HOURS

POST-EXPOSUREPRE-EXPOSURE

B

A

Figure 2. Survival for pre- and post-exposure groups of larvae from female

parents exposed to (A) 100 µg/L Cu and (B) 0 µg/L Cu.

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For both groups, the effects of each Cu test solution concentration was significantly related to time-to-death, and their inclusion significantly improved the fit of the model to the data. The pre-exposure and post-exposure groups of larvae from sham-exposed females were not significantly different in survivorship. For larvae from Cu-exposed females, however, post-exposure survivorship was significantly higher than pre-exposure survivorship; indicative of maternal transfer. We were surprised to find that Cu susceptibility in the adults was not directly correlated with Cu susceptibility in their larvae. One reason may be that the mechanism for Cu susceptibility may be different between these two life history stages. Upon closer examination (our second experiment), it became apparent that maternal transfer had a much more important influence on the relative Cu resistance of larvae than we initially anticipated. Acknowledgements We thank the Biology Department and the University Committee on Research, University of Nebraska at Omaha (UNO) (Omaha, NE, USA) for funding. The University of Mississippi Environmental Toxicology Research Laboratory (Oxford, MS, USA) and the University of Nebraska-Lincoln Groundwater Chemistry Laboratory (Lincoln, NE, USA) assisted in the analysis of water samples for Cu. We also thank laboratory workers Koryn Boss, Sarah Cederstrand, Randy Johnson, Darcy L'Etoile-Lopes, and Tony Mertz as well as the Animal Care Services staff at UNO. References Kolok, A.S., E.P. Plaisance and A. Abdelghani. 1998. Individual variation in the

swimming performance of fishes: An overlooked source of variation in toxicity studies. Environ. Toxicol. Chem. 17:282-285.

Lin, H.C., S.C. Hsu and P.P. Hwang. 2000. Maternal transfer of cadmium

tolerance in larval Oreochromis mossambicus. J. Fish Biol. 57:239-249. Schlueter, M.A., S.I. Guttman, J.T. Oris and A.J. Bailer. 1995. Survival of

copper-exposed juvenile fathead minnows (Pimephales promelas) differs among allozyme genotypes. Environ. Toxicol. Chem. 14:1727-1734.

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U.S. Environmental Protection Agency. 1994. Short-term methods for

estimating the chronic toxicity of effluents and receiving waters to freshwater organisms. 3rd Ed. EPA/600/4-91/002. Environmental Monitoring Systems Laboratory, Cincinnati, OH, USA.

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92

REDUCED GONADAL SOMATIC INDEX AND EXTERNAL

COLORATION FOLLOWING EXPOSURE

TO P,P’-DDE IN ADULT MALE FUNDULUS HETEROCLITUS

Emily Monosson, Box 329/15 North Street Montague MA 01351

phone/FAX 413-367-0052emonosson@forwild.umass.edu

Stephen D. McCormick, Michael F. O’Dea, USGS, Leetown Science Center, Conte Anadromous Fish Research Center, Turners Falls, MA

Gregory M. Weber, USDA/ARS National Center for Cool and Cold Water Aquaculture, 11876 Leetown Road, Kearneysville, WV 25430

EXTENDED ABSTRACT ONLY – DO NOT CITE

The persistent organochlorine, p,p’-DDE (a lipophilic metabolite of DDT) is a potent antiandrogen that binds the androgen receptor (AR) in rats (Kelce et al. 1995). p,p’-DDE also binds ARs in fish in vitro (Thomas 2000) although in vivo studies investigating the antiandrogenic activity of p,p’-DDE in fish have produced conflicting results. Carlson et al. (2000) reported a lack of effects by either p,p’DDE or o,p’DDT on sexual development in rainbow trout embryos, whereas Baatrup and Junge (2001) recently demonstrated that p,p’DDE can act as an antiandrogen in male guppys. We evaluated the effects of p,p’DDE exposure on gonadal growth, coloration (male Fudulus develop yellow coloration when mature), and blood plasma concentrations of testosterone (T) and 11-ketotestosterone (KT). Two similar studies were conducted in 1998 and 1999. Fundulus collected from Stony Brook, NY in the fall of 1997 were held approximately six months prior to the 1998 study (10 ppt salinity, 10oC) and for 1.5 years prior to the 1999 study. In March 1998, 56 Fundulus were allocated to tanks (12 fish per tank, with two tanks per dose), acclimated for five days and injected with either 0, 10 or 100 ppm p,p’DDE dissolved in olive oil. Temperature was then increased on the day of injection over a period of 4 days to 20oC to induce maturation. Fish were sampled 0, 2 and 4 weeks after injection, visually inspected for color development, and condition factor (CF), gonadal somatic index (GSI) and hepatic somatic index (HSI) were recorded. Na+,K+-ATPase activity was

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measured in gill tissue as a general indicator osmoregulatory status. In 1999 we again rated coloration on a scale from 1 (little or no yellow) to 3 (bright yellow), since ratings were qualitative they were confirmed ‘blind’ by others in the lab. Blood plasma concentrations of T and KT were also measured using previously established radioimmunoassays. Statistical results were calculated using ANOVA, with HSD for unequal cell numbers. A nonparametric analysis employing a Kruskal-Wallis ANOVA by ranks post-hoc test in STATISTICA was used for hormone analysis since a large number of samples were below detection limits. Results from both 1998 and 1999 were fairly consistent. In general p’p-DDE did not affect general indicators of health including weight, HSI, or Na+,K+-ATPase activity (range from 6.2 to 7.0 µmole/mg protein/min). There was, however, a decrease in CF in 1998 in fish dosed with 100 ppm p,p’-DDE (from 1.12 to 1.05), although we did not observe a reduced CF in 1999. Exposure to 100 ppm p,p’-DDE caused a 28% decline in GSI in the maturing male fish in 1999 and a 22% decline (which was not significant) in

Fig. 1. Gonadal (testicular) somatic index (GSI) following four weeks of exposure to p,p’-DDE in Fundulus heteroclitus shown as a percent of mean control GSI (+/- STDEV). Data for studies conducted in 1998 and 1999.

94

1998 (Figure 1). Additionally we observed a decrease in yellow coloration of the males in 1999 (Table 1). Differences in coloration, though not quantified, were noted in 1998 as well. Development of coloration is an androgen dependent process in most male fish often involving KT (reviewed in Borg 1994). The results of this work are consistent with effects of p,p-DDE in the guppy (Baatrup and Junge 2001) where high doses of p,p’-DDE caused reduced coloration, sperm count, gonad size and courtship behavior (1 ppm of p,p’-DDE in food was estimated to result in a daily ‘dose’ of 15 µg/g fish [Baatrup and Junge 2001] or a maximum of 450 ppm over 30 days.) Our exposure levels are closer to the lower dose used by Baatrup and Junge (2001) although our route of exposure was quite different (a single injection at the initiation of the study verses a daily exposure throughout the course of the study). Additionally we measured plasma concentrations of T and KT. Unfortunately too many values for T were below the detection limit of our assay to conduct statistical analysis. KT concentrations did not appear to be altered five weeks after p,p’-DDE exposure (Table 1). Since a large number of samples were below detection limits for T and KT it is difficult to draw conclusions regarding the effects of p,p’-DDE exposure on these androgens. Although by taking the conservative approach of assigning the detection limit to those samples below that limit, our data suggests that exposure to p,p’-DDE did not alter concentrations of KT. Nor did concentrations of KT appear to differ among groups when calculated as ng/ml steroid/g gonad (data not shown). The results of this study suggest that p,p’-DDE exposure in adult fish interferes with gonad growth and development of nuptial coloration in Fundulus, possibly by acting as an antiandrogen.

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Table 1. Effects of p,p’-DDE on morphometric and steroidal endpoints in Fundulus heteroclitus following two or four weeks of exposure data are expressed as mean(stdev).

tmt mg/kg

N Wta g

L cm

HSI %

GSI %

CF KTb ng/ml

Tc ng/ml

Color Score

Results 1998 2 weeks after initial exposure 0 8 5.2

(2.2)A 7.7 (1.0)A

1.0 (0.6)B

4.1 (1.1)B

1.10 (0.04)A

n n n

10 8 5.9 (1.3)A

8.1 (0.5)A

1.3 (0.3)B

3.7 (0.8)B

1.08 (0.04)A

n n n

100 8 4.6 (0.9)A

8.1 (0.8)A

1.5 (0.4)B

4.5 (0.9)B

1.19 (0.04)A

n n n

Results 1999 4 weeks after initial exposure 0 5 4.6

(1.3)A 7.4 (0.8)A

0.9 (0.2)B

3.7 (0.2)B

1.12 (0.03)A

n n n

10 7 5.1 (2.0)A

7.5 (1.0)A

0.9 (0.3)B

3.2 (0.7)B

1.15 (0.03)A

n n n

100 8 4.6 (0.9)A

7.6 (0.5)A

0.8 (0.2)B

2.9 (0.7)B

1.05 (0.02)B

n n n

Results 1999 4 weeks after initial exposure 0 12 10.9

(2.0)A 9.5 (0.5)A

1.7 (0.3)B

2.5 (0.5)A

1.25 (0.04)A

0.46 (0.08)B

0.14 (0.02) A

1.9 (0.7)B

10 11 11.0 (3.1)A

9.5 (0.8)A

1.7 (0.3)B

2.8 (0.5)A

1.26 (0.04)A

0.67 (0.30)B

0.21 (0.11) A

1.9 (0.7)B

100 12 12.9 (3.2)A

9.9 (0.7)A

1.8 (0.4)B

1.8 (0.5)B

1.31 (0.31)A

0.48 (0.17)B

0.16 (0.05) A

1.2 (0.5)A

aDifferent letters are significantly different (p<0.05) using ANOVA and HSD for an unequal N post-hoc test.b KT measurements are for individual fish. Sample detection limit was 0.40 ng/ml to be conservative we used 0.40 ng/ml for samples that were below detection limit. There were several samples below detection limit (0, 5, 4 and 8 in the Time 0, 0 ppm, 10 ppm and 100 ppm groups respectively). Because of the large number of samples below detection we used a nonparametric posthoc test for KT analysis (Kruskal-Wallis ANOVA by ranks) cSamples from 3-4 individuals were pooled for analysis of T, so N=4 per

96

treatment except Time 0 where N=3. Sample detection limit was 0.13 ng/ml, to be conservative we used 0.13 ng/ml for samples that were below detection limit. There were 3, 1 and 3 values below detection in the 0, 10 and 100 ppm groups respectively, so no statistical analysis was conducted. d n=not measured in this study. References Baatrup, E, M Junge. 2001. Antiandrogenic Pesticides Disrupt Sexual

Characteristics in the Adult Male Guppy (Poecilia reticulata). Env Health Perspec 109:1063-1070

Carlson DB, Curtis LR, Williams DE. 2000. Salmonid sexual development is

not consistently altered by embryonic exposure to endocrine-active chemicals. Env Health Perspec 108:249-255

Kelce WR, Stone CR, Laws SC, Gray LE, Kemppainen JA, Wilson EM. 1995.

Persistent DDT metabolite p,p’-DDE is a potent androgen receptor agonist. Nature 375:581-585

Kodric-Brown A. 1998. Sexual dichromatism and temporary color changes in

the reproduction of fishes. Am Zool 38:70-81 Thomas P. 2000. Chemical interference with genomic and nongenomic actions

of steroids in fishes: role of receptor binding. Mar Env Res 50:127-134

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98

BLOOD CELL RESPONSES

OF THE TROPICAL FISH PROCHILODUS SCROFA

TO ACUTE COPPER EXPOSURE AND SUBSEQUENT RECOVERY

Carla C. C. Cerqueira Federal University of São Carlos

Post-Graduate Program in Ecology and Natural Resources C. Postal 676, Phone: 55 16 260-8314

Marisa N. Fernandes

Federal University of São Carlos Department of Physiological Sciences C. Postal 676, Phone: 55 16 260-8314

email: dmnf@power.ufscar.br

Abstract Changes in Prochilodus scrofa blood cells were investigated after 96-h of exposure to copper and following transference to clean water. Hematocrit, red blood cells and hemoglobin concentrations showed a significant increase after copper exposure, remaining high until the 7th day after transference to clean water. The immature blood cells (erythroblasts) also increased significantly, but did not differ from the controls on the 7th day in clean water. The changes in leukocytes occurred only in the percentage of lymphocytes, which was significantly reduced after 96-h copper exposure, remaining lower on the first and second day in copper-free water. Thrombocytes increased significantly in fish exposed to copper and remained high on the 7th day in clean water. The changes in the blood cells of P. scrofa reflect the animals’ responses to stress caused by copper; however, after their transfer to clean water, most of the changes involved a compensatory physiological mechanism that allowed the fish to recover from copper-related damage. Introduction Blood cell responses are important indicators of changes in the internal and/or external environment of animals. In fish, exposure to chemical pollutants can induce either increases or decreases in hematological levels. The growing use of

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copper in the metallurgic industry has resulted in an increase of copper ions in the natural waters of southeastern Brazil. Previous studies of the neotropical fish Prochilodus scrofa exposed to copper revealed drastic changes in red and white blood cells as well as in the thrombocytes (Cerqueira, 2000; Mazon et al., 2003). Few studies, however, have focused on the process of recovery of these cells after copper was removed from water. Hence, the main purpose of this study was to evaluate the effect of acute copper exposure on P. scrofa blood cells and to discover how long the fish took to recover after returning to an environment of improved water quality. Materials and Methods Three to five-month-old juvenile Prochilodus scrofa (W = 15-75 g) were provided by the Hydrobiology and Aquaculture Station of Furnas Hydroelectric Power Plant, Furnas, MG, Brazil, and kept in tanks at 25 ± 1oC (1000 L) with a continuous flow of dechlorinated tap water (water composition: pH = 7.3 ± 0.2; alkalinity = 23.7 ± 1.9 mg L-1 as CaCO3; conductivity = 8.3 ± 0.3 µS and hardness = 24.5 ± 0.2 mg L-1 as CaCO3) and aeration (100% O2 saturation) for at least one month prior to the experiments. The laboratory photoperiod was 12D:12L. The fish were fed with balanced fish food suitable for this species provided by the Aquaculture Research and Training Center - CEPTA/IBAMA. Water temperature, pH, hardness and alkalinity were the same as the mean values found in P. scrofa’s natural habitat (CETESB 1992-2000). After acclimation to laboratorial conditions, the fish were randomly divided into two groups and each group transferred to static test aquariums not exceeding 1g fish L-1. Group 1 (the control group) was kept in copper-free water while group 2 (the group exposed to copper) was exposed to 29 µgCu L-1 (96-h LC50 for P. scrofa; Mazon and Fernandes, 1999). After 96 h, each group was transferred to an aquarium with clean flowing water (recovery period). The physicochemical characteristics of the water in the aquariums of both groups were maintained constant throughout the experimental period and were the same as those prevailing during the acclimation period (except for the copper concentration in the aquarium of group 2). The copper agent used was CuSO4.5H2O and the copper concentration in the water was measured using Atomic Absorption Spectrophotometry. No copper was detected in the water of the control and recovery aquariums.

100

To evaluate the changes in blood cells after exposure to copper and their reversibility following transference to copper-free water, random fish samples (n = 8) from each group, i.e., the control group and the one exposed to copper, were taken after the fish had been held for 96 h in a static system and 1, 2, 7, 15, 30 and 45 days after their transference to clean water. The fish were anaesthetized with 0.01% benzocaine and blood samples were withdrawn from the caudal vein into heparinized plastic tubes. Hematocrit (Hct), red blood cell count (RBC) and hemoglobin concentration ([Hb]) were conducted immediately. Hct was determined by spinning the blood sample contained in heparinized capillary tubes in a microhematocrit centrifuge. The RBC count was carried out in a modified Neubauer chamber after saline dilution of the blood, while the [Hb] was determined by the cyanomethaemoglobin method. The mean corpuscular volume (MCV), mean corpuscular hemoglobin (MCH) and mean corpuscular hemoglobin concentration (MCHC) were calculated from previous blood measurements. Blood smears were fixed with methanol and stained with Leishman solution for counts of immature red blood cell, thrombocytes and leukocytes by 5000 cell count, according to the method described by McKnight (1966). To prevent errors arising from uneven cell distribution, the slides were divided into four segments and cells were counted in fields contained in parallel rows commencing from outside edge of the slide toward the inside. Differential leukocyte counts were made by identifying 200 leukocytes in each slide (Dick and Dixon, 1985). The leukocytes were classified according to their general shape and affinity to the dye (Takashima and Hibiya, 1995). The data are presented as mean ± SEM. The control group data are given all together, since no significant changes were found among them. After the uniformity of the groups’ data was verified using the Bartlett test, the parametric analysis of variance (ANOVA) was applied to determine differences in the level of significance among the groups. Tukey’s test with a 95% confidence limit was applied to compare the mean values whenever a level of significance occurred (GraphPad InStat Software, San Diego, CA). Results No fish from the control group died during the experiment; however, 48% of the fish from the group exposed to copper died during its 96-h exposure. After its

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transfer to aquariums with clean flowing water, no further mortality occurred in this group. Figures 1 and 2 and Table 1 show the changes in the blood cells after 96 h of copper exposure and following transference to clean water. Hct, RBC and [Hb] were significantly higher in fish exposed to copper, particularly on the 1st and 2nd days following transference to clean water (Fig 1). On the 7th day in clean water, RBC and [Hb] were still significantly higher than in the control fish, but significantly lower than they had been immediately following 96 h of exposure to copper. VCM and CHCM increased after 96 h of exposure to copper but were similar to the values of the controls following transference to clean water. The percentage of immature erythrocytes (polychromatophilic and orthochromatophilic erythroblasts) increased after copper exposure, remaining high on the 1st and 2nd days after the transference to clean water (Table 1).

Table 1. Mean values ± SEM of polychromatophilic erytroblasts (PCE) and orthochromatophilic erytroblasts (OCE) of P. scrofa after 96h exposure to

copper and subsequent recovery n PCE

(%) OCE (%)

Control 56 2.7 ± 0.05 1.3 ± 0.08 96h LC50 copper exposure 8 4.5 ± 0.12* 3.8 ± 0.07* Recovery (days) 1 8 3.8 ± 0.06*o 2.1 ± 0.07*o

2 8 3.5 ± 0.08*o 1.9 ± 0.10*o

7 8 2.6 ± 0.13o 1.4 ± 0.07o

15 8 2.5 ± 0.13o 1.4 ± 0.07o

30 8 2.6 ± 0.09o 1.3 ± 0.10o

45 8 2.7 ± 0.09o 1.4 ± 0.08o

* indicates significant difference from the controls (p<0.05) and o indicates significant difference from 96h LC50 copper exposure (p<0.05)

102

0

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(%)

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Cont 96h 1 2 7 15 30 45 Recovery (days)

Cont 96h 1 2 7 15 30 45 Recovery (days)

* * *

*

* * * *

** * *

*

oo

o o o

o o oo o o

o o o

o o o ooo

oo

o

Figure 1. Changes in hematocrit (Hct), red blood cells (RBC), whole blood

hemoglobin concentration ([Hb]), mean cell volume (MCV), mean corpuscular hemoglobin (MCH) and mean corpuscular hemoglobin content (MCHC) of P. scrofa after 96h of copper exposure and subsequent recovery in clean water. The bars represent the mean values (± SEM). Control fish (n = 56; open bars); 96h copper exposed fish (n = 8; black bars) and recovery (n = 8 each time; stippled bars). * indicates significant difference from the controls (p < 0.05); o indicates significant difference from 96h copper exposed fish (p < 0.05)

103

Differential leukocyte counts (Fig. 2A) showed that lymphocytes were the most frequent white blood cells in the control P. scrofa (63 ± 1 %) and the proportion of these cells was reduced to 50% in fish exposed to 29 µgCu L-1. A further significant reduction was found during the first two days of the recovery period and, on the 7th day, the percentage of lymphocytes was similar to that of the controls. The percentage of neutrophils was low compared to monocytes in the control fish. Neither of the cell types showed significant changes after exposure to copper. Basophils were not found in the prepared smears and eosinophils were very rare (less than 0,20 %). The percentage of thrombocytes increased in fish exposed to copper and was found to be similar to the control fish on the 15th day in clean water (Fig. 2B). A B

0

40

80

Cont 96h 1 2 7 15 30 45 Recovery (days)

* * *

ooo o

o o

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*

* *

*

Thro

mbo

cyte

s (%

)

oo

o

oo

o

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)

te

s

ocy

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Le

Figure 2. A. Lymphocyte (open bars), monocyte (black bars) and

neutrophil (stripped bars) percentages of P. scrofa leukocytes after 96h copper exposure and subsequent recovery in clean water. B. Thrombocyte percentage of P. scrofa leukocytes after 96h copper exposure (black bar) and subsequent recovery in clean water (stripped bars); control fish (open bar). The bars represent the mean values (± SEM). Control fish (Cont, n = 56); 96h copper exposed fish (n = 8) and recovery (n = 8 each time). * indicates significant difference from the controls (p < 0.05); o indicates significant difference from 96h copper exposed fish (p < 0.05)

104

Discussion

The direct effects of copper on circulating blood cells were usually associated with an increased disintegration of erythrocytes or, in the case of more sensitive species, to damage of the hemopoietic system (Svobodová et al., 1994). However, some contradictory responses have been found (Wilson and Taylor, 1993; Heath, 1995; Nussey et al., 1995a,b) even in the same species. In P. scrofa, the increase of Hct, RBC and [Hb] with significant changes in the MVC and MCHC blood indices suggests a possible hemoconcentration after copper exposure for 96h. Similar increases of Hct, RBC and [Hb] were also reported by Mazon et al. (2003) in P. scrofa exposed to copper, but the changes were not coupled to changes in the blood indices. During the 7 days of the recovery period, the changes in the red blood cells (increase in the RBC and Hb concentration with no significant changes in the MCH and MCHC blood indices and cell size) suggest a compensatory response of this species to heighten the blood’s O2 carrying capacity. The reduction of the lymphocyte percentage seems to be a general response to metal exposure (Mishra and Srivastava, 1980; Dick and Dixon, 1985; Svobodová et al., 1994) caused by increasing corticosteroid levels in the blood. Neutrophil and monocyte percentages in blood are expected to decrease during acute copper exposure (Svobodová et al., 1994), since these blood cells are vital to protect the body against bacterial infection in damaged tissue. However, no changes were found in P. scrofa either following acute copper exposure in which cell degeneration, rupture and peeling of lamellar epithelial are known to be intense (Mazon et al., 2002), nor during the recovery period. Because thrombocytes are the blood cells involved with blood clotting, their increased percentage in P. scrofa exposed to copper may evidence a compensatory response to reduce bleeding from the damaged branchial vascular tissue. The return of thrombocyte percentages to the levels of the control fish in clean water coincided with the main changes observed in the restoration of gill tissue (Cerqueira et al., 2002). In conclusion, the changes in the blood cells reflect the responses to the effects of stress caused by copper and, after transference to clean water, most of the changes are evidence of compensatory responses that enable fish to recover from copper-related damage.

105

Acknowledgments This research was supported by Fundação de Amparo a Pesquisa do Estado de São Paulo (FAPESP) and Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq), Brazil. C.C.C. Cerqueira acknowledges CAPES for the award of a scholarship. References Cerqueira, CCC 2000 Recuperação do tecido branquial, parâmetros

hematológicos e iônicos de curimbatá, Prochilodus scrofa Steindachner, 1881 (Characiformes, Prochiloontidae) após exposição ao cobre. Ms thesis. Universidade Federal de São Carlos, São Carlos, SP, Brazil, 86 p.

Cerqueira, CCC and Fernandes, MN 2002 Gill tissue recovery after copper

exposure and blood parameter responses in the tropical fish Prochilodus scrofa. Ecotoxicol. Environm. Safety 52, in press.

CETESB 1992-2000 Relatório de Qualidade das Águas Interiores do Estado de

São Paulo. São Paulo, Brasil Dick, PT and Dixon, DG 1985 Changes in circulating blood cells levels of rainbow

trout, Salmo gairdneri Richardson, following acute and chronic exposure to copper. J. Fish Biol. 26: 475-484.

Heath, AG 1995 Water Pollution and Fish Physiology. CRC Press, Boca Raton,

Fl. Mazon, AF and Fernandes, MN 1999 Toxicity and differential tissue

accumulation of copper in the tropical freshwater fish, Prochilodus scrofa (Prochilodontidae). Bull. Environ. Contam. Toxicol. 63: 797-804.

Mazon AF, Cerqueira, CCC and Fernandes, MN 2002 Gill cellular changes

induced by copper exposure in the South American tropical freshwater fish Prochilodus scrofa. Environ. Res. 88: 52-63.

Mazon, AF, Monteiro, EAS, Pinheiro, GHD and Fernandes, MN 2003

Hematological and physiological changes induced by short-term

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exposure to copper in the freshwater fish, Prochilodus scrofa. Braz. J. Biol. 63: in press.

McKnight, IM 1966 A hematological study on the mountain whitefish,

Prosopium williamsoni. J. Fish. Res. B. Can. 23: 45-64. Mishra, S and Srivastava, AK 1980 The acute toxic effects of copper on the blood

of a teleost. Ecotoxicol. Environ. Safety 4: 191-194. Nussey, G, Van Vuren, JHJ and Du Preez, HH 1995a Effect of copper on

haematology and osmoregulation of the Mozambique tilapia, Oreochromis mossambicus (Cichlidae). Comp. Biochem. Physiol. 111(C): 369-380.

Nussey, G, Van Vuren, JHJ and Du Preez, HH 1995b Effect of copper on the

differential white blood cell counts of the Mozambique tilapia (Oreochromis mossambicus). Comp. Biochem. Physiol. 111(C): 381-388.

Svobodová, Z, Vykusová, B and Máchová, J 1994 The effects of pollutants on

selected haematological and biochemical parameters in fish. In Sublethal And Chronic Effects Of Pollutants On Freshwater Fish (Müller, R and Lloyd, R, eds). Fishing New Books, London.

Takashima, F and Hibiya, T 1995 An Atlas Of Fish Histology. Normal And

Pathological Features. 2nd. Ed. Kodansha, Tokyo. Wilson, R and Taylor, EW 1993 The physiological responses of freshwater

rainbow trout, Oncorhynchus mykiss, during acutely lethal copper exposure. J. Comp. Physiol. 163(B): 38-47.

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108

TOXICITY OF CADMIUM: A COMPARATIVE STUDY IN THE

AIR BREATHING FISH, CLARIAS BATRACHUS AND IN

NON-AIR BREATHING ONE, CTENOPHARYNGODON IDELLUS

P.K. Joshi Head of the Department, Dept. of Zoology

Govt. P. G. College. MHOW Dist.Indore (M.P.) INDIA

Phone-91-0731-478204 E-mail – PRAMOD772000@yahoo.com

Manjushree Bose Ram Rahim Colony RAU Dist.Indore (M.P.) INDIA Phone-91-0731-856219

E-mail - manjushree_bose@rediffmail.com Abstract Heavy metals and their salts constitute a very important group of environmental pollutants since they are potent metabolic inhibitors. The inherent toxicity of a metal depends upon its capacity to disturb the dynamic life processes in biological system by combining with cell organelles, macromolecules and metabolites. Cadmium is considered as non-essential element. This study was performed to begin an assessment of effect of the heavy metal on biochemistry of blood serum because blood is a good patho –physiological indicator. Test animals used for the study were Clarias batrachus and Ctenopharyngodon idellus. Both the fishes responded differently to the same toxicant and for same duration of time. In Clarias there was a decrease in glucose, cholesterol, total protein, urea and creatinine value. While in Ctenopharyngodon only glucose and sodium showed a decrease but all the other parameters showed elevation in values. After studying the result of present work it is clear that Cadmium very much affects the energy metabolism, which in long term cause the death of the individual organism and affects the whole community. Introduction It is now well realized that environmental problems have increased exponentially in recent decades mainly because of rapid growth in human population and increased demand for several household materials. While on one hand technological development has improved the quality of life, on the

109

other hand it has created a number of health hazards. The toxic chemicals discharged into air, water and soil get into food chain from the environment. By entering into the biological system they disturb the biochemical processes leading to health abnormalities, in some cases to fetal consequences (Pratima Gupta, 1998). In 1975 U. S. Environmental Protection Agency (U S E P A), Occupational Safety and Health Administration (O S H A), Consumer Product Safety Commission (C P S C) listed 24 extremely hazardous substances. These include heavy metals also. One such important heavy metal is cadmium (Cd). Cadmium is a well-known cumulative poison in animals that belongs to group ΙΙ b of the periodic table. Cadmium enters surface water with the discharge of industrial wastes or by leaching of soil, to which sewage sludge is added. It is biologically very reactive and therefore gives rise to both acute and chronic poisoning. Nariagu (1983) emphasized elaborately on effects of cadmium on aquatic organisms. Many reports are available on the effect of Cd on fish blood. Blood is a good bio indicator or a diagnostic tool to study the problem in organ function. The measurement of biochemical changes in blood of fish under exposure to any toxicant may be used to predict effects upon chronic exposure. Present work was a comparative study with a siluroid air-breathing catfish Clarias batrachus and a cyprinoids non air-breathing fish Ctenopharyngodon idellus under sub lethal Cd intoxication. Effects were studied on some serum biochemical parameters. Materials and Methods Live and healthy Clarias batrachus were purchased from local fish market and Ctenopharyngodon idellus were collected from the pond of village Santer near MHOW. Fishes were checked for injury and disease, and then washed in .1% KMnO4 solution for 5 minutes. After acclimation of 15 days, 42 fishes of each category were selected for experiment, irrespective of their sex. The average length and weight of Clarias batrachus was 17+- cm. and 100+_ gm. The average length of Ctenopharyngodon idellus was 14+_ cm. and weight 100+_ gm. Prior to experiment toxicity tests were conducted to determine the LC50 and safe concentration values of CdCl2 for 96 hours. The physico-chemical analysis of water was done according to Standard Methods published by A.P.H.A. (1992). Both the fishes were divided into 4 equal groups of 6 fishes each. First 3 groups of both the fishes were maintained in sub lethal concentration of CdCl2 separately for 96 hours, 15 days and 30 days. Sub lethal concentration of CdCl2 for Clarias batrachus was 2.5 ppm. and for Ctenopharyngodon idellus was 2.0 ppm. Fourth groups were served as control for respective groups. All control and treated fishes were fed once daily during the tenure of experiment. Both control and treated fishes were sacrificed at time intervals and blood was collected by serving the caudal peduncle using a sharp knife. Serum was separated from

110

the formed elements through the centrifugation at 3000 rpm. for 15 minutes. Seven biochemical parameters were analyzed in serum, are:

1. Glucose (mg/100 ml) - GOD POD method of Trinder. 2. Cholesterol (mgm/dl) - CHOD PAD method of R. Kettermann. 3. Urea (mg%) - DAM method of D. R. Wybenga. 4. Total Protein (gm%) - Biuret method of T. E. Welchelbaum. 5. Creatinine (mg/dl) - Jaffs method of H. P. Seilig and H. West 6. Sodium (mMol/li) - Flame Photometry method of John D. Baur. 7. Potassium (mMol/li) - Flame Photometry method of John D. Baur.

Results During the course of experiments no mortality were recorded in both the types of fishes exposed to sub lethal concentration of Cadmium chloride. Certain changes were observed in the coloration, feeding behavior and activeness of the fishes. Both the types of fishes initially became more active but later their activity ceases. In both the types of fishes coloration fades a little, fluctuating responses were observed in feeding behavior. Table 1. shows the biochemical indices recorded from exposing Clarias batrachus to 2.5 ppm Cadmium chloride for 96 hours, 15 days and 30 days. Differences were measured against the control values determined under controlled laboratory conditions. The value of glucose shows a gradual fall of 54%, cholesterol, total protein, creatinine, urea and potassium values show a regular increase while sodium levels show an initial increase of 11% but at the end of experiment it lowers to .7%. Table 1. Serum biochemical data of the Clarias batrachus exposed to Cadmium chloride (2.5 ppm).

SNO PARAMETER Control 96 hrs 15 days 30 days 1Glucose 47.0 32.3 27.4 21.42Cholesterol 170.0 185.0 192.0 209.03Total protein 4.0 4.2 4.2 5.44Urea 16.300 16.500 17.100 18.4005Creatinine 0.200 0.300 0.400 0.5006Sodium 134.000 149.000 138.000 133.0007Potassium 6.000 7.100 8.100 8.500

Ctenopharyngodon idellus responded slightly different from Clarias batrachus. (Table 2.) In this fish all the parameters gave fluctuating results. Glucose level initially increases up to 29% but later it decreases lower than normal level. Cholesterol, urea and creatinine had a regular decrease up to 66%. Total protein value increases 14% up to 15 days then it lowers to

111

normal in 30 days, while sodium and potassium levels gradually elevated during the total period of experiment. Table 2. Serum biochemical data of the Ctenopharyngodon idellus exposed to Cadmium chloride (2.0 ppm). SNO PARAMETER Control 96 hrs 15 days 30 days

1Glucose 66.6 86.0 72.0 60.02Cholesterol 266.6 210.0 170.0 104.73Total protein 7.0 7.5 8.0 7.04Urea 20.600 20.500 18.300 15.2005Creatinine 0.900 0.600 0.500 0.3006Sodium 112.600 120.000 124.000 129.0007Potassium 9.200 9.800 10.200 11.200

112

Discussion Heavy metals are widely distributed in free water sources and are harmful to aquatic fauna. Biochemical parameters are the best indicators of stress situations caused by heavy metals. In Clarias there was a decrease in glucose value. Cadmium like heavy metals have affinity for ligands like phosphate, cystenyle and histidyl side chains of proteins, can bind with carrier protein molecules resulting in inhibition of sugar and amino acid transport (Alvarado, 1966). According to Passov et al,. (1966) metal ions block the active absorption of glucose by the intestinal epithelial cells. Many other workers reported hypoglycemic condition in air breathing fishes due to contaminants (Kurde1990, Sastry 1984). This may be to cope with high-energy demand in stress situations. Clarias is more active than Ctenopharyngodon, toxicity tests showed that cadmium is more toxic to non-air breathing fishes. (Figure 1) In glucose levels of Ctenopharyngodon, showed initial increase and then a decrease. It may be due to liver impairment to utilize glucose for glycogenolysis (Shastry and Sunita, 1982). Such a situation may be attributed to higher activities of enzymes participating in gluconeogenetic mechanisms, since enzymes of gluconeogenesis are reported to be induced by various toxicants (Shaikh and Hiradhar, 1985).

Glucose

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Figure 1. Alterations in blood glucose levels in both the fishes. Clarias batrachus showed increase in cholesterol value while a slow decrease was observed in Ctenopharyngodon. (Figure 2) According to (Kurde, 1990) 60 – 80 % of total serum cholesterol is in esterified form & esterification occurs mainly in liver. Cadmium damages the liver, proportion of esterified cholesterol decreases. Hyper cholestrolemia observed in Clarias it may be due to impairment of liver and inhibition of enzymes, which converts cholesterol into bile acid (Murrey 1990). Reduced lipoprotein lipase activity plays a role in the increment of plasma lipid (Asha Agrawal and Poonam Sharma, 1999).

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Cholestoral

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Figure 2. Changes in serum cholesterol level. Proteins play a vital role in physiology of living organisms. All biological activities are regulated by enzymes and hormones, which are also proteins. Assessment of protein content can be considered as a diagnostic tool to determine the physiological phases of the cells. (Kapila Manoj 1999) Cadmium competes with Zn for the same sulphahydral group and binds more firmly. Proteins are too sensitive and early indicators of heavy metal poisoning. Kapila Manoj, 1999 & Kurde, 1990 observed elevation in protein content of rat serum due to textile mill effluents. B.Rajanna et al,. (1981) reported enhancement in protein content due to cadmium. (Figure 3) The increase in protein content was due to enhancement of microsomal protein synthesis suggested by many workers. Kidney is the target organ of cadmium poisoning.

Total Protein

02468

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30days

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Figure 3.Changes in serum total protein level. Teleost fishes are primarily amminotelic but their blood contains significant amount of urea and indeed in some teleosts it may account for 20 % or more of the total nitrogen excreted. Occurrence of uremia was reported by many workers (Gupta and Bhargava 1985, Kurde 1990). Renal disorders also

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elevate serum urea values. The level of urea was influenced by protein content of diet.

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Figure 4. Changes in serum urea levels. A high protein diet raises the serum urea and low protein diet lowers it. (Figure 4) Clarias is an active fish in stress condition it shows higher activity while Ctenopharyngodon become sluggish and doesn’t show more activity, this type of expression in serum urea level was may be due to their feeding habits. Creatinine is another nitrogenous waste product that is eliminated by the kidneys, when excretion is suppressed in renal insufficiency. The value is unaffected by protein intake. According to Lall et.al,. (1997) rise in creatinine value is an indication of renal tubular damage due to cadmium-induced naphrotoxicity (Kazuo et al,. 1980) Figure 5. Shows the regular increase in serum creatinine value in Ctenopharyngodon it proves that Cadmium is more nephrotoxic to non-air breathing fishes.

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Figure 5. Alterations in serum creatinine value. Minerals are mainly responsible for the maintenance of osmotic pressure in blood and proper function of all types of tissues. (Mohanty & Mishra 1983). The presences of activator ions of alkali metal series (Na+, K+) are essential

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for activity of many enzymes. Toxic metals can alter the concentration of electrolytes in blood. ATP and its related systems have been documented well to participate in several metabolic processes. Na+ -K+ ATPase, located in the cell membrane, has been implicated in the active transport of Na+ and K+ across the cell membrane (B.Rajanna et al., 1981); changes in the levels of plasma ions in present study were due to gill damage and inhibition of

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Figure 6. Changes in blood sodium level.

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Cotrol 96 hrs 15 days 30 days

ClariusIdellus

Figure 7. Changes in blood potassium level. enzyme activity. Woodling (1999) suggested that plasma ion decrease or increase is due to kidney damage and altered enzyme activity and it has been an indicator of impending death. Changes in the value of sodium in both the fishes was may be due to the structural difference in gills and tolerance to toxicant. Through reviewing the available literature it may be due to lateral line imbalance and hormonal disorder by affected endocrine organs through heavy metals.

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Conclusion After the above discussion it had been concluded that heavy metals causes deleterious effects on fishes and very much altars the biochemical characteristic of blood. In sub lethal concentration it may not be fetal for an individual organism but it does affect the growth rate and reproduction resulting in disturbance to whole community and tropic levels of food chains, ultimately the ecosystem. References Alvarado, F.1966.Transport of sugars and amino acids in the intestine.

evidences for a common carrier. Science.151: 1011-1012. A.P.H.A., A.W.W.A.and W.P.C.F. 1970. Standard methods for the

examination of water and wastewater. 18th Ed. American Public Health Association. Washington.

Asha Agrawal and Poonam Sharma. 1999. Effect of sulphur dioxide on total

lipid and cholesterol level in the blood of albino rats. Journal of Environ.Biol.20 (4). 335-338.

Gupta, R.C. and S. Bhargava. 1985. Practical Biochemistry. CBS Publishers

And Distributors, Delhi (India). Gupta Pratima. 1998. Cadmium toxicity and thyroid function with special

reference to 5’-monodeiodinase enzyme activity a comparative study in birds and mammal. Ph.D. Thesis.

Kapila Manoj and G.Raghothaman. 1999. Mercury, copper and cadmium

induced changes in the total protein level muscle tissue of an edible estuarine fish Boleophthalmus dessumieri. Cuv.J.Envi. Biol.20 (3), 231-234.

Kazuo T.Suzuki, Mitsuru Yamamura, Yasuko K. Yamada and Fujio

Shimizu. 1980. Decreased copper content in rat kidney Metalothionine and its relation to acute cadmium toxicity. Toxicology Letters, 7.137-142.

Kurde Sushama.1990. Effect of textile mill effluents and dyes on the

heamatological parameters in albino rats. Ph.D.Thesis. Lall, S.B., N.Das, R.Rama, S.S.Peshin, S.Khatter, K.Gulati and S.D.Seth.

1997. Cadmium induced nephrotoxicity in rats. Indian. J. Exp. Biol. Vol 35, pp 151-154.

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Mohanty, B.K. and B.N.Mishra. 1983. Effect of mercurial drug (Kajyoli) on

Albino rat blood.J.Envi.Biol. 4 (4) 201-206. Murray,R.K. 1991.Harpers Biochemistry 22 nd edition Prentice Hall

International Inc. pp-678. Nariagu, O.Jerome and John.B.Sprangu. 1983. Cadmium on the aquatic

environment. Vol.9, A. Wietly Interscience Publication. Rajanna, B., K.D.Chapatwala, D.D.Vaishnav and D.Desaiah.1981.Changes

in ATPase activity in tissues of rat fed on cadmium. J. Envi. Biol. 2(1) 1-9.

Sastry, K.V. and Sunita. 1983. Alterations in the intestinal absorption of

Xylose induced by heavy metals in fresh water teleost fish Channa punctatus. Poll.Res.Vol. 2(2): 45-48.

Sastry and Sunita. 1982. Effect of cadmium chromium on the intestinal

absorption of glucose in snake head fish Channa punctatus.Toxicology Letters,10.293-296.

Shaikh, Y.A. and P.K.Hiradher. 1985. Fluoride induced changes in blood

Glucose, tissue glycogen and succinate dehydrogenase (SDH) Activity in the mudskipper Beleophthalmus dessumeiri. Proc. Symp. Assess. Envir. Pollu, 93-99.

Woodling, John. D. 1999.Physiological and weight changes of wild brown

Trout inhabiting water with acutely toxic cadmium and zinc concentration : An in situ study. Proc.of World Fish Convention.

Acknowledgement Authors wish to express their humble gratitude to Dr. Lokesh K. Agrawal for providing facilities for present work.

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EFFECTS OF VANADATE OLIGOMERS

ON LIPID PEROXIDATION AND ANTIOXIDANT ENZYMES

IN THE LUSITANIAN TOADFISH KIDNEY AND LIVER:

SHORT-TERM EXPOSURE

Inês Figueiredo 1

1Group of Comparative Cardiovascular , CCMar, Faculty of Marine and Environmental Sciences, University of Algarve, Campus de Gambelas,

8000-810 Faro, PORTUGAL Phone: 351 89 800 900; Fax: 351 89 818353

E-mail: inesu2@hotmail.com

Sandra Soares1; Gisela Borges1; Natércia Joaquim1; Manuel Aureliano2; Josefina Coucelo1

2CMQA, Chemistry Dept., FCT, University of Algarve

EXTENDED ABSTRACT ONLY - DO NOT CITE

Introduction

Vanadium is a transitional metal to which a special attention is given in questions of environmental management and health (Nriagu, 1998). Several animal studies associate vanadium with oxidative stress and pointout liver and kidney as major targets of metal toxicity (Stohs and Bagchi, 1995). As well as other toxic metals, vanadium is known to exhibit the ability to produce reactive oxygen species, resulting in lipid peroxidation and antioxidant enzymes alterations, namely superoxide dismutase, catalase and glutathione peroxidase (Byczkowski and Kulkarni, 1998). However, the contribution of vanadate oligomers, in this case “meta” and “decavanadate”, for vanadium toxicity in these tissues is not clarified. Thus, the objective of this work was to evaluate antioxidant defence system responses induced by an acute exposure to a sub-lethal concentration (5mM) of “meta” and “decavanadate”, on the kidney and liver of Halobatrachus didactylus (Lusitanian toadfish).

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Methodology

The H. didactylus individuals were collected from Ria Formosa (South coast of Portugal) and divided in three groups: Control (CTRL), injected intraperitoneously (i.p.) with 0.9% NaCl; Metavanadate (Meta V), injected i.p. with 1 ml/Kg of “metavanadate” (5mM); Decavanadate group (Deca V), injected i.p. with 1 ml/Kg of “decavanadate” (5mM) Subgroups of 3 individuals were sacrificed 0, 1 and 8 days after intoxication.

Liver and kidney were collected after sacrifice and cytosolic and mitochondrial fractions were prepared for determination of catalase (CAT), superoxide dismutase (SOD) and glutathione peroxidase (Total GPx and Se-GPx) activities. Lipid peroxidation products were determined in homogenates using TBA method.

The results are shown in percentual variation of group averages, compared with CTRL group.

Results and Discussion

Different effects for both vanadate solutions in liver and kidney, were observed. In the kidney (Table 1), antioxidant enzymes activities and lipid peroxidation increased both in Meta V and Deca V groups. Major alterations occured in Deca V CAT cytosolic, after 8 days, SOD mitochondrial, after 24 hours and Se-GPx activities (Table 1). Also, there was a significant increase in lipid peroxidation on Deca V group, which indicates an ineffective response of the cellular defence mechanisms against oxidative stress caused by this metal. The same study applied to the cardiac muscle (Aureliano et al., 2002) also revealed significant changes in antioxidant enzymes activities and lipid peroxidation, indicating that decameric vanadate species induce stronger toxic effects than other vanadate species.

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Table 1. Percentual variation of kidney antioxidant enzymes activity and lipid

peroxidation, after 24 hours and 8 days of “meta” or “decavanadate” exposure.

% of Variation 24 Hours 8 Days Parameter Fraction

Meta V Deca V Meta V Deca V Cytosolic 70.9 39.2 -57.8 295.0 CAT Mitochondrial 23.0 0.5 -42.6 -53.2 Cytosolic -29,6 -48.1 -18.3 33.1 SOD Mitochondrial 24.1 139.2 65.2 85.0

Total 9.1 43.1 4.2 64.7 GPx Se-GPx 11.6 115.3 70.0 137.0 TBARS Total 61.4 8.8 34.7 257.3

In the liver (Table 2), CAT and SOD activities were in general stimulated in both groups. Deca V group, after 24 hours, has shown the highest difference in comparison to CTRL group (139.4%). There were no significant alterations in lipid degradation products. These results indicate that, in the liver, the antioxidant enzymes play an important role against oxidative stress. Table 2. Percentual variation of liver antioxidant enzymes activity and lipid

peroxidation, after 24 hours and 8 days of “meta” or “decavanadate” exposure

% of Variation 24 Hours 8 Days Parameter Fraction

Meta V Deca V Meta V Deca V Cytosolic 7.9 23.7 55.3 19.9 CAT Mitochondrial 14.1 139.4 -44.9 11.0 Cytosolic 66.8 13.0 21.9 6.8 SOD Mitochondrial 18.8 10.6 6.3 9.5

Total -34.1 -16.1 -36.7 58.0 GPx Se-GPx -8.2 -19.9 -20.4 -33.5 TBARS Total -42.1 -30.1 -49.5 -33.6

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A similar study with cadmium in the heart, kidney and liver (Coucelo et al., 2000) also report an increase of CAT and SOD activities in the liver. The antioxidant enzymes activities in kidney had the same pattern, except in CAT activity, that decreases after 24 hours and an increase after 7 days.

All oligomeric species of vanadate studied induced oxidative stress in both tissues, but have also shown to affect differently antioxidant enzymes activities and lipid peroxidation. Apparently, “decavanadate” induces stronger antioxidant responses than “metavanadate” and stronger effects, as well as lipid peroxidation, in the kidney.

References

Aureliano, M., N. Joaquim, A. Sousa, H. Martins and J.M. Coucelo. 2002. Oxidative stress in toadfish (Halobactrachus didactylus) cardiac muscle: acute exposure to vanadate oligomers. J. Inorg. Biochem. 90: 159-165

Byczkowski, J.Z. and A.P. Kulkarni. 1998. Oxidative stress and pro-oxidant biological effects of vanadium. In "Vanadium in the environment. Part 2: Health Effects”. John Willey & Sons, Inc. N.Y. pp. 235-264

Coucelo, J. M., N. Joaquim, V. Correia, M.J. Bebianno and J.A. Coucelo. 2000. Cellular responses to cadmium toxicity in the heart, kidney and liver of Halobatrachus didactylus. Ecotox. Environ. Rest. 3: 29-35

Nriagu, J. O. (1998). History, occurence and uses of vanadium In "Vanadium in the environment. Part 1: Chemistry and Biochemistry”. Willey & Sons, Inc. , N.Y. pp. 1-22

Stohs, S. J. and D. Bagchi. 1995. Oxidative mechanisms in the toxicity of metal ions. Free Rad. Biol. Med. 18: 321-336

INCUBATION AND PH-DEPENDENT EFFECTS

OF VANADATE OLIGOMERS AND CADMIUM

WITH Halobatrachus didactylus SARCOPLASMATIC RETICULUM

CALCIUM PUMP

Lília Leonardo Biochemistry Lab., CMQA, Chemistry Dept., FCT

University of Algarve, Campus de Gambelas, 8000-810 Faro, PORTUGAL Phone: +351 89 800 900; Fax: +351 89 818 353

E-mail: lilialeonardo@hotmail.com

S.S. Soares1, N. Joaquim1, J.M. Coucelo. 1 and M. Aureliano2 1 Group of Comparative Cardiovascular , CCMar, Faculty of Marine and

Environmental Sciences 2 Biochemistry Lab., CMQA, Chemistry Dept., FCT

University of Algarve, Campus de Gambelas, 8000-810 Faro, PORTUGAL

EXTENDED ABSTRACT ONLY – DO NOT CITE

Introduction The Ca2+-ATPase of sarcoplasmatic reticulum (SR) membrane is a transmembranar protein that catalyses both the hydrolysis and synthesis of ATP (Chini et al; 1993) playing an essential role in relaxation and contraction of the muscle. The activity of Ca2+-ATPase is normally described by a cycle postulating the existence of two protein conformations, E1 and E2, with high and low affinity for Ca2+ and ATP, respectively (Inesi, 1985). The alternation between the two distinct conformations of Ca2+-ATPase during the transport cycle seems affect the interactions with different oligomeric species of vanadate (e,g., decameric and tetrameric) that affects the activity of the calcium pump (Aureliano and Madeira, 1998). Such as, vanadium, cadmium is well known by its toxic effects in live organisms, being an inhibitor of the calcium pump even at very low concentrations.

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In this work we compare the effects of cadmium and vanadium on the ATP hydrolysis by the calcium pump at different pH values and incubation times from skeletal muscle of Halobatrachus didactylus (Schneider, 1801). Material and Methods Ca2+-ATPase isolation and characterization SR vesicles derived from H. didactylus – Lusitanian toadfish – skeletal muscle were prepared as described elsewhere (Aureliano and Madeira, 1994). Metal stock solutions and vanadate stability Cadmium stock solution (50 mM) was prepared from cadmium chloride. Vanadate stock solutions (50 mM) – “metavanadate” and “decavanadate” – were prepared from ammonium metavanadate, according to described elsewhere (Aureliano and Madeira, 1994). The stability of vanadate solutions at the different experimental conditions was analysed by measuring the absorbance at 400 nm. Hydrolysis of ATP by calcium pump ATP hydrolysis was measured by colorimetry, through inorganic phosphate analysis. Experiments were proceeded at 25ºC in a reaction medium containing 0.1 M KCl, 25 mM HEPES, 5 mM MgCl2, 50 mM CaCl2, pH 6.0, 7.0 or 8.0, in the absence and presence of 2 µM of cadmium or 2 mM (total vanadium) of “metavanadate” or “decavanadate”, with or without 1 hour incubation with 0.285 mg/ml protein. The reaction was started by the addition of 420 mM Mg-ATP. Results pH effects on the vanadate solutions stability The “decavanadate” solutions stability decreases with pH (pH 6.0>7.0>8.0) whereas “metavanadate” solutions are stable.

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pH effects on Ca2+-ATPase activity in vanadium or cadmium inhibition The ATP hydrolysis by the SR Ca2+-ATPase is more strongly affected by “decavanadate” than “metavanadate” and it depends on the pH. The relative inhibitory effects for “metavanadate” and “decavanadate” regarding pH was: pH 6.0>8.0>7.0, whereas for cadmium the ATP hydrolysis was more inhibit at pH 8.0 (pH 8.0>7.0»6.0) (Figure 1).

deca 2 mM

0

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80

100

pH 6.0 pH 7.0 pH 8.0

Inhi

bitio

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Figure 1. Effects of pH on the inhibition of calcium pump ATPhydrolysis by “metavanadate”, “decavanadate” 2 mM orcadmium 2 µM (n=4).

cádmio 2 mMcadmium 2 µM

Effects of Ca2+-ATPase incubation with vanadium or cadmium After 60 minutes of incubation “decavanadate” inhibition on ATP hydrolysis is always higher than “metavanadate”. For both vanadate solutions the relative order of inhibition upon pH is: pH 6.0>8.0>7.0. In fact, the lowest inhibition is observed at pH 7.0 where “metavanadate” and “decavanadate” affect by 49% and 80%, respectively the ATPase activity. Unlike vanadium, the protein incubation with cadmium increases the inhibition of ATP hydrolysis within pH increase from 3% to 24% (Figure 2).

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0

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deca 2 mM

Figure 2. Effects of pH on the inhibition of calcium pump ATPhydrolysis by “metavanadate”, “decavanadate” 2 mM orcadmium 2 µM after 1 hour of incubation (n=4).

Conclusions It is concluded that, at pHs 6.0, 7.0 and 8.0, cadmium inhibits strongly calcium pump ATP hydrolysis and the cadmium incubation with the protein favours enzymatic inhibition. “Decavanadate” affects more strongly the ATP hydrolysis by Ca2+-ATPase than “metavanadate”, being the relative order of inhibition affected by pH as followed: pH 6.0>8.0>7.0. It is suggested that oligomeric species of vanadate inhibition of ATP hydrolysis is favoured by the E2 conformation (pH 6.0). Further studies are in course to clarify contribution of vanadate species to vanadium interaction with the calcium pump of Halobatrachus didactylus. Acknowledgements Sandra Soares thanks to Portuguese Foundation for Science and Technology, FCT and by POCTI program financed through FEDER. Research project 38191/QUI/2001 and University of Algarve.

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References Aureliano, M. and V.M.C. Madeira. (1994). Interactions of vanadate oligomers

with sarcoplasmic reticulum Ca2+-ATPase. Biochim. Biophys. Acta 1221: 259-271

Chini, E.N., F.G.S. de Toledo, M.C. Albuquerque and L. de Meis (1993). The

Ca2+-transporting ATPases of rabbit and trout exhibit different pH- and temperature-dependences. Biochem. J. 293: 469-473

Inesi, G. (1985). Mechanism of calcium transport. Annu. Rev. Physiol. 47: 573-604

M. Aureliano and V. M. C. Madeira (1998) Vanadium in the Environment. Part

1: Chemistry and Biochemistry, Wiley & Sons, Inc., New York, pp. 333—357

127

128

HISTOLOGICAL ANALYSIS OF VANADATE OLIGOMERS

EFFECTS ON HEART, KIDNEY AND LIVER

OF THE LUSITANIAN TOADFISH:

AN ACUTE EXPOSURE STUDY

Gisela Borges1

1Group of Comparative Cardiovascular , CCMar, Faculty of Marine and Environmental Sciences, University of Algarve, Campus de Gambelas,

8000-810 Faro, PORTUGAL Phone: +351 89 800 900; Fax: +351 89 818353

E-mail: gisela_borges@megamail.pt

Paula Mendonça1; Natércia Joaquim1; Manuel Aureliano2; Josefina Coucelo1

2CMQA, Chemistry Dept., FCT, University of Algarve

EXTENDED ABSTRACT ONLY - DO NOT CITE Introduction The transitional element vanadium is an anthropogenic toxic metal produced by the burning of fossil fuels (Barceloux, 1999). In trace concentrations vanadium may exert benefit effects (1-10 nM), although in higher concentrations it becomes toxic (>100 mM) (Stohs and Bagchi, 1995; Crans et al, 1998).Vanadium toxicity has been associated mainly to oxidative stress in kidney and liver (Stohs and Bagchi, 1995), but several effects on cardiovascular system were also reported. However, histopathological effects of vanadium on these tissues are not clarified. Therefore, the objective of this work is to study the effects of an acute exposure to a sub-lethal concentration (5 mM) of two different vanadate solutions in cardiac, renal and hepatic tissues of the Lusitanian toadfish Halobatrachus didactylus (Schneider, 1801). It is also our objective to verify if these histopathological effects depend on the vanadate oligomers administered: “metavanadate” or “decavanadate”. Material and Methods Specimens of H. didactylus were divided in three groups: Control, Meta and Deca groups, consisting of 10 individuals each. The Meta and Deca groups were injected with 1ml/Kg of a “metavanadate” solution (5 mM), containing

129

mainly metameric species and a “decavanadate” solution (5 mM) containing mainly decameric species, respectively. Five specimens of each group were sacrificed 1 and 7 days after the exposure. The organs were collected and the hearts were weighted in order to calculate the relative ventricular mass (RVM). Sections of ventricular, renal and hepatic tissues were subjected to histological procedures, stained with Haematoxylin-eosin and examined through light microscopy in order to evaluate tissue alterations induced by vanadium. Ventricular sections were also stained with Picrosirius and observed through bipolarising microscopy to calculate the ventricular wall structural elements area fraction (percentage): collagen type I (coll. I), collagen III (coll. III) and muscular tissue, as described previously (Coucelo et al., 2000). Mann-Whitney test, a non-parametric variance test with a significance level of 0.05 was applied to compare contaminated groups with control. Results and Discussion In the heart, 1 and 7 days after exposure to “metavanadate” the RVM was not affected (p>0.05), while “decavanadate”, induced a significant increase at day 7 (p<0.05), relatively to the Control group. Although no evident cardiac tissue lesions were observed in the contaminated groups, the study of the cardiac tissue components area fractions revealed significant alterations (p<0.05) in the myocardial tissue organization. “Metavanadate” induced a decrease of coll. III at day 1 and a decrease of coll. I after 7 days, while “decavanadate” decreased coll. I and III, just after 1 day. The muscle tissue area fraction was increased at day 1 by “metavanadate”, while “decavanadate” augmented this component after 1 and 7 days of exposure (Figure 1). The observed results indicate that, in general, there was a collagen decrease and a muscle tissue increase. Studies indicate that a collagenase activation and fibrillar collagen breakdown are responsible for the dilation, the change in shape and the increase in distensibility of the cardiomyopathic left ventricle (Caulfield and Janicki, 1997), which is in concordance with our results. However, “metavanadate” contrarily to “decavanadate” did not induce RVM increase, although a collagen decrease was registered.

130

50.84 ± 0.945 *

012345

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)

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)

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0.33 ± 1.12 0.33 ± 0.14

0.23 ± 0.19 * 0.22 ± 0.08

*

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1.70 ± 0.75

2.49 ± 1.59 1.68 ± 2.50

* 0.70 ± 0.23

*

50.84 ± 0.945 *

a) b)

Figure 1. Cardiac tissue components area fraction (%) in the images containing the myocardial region: a) collagen type I; b) collagen type III; c) muscle tissue; * p<0.05.

In the kidney of contaminated individuals, the renal tubules presented no lumen and showed complete disorganization and different necrosis states of epithelial cells (picnosis and karyolysis). The interstitial tissue was

131

132

hypercromatic. At day 7, these features were more severe than at day 1, although “metavanadate” and “decavanadate” produced similar effects. The hepatic tissue alterations included hypertrophied nuclei and hepatocytes, necrotic hepatocytes and diminished cytoplasmic content. “Metavanadate” and “decavanadate” induced more severe effects at day 7, and the effects due to “decavanadate” were also more severe. Conclusion The studied oligomeric species of vanadate promoted evident tissue lesions in the kidney and liver but not in cardiac tissue. However, “decavanadate” induced a dilatation of the ventricle due to a decrease in the myocardial collagen fibers percentage area, which may be related to ventricular dysfunction. In general, “decavanadate” induced stronger histopathological changes than “metavanadate”. Acknowledgements Gisela Borges was supported by a research grant from UIC (Unidade de Intervenção Cardiovascular , Hospital Particular do Algarve). References Barceloux, D. 1999. Vanadium. Clin.Toxicol. 37 (2): 265-278 Coucelo, J.M., N. Joaquim, V. Correia, M.J. Bebianno and J.A. Coucelo.

2000. Cellular responses to cadmium toxicity in the heart, kidney and liver of Halobatrachus didactylus. Ecotox. Environ. Rest. 3 (1): 29-35

Crans, D., S. Amin and A. Keramidas 1998. Chemistry of relevance to

vanadium in the environment in Nriagu, J. (Ed.). Vanadium in the environment. Part 1: Chemistry and biochemistry. John Wiley & Sons, Inc. New York. pp. 73-95

Caulfield, J. B. and J. S. Janicki. 1997. Structure and function of myocardial

fibrillar collagen. Technol. Health Care. 5: 95-113 Stohs, S.; Bagchi, D. (1995). Oxidative mechanisms in the toxicity of metal

ions. Free Rad. Biol. Med. 18: 321-336

CADMIUM AND VANADATE OLIGOMERS COMPARATIVE

EFFECTS ON THE TOADFISH ERYTHROCYTE

S.S. Soares1

1Group of Comparative Cardiovascular , CCMar, Faculty of Marine and Environmental Sciences, University of Algarve, Campus de Gambelas, 8000-

810 Faro, PORTUGAL Phone: +351 89 800 900; Fax: +351 89 818 353

E-mail: soares_ss@hotmail.com

M. Aureliano2, N. Joaquim1, J. M. Coucelo1

2CMQA, Chemistry Dept., FCT, University of Algarve

EXTENDED ABSTRAT ONLY – DO NOT CITE

Introduction

The increasing levels of toxic metals in the environment, specially in marine environments, due to anthropogenic activities over the last years, make it important to study its toxicity mechanisms. Heavy metals, such as cadmium and vanadium, are known to cause extremely harm on biological systems. Cadmium presents a high potential as a toxic substance, even in low concentrations (Hu, 2000). On contrary, vanadium is considered benefit to live organisms in vestigial concentrations (1-10 nM), although it becomes toxic in higher concentrations (>100 µM) (Harland and Harden-Williams, 1994). Several haematological changes, compromising oxygen transport efficacy (cell destruction, hemolytic anemia and methaemoglobinemia), have been described upon metal intoxication. In teleost species this is specially important since 10% of haemoglobin is in the form of methaemoglobin (Lewis and Morris, 1986). In the present study, biochemical and morphological in vivo toxic effects of cadmium and vanadium were evaluated in red blood cells from Halobatrachus didactylus (Schneider, 1801), in order to: 1) compare cadmium and vanadium effects in methaemoglobin reductase activity; 2) evaluate the contribution of different vanadate oligomeric species to vanadium toxicity; 3) study cadmium and vanadium interaction with haemoglobin and 4) determine morphological changes induced in red blood cells by cadmium and vanadium.

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Material and Methods

H. didactylus – Lusitanian toadfish – individuals were collected in Ria Formosa lagoon (Portuguese south coast) and divided into five groups: Control 1, non injected; Control 2, injected intraperitoneously (i.p.) with 0.9% NaCl (placebo); Cd, injected i.p. with 5 mM of Cd (CdCl2); Meta, injected i.p. with a “metavanadate” solution containing 5 mM of total vanadium; and Deca, injected i.p. with “decavanadate” containing 5 mM of total vanadium. Metal solutions were diluted into final concentration (5 mM) in 0.9% NaCl. “Metavanadate” and “decavanadate” solutions were prepared from ammonium metavanadate, according to described elsewhere (Aureliano and Madeira, 1994). All solutions were administrated in a dosage of 1 mL solution/Kg of body weight. Subgroups of 4 individuals of each group were sacrificed after 1 and 7 days. Blood samples were collected using heparin as an anticoagulant.

Methaemoglobin reductase activity was determined by Board (1981) method.

Cadmium and vanadium interactions with haemoglobin were studied in red blood cells haemolysates, from a group of 6 non-injected fishes. The obtained supernatant was incubated with metal concentration ranging from 0.5 to 5 mM, at 25 ºC, up to 90 minutes and haemoglobin spectrum changes were analyzed by spectroscopy.

Erythrocyte number, haemoglobin level and haematocrit were determined, in blood samples, by routine methods. Morphological changes, induced by cadmium and vanadium on red blood cells were studied on haematological preparations with Giemsa stain.

The Mann-Whitney test was applied to test differences between groups, on all the parameters analysed. The significant level used was p <0.05. Control 1 and Control 2 showed no significant differences and for result analysed were considered together as Control.

Results

Cadmium and vanadium effects on methaemoglobin reductase activity

It was observed a 20% decrease, relatively to Control, in methaemoglobin reductase activity 1 day after cadmium injection, while “metavanadate” induced

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a 15 to 67% increase. “Decavanadate” had no significant effects on methaemoglobin reductase activity (Figure 1).

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Figure 1. H. didactylus methaemoglobin reductase activity variation1 and 7 days after 5 mM cadmium or vanadium (“metavanadate” or“decavanadate”) in vivo administration (n=4). Standard deviationranging from 2 to 25 µmol NAD+/min/g Hb.

Cadmium and vanadium interactions with haemoglobin

A 0.5 mM “decavanadate” solution (which contains only 50 µM of decameric species) induced haemoglobin oxidation to methaemoglobin after 30 minutes

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incubation, whereas a slight spectrum change was detected for 5 mM of “metavanadate” and 5 mM of cadmium does not affect haemoglobin spectrum (Figure 2).

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Figure 2. A: a) Haemoglobin and b) methaemoglobin spectra; B:Haemoglobin spectra after 90 minutes of incubation with a) Cd 5mM, b) Meta 5 mM and c) Deca 50 µM (0.5 mM total vanadium), at 25 ºC in phosphate buffer 0.5 M pH 7.2.

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Cadmium and vanadium effects on haematological parameters

After 7 days of metals administration distinct changes were found in the haematological parameters. Only “decavanadate” solution induced a decrease in erythrocytes count and haematocrit relatively to Control – 14% and 31%, respectively, whereas the haemoglobin concentration decreased 29, 22 and 36% in Cd, Meta and Deca groups, respectively, as it was observed an increase of erythrocytes cellular volume. In Cd group, the cell hypertrophy was followed by a nuclear volume increase.

Conclusions

It is concluded that different vanadate oligomers contributed differently to vanadium toxicity in H. didactylus. Stronger effects were observed for decameric species in causing haematological changes in the toadfish erythrocyte. Upon administration, apparently cadmium induced an enzymatic activity decrease of the methaemoglobin reductase in vivo, while “metavanadate” solution promoted its stimulation and “decavanadate” had no effect. Vanadate species seems to be more effective causing haemoglobin oxidation in vitro than cadmium and, among vanadate species, “decavanadate” promoted a stronger effect in this process. Moreover, cadmium and vanadate oligomers induce a reduction of haemoglobin concentration, as well as the increase of erythrocytes cellular volume. However, decameric species of vanadate exhibit a more deleterious effect, promoting the reduction of erythrocytes count and hematocrit.

Acknowledgements

S.S. Soares thanks to Portuguese Foundation for Science and Technology, FCT and by POCTI program financed through FEDER. Research project 38191/QUI/2001 and University of Algarve.

References

Aureliano, M. and V.M.C. Madeira. (1994). Interactions of vanadate oligomers with sarcoplasmic reticulum Ca2+-ATPase. Biochim. Biophys. Acta 1221: 259-271

Board, P.G. 1981. NADH-ferricyanide reductase, a convenient approach to the evaluation of NADH-methaemoglobin reductase in human erythrocytes. Clin. Chim. Acta 109: 233-237

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TOXICITY OF LAKE ENRICHMENT NUTRIENTS

TO AQUATIC LIFE

Don MacKinlay, Fisheries & Oceans Canada

555 West Hastings St. Vancouver V6B 5G3 Canada E-mail: mackinlayd@pac.dfo-mpo.gc.ca

Craig Buday,

Pacific Environment Laboratory, Environment Canada

Introduction The Lake Enrichment Program adds nutrients to lakes in British Columbia to increase their productivity as nurseries for sockeye salmon fry (Stockner, 1987). Since 1985, the nutrients used have been a mixture of urea ammonium nitrate (32-0-0 or 28-0-0) and ammonium polyphosphate (10-34-0), both of which are supplied in concentrated liquid form. Before 1997, these nutrients were added to the lakes by spraying a fine mist onto a designated 'application zone' from aircraft (twin engine airplanes or helicopters) fitted for crop dusting. Currently, nutrients are added by introducing the liquid nutrient mix directly into the propeller wash from a boat cruising on the lake. The goals of this study were to: 1. Determine the toxicity of the concentrated nutrient solutions to aquatic life, 2. Determine which chemical species is the toxic fraction, and 3. Compare the toxicity levels to the concentrations that would be found

during an application of nutrients to the lake. Methods Toxicity Tests Toxicity testing was carried out at the Pacific Environmental Science Centre in North Vancouver, Canada in 2001. 96-hr acute lethality tests were performed on

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rainbow trout (Oncorhynchus mykiss) for both the 28-0-0 and the 10-34-0 solutions. A 48-hr a acute lethality test was performed on the freshwater microcrustacean, Daphnia magna (Figure 1). The 96 hour LC50 is the concentration of sample that is calculated to be lethal to 50% of the test fish over an exposure period of 96 hours.A 72-hr IC50/IC25 growth inhibition test was performed on the green alga, Selenastrum capricornutum using the 10-34-0 solution.

Figure 1. Daphnia magna, a test organism for freshwater aquatic life

toxicity studies. The rainbow trout LC50 tests followed protocols outlined in McLeay (1990, 2001) and the Dapnia LC50 test followed protocols outlined in (Miller et al. 2000). Both tests were conducted using well water as the contol and the diluent for the test concentrations. Five test concentrations were used: 5600, 3200, 1800, 1000, 560 and 320 mg/L and a control. Each concentration had one replicate with 10 test organisms per replicate in 30 kg of test solution for the Rainbow trout and 200 mL for the Daphnia. Rainbow trout mortality was recorded after 5, 10, 20, 40, 80 minutes and 24, 48, 72, 96 hours of exposure at 15+1oC. Cumulative Daphnia mortality was recorded after 24 and 48 hours of exposure at 20+2oC. The 96-hr and 48-hr mortality data, respectively for rainbow trout and Daphnia, were analyzed using the Stephan Program to calculate the LC50 and 95% confidence intervals. A phenol reference test was conducted for the rainbow trout, and a sodium chloride reference test was

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conducted for the Daphnia, to ensure that the organism sensitivity was within acceptable quality control warning chart limits. The Selenastrum growth inhibition test followed protocols outlined in (McLeay, 2000). Tests were conducted using de-ionized water as the control and the diluent for the test concentrations, which were: 10000, 5000, 2500, 1250, 625, 312.5, 156.3, 78.1, 39.1, 19.5 mg/L and a control. Each concentration had five replicates with an initial approximate concentration of 10,000 organisms per replicate in 220 µL of test solution. Algal cell yield was recorded using a Coulter (particle) Counter after 72 hours at 23+1oC. Algal cell yield data was analyzed using ToxStat V3.5 to calculate the IC50, IC25 and 95% confidence intervals. The 72 hour IC50 and IC25 is the concentration of sample estimated to cause a 50% and 25% inhibition in growth of the algae over an exposure period of 72 hours. A copper reference toxicant test was conducted to ensure that the organism sensitivity was within acceptable quality control warning chart limits. Chemical Analyses Nutrient solution chemistry samples from both solutions were prepared at the LC50 or IC50 concentrations following the toxicity tests. The chemistry samples were analyzed for total metals, nitrogen ammonia, nitrogen nitrate and nitrite, and total nitrogen using an ICAP spectrophotometer and standard methods, respectively. Application Dilution The nutrient solution is added to the lake by filling tanks or the hold on a boat with the solution and cruising down the middle of the lake while pumping the nutrients into the propeller wash behind the boat (Figure 2). The propeller wash is an extremely turbulent zone of water that spreads out behind the boat in three dimensions. Application concentrations were calculated from conditions found in lake enrichment projects in British Columbia. The diluted concentration of nutrients when it first enters the water can be calculated from the total amount of nutrient solution added per trip or from the instantaneous addition rate:

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XST

TN AD

NC*

=

XSB

AN AV

FC*

60** ρ=

where: CN is the final concentration of the nutrient solution (in mg/L), NT is the total amount of nutrients added (in kg), DT is the total distance travelled during the application (in km), AXS is the cross sectional area influenced by the propeller wash (in m2), FA is the flow of nutrients into the lake during application (in L/min), ρ is the specific gravity of the nutrient solution (about 1.4, dimensionless), VB is the velocity of the boat (in km/hr), and 60 is a factor that converts hours to minutes. The concentration of the nutrient elements (N and P) would be proportionately less, according to which source nutrient was used. The 28-0-0 contains 28% N as a combination of ammonia, nitrate and urea. The 10-34-0 contains 10% nitrogen and 34% phosphorus as P2O5, which translates to about 14.6% elemental P.

Figure 2. Application vessel for the Adams Lake Enrichment Project, 1997.

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Results and Discussion Toxicity Tests The test results indicate that the N-rich 20-0-0 is about twice as toxic to trout as the P-rich 10-34-0 (Table 1). Table 1. Toxicity levels of fertilizers used in Lake Enrichment Program Chemical Test

Organism Type of Test

Length of Test

Effective Concentration

95% C.I. (mg/L)

28-0-0 Trout. LC50 96 hrs 585.1 mg/L 454-745 10-34-0 Trout. LC50 96 hrs 1341.6 mg/L 1000-1800 10-34-0 Daphnia LC50 48 hrs 1229.4 mg/L 1000-1800 10-34-0 Selenastrum IC50 72 hrs 1745.0 mg/L 1674-1818 Chemical Composition of Test Solutions Heavy metals were all below detection limits (0.03 mg/L or 0.005 mg/L, depending on the element) for all test solutions (including the toxic heavy metals: aluminum, arsenic, cadmium, chromium, cobalt, copper, iron, lead, manganese, tin and zinc). The total nitrogen concentration of the ‘lethal’ solutions is very similar for rainbow trout and Daphnia at approximately 130-150 mg/L (Table 2). This indicates that it is probably nitrogen, in some form, that causes the lethality. The nitrogen content of the 28-0-0 is approximately 1/3 ammonia and 1/3 nitrate. The other major component is urea, which was not analysed from these samples. It is possible that the urea degraded into ammonia, which is known to be toxic (Sigma, 1983), over the duration of the test.

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Table 2. Concentrations of nitrogen compounds in the test fertilizers at the lethal concentrations.

Organism Rainbow

Trout Rainbow Trout

Daphnia Selenastrum

Fertilizer 28-0-0 10-34-0 10-34-0 10-34-0 Test Solution Concentration

585 mg/L 1341 mg/L 1229 mg/L 1745 mg/L

Ammonia 42 mg/L 115 mg/L 112 mg/L 160 mg/L Nitrite <0.002 mg/L <0.003 mg/L <0.003 mg/L <0.003 mg/L Nitrate 42.2 mg/L 8.3 mg/L 11.6 mg/L <0.2 mg/L Total N 130 mg/L 149 mg/L 139 mg/L 209 mg/L In-Lake Dilution At the Adams Lake Enrichment Project, the boat travels approximately 20 km in the application zone, discharging 7000 kg of nutrient solution. The boat travels at 10 km/hr and discharges the nutrient solution at about 42 L/min. I estimate that the propeller wash of the vessel shown in Figure 2, fully loaded with nutrient solution, causes a turbulent zone approximately 4 m wide by 3 m deep. The smaller vessel used on the Great Central Lake Enrichment Project causes a smaller turbulent zone (~3m wide by 2 m deep), but it discharges nutrients at less than half the rate of the Adams Lake project. Inserting the Adams Lake conditions into the formulae shows that the nutrient solution is diluted to about 29 mg/L within the propeller wash. This value is about 5% of the LC50 value for the most toxic of the nutrient solutions on the most sensitive species. However, this is the highest concentration of the nutrient entering the lake water. The nutrient solutions are extremely soluble and continue to disperse and dilute so rapidly that water sampling a day or two later cannot find a gradient in nutrient concentrations in a lake being enriched. Conclusions The inorganic fertilizers that are used to enrich lakes can be toxic to aquatic life in high concentrations. However, such high concentrations are not likely to be encountered during standard techniques of adding nutrients to lakes because of the turbulence of the immediate application zone and the rapid dilution with surrounding lake water.

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method for determining acute lethality of effluents to rainbow trout. Report EPS 1/RM/13 2nd Edition, Environment Canada 23 p.

McLeay, D.J. and J.B. Sprague. 1990. Biological test method: acute lethality test

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Miller, J.A., R.P. Scroggins and D.J. McLeay. 2000. Biological test method:

reference method for determining acute lethality of effluents to Daphnia magna. Report EPS 1/RM/14 2nd Edition, Environment Canada 21 p.

St. Laurent, D., G.L. Stephenson and K.E. Day. 1992. Biological test method:

reference method for determining acute lethality of effluents to rainbow trout. Report EPS 1/RM/25, Environment Canada 41 p.

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