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1. GENERAL INTRODUCTION

1.1 Marine Pollution in Malaysia

The west coast of Peninsular Malaysia plays a major role in the maritime trade of

the country (Naidu, 1993). It is the hub for the major urban centres, industries and

plantation. These population centres are estimated to account for at least 8.5 million (70%)

of Malaysia’s 20 million population. The rivers, which run through most of the

urban/industrial areas, are the main repository of domestic and industrial wastes, sewage,

organic and inorganic loadings. Siltation from land-based sources, oil and grease from

shipping activities and other contaminants that result from man’s economic activities also

contribute to the general pollution of the aquatic environment (Department of

Environment, 1997). Choo et al. (1994) published an overall assessment of the state of the

coastal marine environment for the west coast of Peninsular Malaysia. The marine waters

of the west coast of Peninsular Malaysia are considered to be most prone to land-based

pollutants (Phang et al., 1991) since 75% of the population and 85% of industries are

concentrated there (Maheswaran and Godwin, 1988). Jaafar (1991) discussed the

management issues related to the marine environment of the Strait of Malacca.

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The coastal marine environment of the west coast Peninsular Malaysia, which also

forms the eastern portion of the Strait of Malacca, has been the main repository of both

land-based pollution and pollutants derived from shipping activities (Choo et al., 1994;

MIMA, 1994). This portion of the Strait of Malacca, within Malaysian territorial waters,

has also been traditionally the major fishing ground for Peninsular Malaysia, and as much

as 70% of capture fisheries come from these waters (Department of Fisheries Malaysia,

1993). Landings of marine fish showed an increasing trend between 1970 and 1980, but

then declined until 1986. Although from 1986 onwards there was an increase in landings,

this increase was attributed to the introduction of deep-sea or offshore fishing activity. Lui

(1992) hypothesised that the fisheries resource within the inshore waters of the Strait of

Malacca had reached its maximum level of exploitation. There was a steep decline in fish

catch per unit population from 1970 to 1990, which indicated that the fish resource was

being exploited beyond its maximum sustainable yield (Sheppard, 1992). Overexploitation

has been suggested as one of the main reasons for the decline in fish resource. However,

the role of mainly land-based pollution and destruction of natural habitats have been

suggested as other major factors responsible for the current decline in inshore fish

resources (Sasekumar, 1980; Phang, 1990). It was also suggested that oil spills from

tankers and bilge from shipping operations contribute to pollution originating from the sea

itself (Absil et al., 1987; Maheswaran and Godwin, 1988).

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The rivers that flow into the marine coastal waters inadvertently affect critical

marine habitats such as mangroves, estuarine mudflats, seagrasses and coral reefs (Peters

et al., 1997). Mangrove swamps on the west coast cover 103,000 ha and comprise more

than 90% of the total mangrove coverage for Peninsular Malaysia (Tang et al., 1980). The

mangrove habitat is a nursery ground for many marine organisms including several

commercial species of fish and shrimps (Robertson and Duke, 1987). Mangrove detritus

forms a sizeable portion of the food of several fish species (Thong and Sasekumar, 1984).

Sources of Pollutants

The Strait of Malacca is the second busiest sea-lane in the world. In 1994, it was

estimated that about 3000 vessels/day plied the maritime waters of the Strait of Malacca.

This number did not include thousands of fisheries vessels navigating the waters. Rapid

industrialisation coupled with chemical-dependent modern agricultural activities have led

to increasing loading of xenobiotics and other contaminants to the coastal waters.

Treatment and mitigating technologies currently in place have not been able to cope

adequately with the rapid pace of industrialisation.

Studies by Law et al. (1993) and Abdullah (1995) showed that the marine waters

of the north-eastern part of the Strait of Malacca around the Langkawi Group of Islands

were contaminated with significant levels of toxic dissolved/dispersed hydrocarbons.

Abdullah reported levels of hydrocarbon concentrations of 1.73-1.97 mg L-1 around

Langkawi Island. Abdullah et al. (1996), detected in samples of coastal sediments low

concentrations of polycyclic aromatic hydrocarbons (PAHs), breakdown by-products of

petroleum.

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In 1997, 117 major rivers were monitored in Malaysia, generating 4078 samples

from 908 stations (DOE, 1997). Twenty-five rivers (11 in 1993, 13 in 1996) were

considered badly polluted in 1997, based on water quality index (WQI) criterion (The

WQI classification is based on biochemical oxygen demand (BOD), chemical oxygen

demand (COD), dissolved oxygen (DO), ammoniacal nitrogen, suspended solids and pH

levels). This is an increase from 12 rivers the previous year. The marine environment also

deteriorated over the same period. The main pollutants were siltation from earthworks,

sewage and oil & grease. In 1997, a total of 4,075 industries were identified as significant

water pollution sources for the west coast of Peninsular Malaysia. The main industries

were Food & Bevcrages, paper treatment, electrical & electronic and metal finishing.

Other significant sources of pollution were sewage treatment plants (4,359) and animal

husbandry (2,029). The main source of organic pollution load was sewage (776.4 metric

tonnes BOD/day), wastes from pig-farms (304.5 metric tonnes BOD/day), the

manufacturing sector (5.9 metric tonnes BOD/day) and the agro-based industries (8.4

metric tonnes BOD/day). The largest contribution of BOD loading from the

manufacturing sector came from the Food & Beverages industry amounting to 1,166.4

kg/day (20%).

Coastal marine water also showed considerable deterioration in 1997. Eighty-four

of the 226 monitoring stations were found to have exceeded the Proposed Marine Interim

Standards for of oil & grease (0 mg L-1), suspended solids (50 mg L-1) and Escherichia coli

(100 MPN /100 mL-1). Copper levels exceeding the proposed standard of 0.1 mg L-1 were

recorded in Sarawak. Mercury and arsenic exceeded the standards (0.001mg L-1 for Hg

and 0.1 mg L-1 for arsenic) in coastal waters off Negeri Sembilan.

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Pesticides in the aquatic environment

Pesticide use became widespread after WWII when DDT and other synthetic

chemicals were introduced to offer an easy solution to the age-old problem of pest control

in agriculture. Amongst the many xenobiotics that enter the aquatic ecosystem, pesticides

pose considerable threat to the biota. In 1997, the Pesticide Board of the Ministry of

Agriculture Malaysia registered 1,758 types of pesticides of which 535 (30.5%) are

classified as very hazardous and 574 (32.7%) as moderately hazardous. It was estimated

that 12 million kg of pesticides used in agriculture were in solid form while 8 million litres

were liquid (MACA, 1997). Ninety percent pesticides used in Malaysia were for the

rubber, oil palm and rice plantations, of which herbicides account for 75%, insecticides

13% and 3% were rodenticides (MACA, 1996). In 1985, the total value of pesticides sold

in Malaysia was US$140 million. Of the organophosphate and carbamate insecticides, the

more important were methamidophos, malathion, chlorpyrifos, fenvalerate, carbofuran and

carbaryl. Diazinon and isoprocarb are used mainly for vegetables. The usage of pesticides

in Malaysia is being regulated and monitored by the Pesticides Board of the Ministry of

Agriculture. The Pesticides Act 1974 regulates the import, manufacture, distribution, sale

and use of pesticides. The Act requires all pesticides to be registered in accordance with

standards and criteria of FAO/WHO. The safety levels of pesticide residues in foods are

determined by the Food Act 1983. The enforcement of this Act is under the purview of

the Ministry of Health.

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Pesticides enter the aquatic environment through their application in the field. The

distribution, fate and mobility of pesticides when applied in the field and the factors

affecting their entry into the aquatic environment have been reviewed (Nicholls, 1988,

Jaffe, 1991). Studies on pesticide contamination in Malaysia have focussed on the

persistent organochlorine pesticides (OCPs). The occurrence of OCPs had been detected

in Malaysia’s aquatic environment. Tan et al.(1990) reported contamination of Kelang

River (one of the most polluted rivers in 1997) by OCPs such as aldrin (0.005-0.061 ng L-

1), endosulfan (0.009-0.256 ng L-1), and heptachlor (0.039-0.742 ng L-1). The same study

also reported the occurrence of small quantities of DDE, DDT and heptachlor in all major

rivers surveyed, at levels below the critical values of the Malaysian Interim Standards for

Water Quality (1992). Cavalho (1993) suggested that land-based pesticide contamination

might still pose a considerable threat to marine biota as the major agricultural areas are

located in coastal plains and river valleys. Contamination by OCPs in the marine biota,

including fish and shellfish, has been documented (Jothy et al., 1983; Rohani et al., 1992).

The study by Jothy et al. (1983) showed that OCP residue levels were found to be low in

all samples analysed except for the blood cockle, Anadara granosa, from Penang and

Perak where the level for total DDT ranged from 0.027-0.05 mg L-1. The OCPs detected in

fish were lindane (0.001-0.012 mg L-1), dieldrin (< 0.001-0.004 mg L-1) and total DDT

(0.003-0.016 mg L-1). Rohani et al. showed that OCP residues were generally low except

for lindane which were detected at 180.9 and 123.7 g L-1 in the mussel, Perna viridis

collected from two sites in Penang. Rohani et al. (1992) reported lower levels of OCPs

compared with similar species analysed by Jothy et al. (1983). It was suggested that the

lower values detected could be explained by the phasing out of OCPs in Malaysian

agriculture.

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The phasing out of OCPs coincided with the introduction of the more toxic but less

persistent organophosphorus and carbamate pesticides into Malaysia. However, it was

suggested that the more toxic OPs and carbamates could be sufficiently persistent to exert

some effect in marine biota (Cavalho, 1993). Efforts to assess impacts of OPs and

carbamates on the freshwater aquatic ecosystem were initiated by measuring their

biological effects – especially acetylcholinesterase activity in freshwater fish (Sulaiman et

al., 1989; Abdullah et al., 1993, 1994).

Heavy Metals in aquatic environment

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Heavy metals are regarded as serious pollutants in marine ecosystems because of

their environmental persistence, their toxicity at low concentrations and their ability to be

incorporated into food chains and concentrated by organisms (Negilski, 1976). Metals

such as zinc and copper are essential trace metals required for biological processes. Other

essential trace metals are Fe, Co, Mn, Cr, Mo, V, Se, Ni and Sn (Bryan, 1980). Metals may

become toxic when accumulated from seawater and concentrated in the tissues of aquatic

organisms. Cadmium is not known to be an essential metal for any organism. In Malaysia,

monitoring of trace metals had concentrated on their levels in the tissues of fish, shellfish,

and in water and sediments (Jothy et al., 1983; Liong, 1986; Devi, 1986; Din, 1992; Din

and Jamaliah, 1994; Din et al., 1996). Rakmi and Salmijah (1987) estimated that 220 000

m3 of toxic waste was produced from the manufacturing industry which contained lead

(5.9 mg L-1), copper (224.0 mg L-1), Cr6+ (6.6 mg L-1), Cr3+ (94 mg L-1), Ni (159.0 mg L-1),

Zn (567.0 mg L-1) and Fe (2661.0 mg L-1). Results from a study carried out in 1990

involving 152 sampling stations showed that concentrations of heavy metals in coastal

seawater of Cd, Cu, Cr, Pb, Hg and As surpassed the Malaysian interim marine water

quality standards (DOE, 1991). Trace metal contents in estuarine and marine sediments

along the coast of Penang Island showed elevated levels of As, Cd, Cr, Pb and Zn

compared to control sites (Din and Jamaliah, 1994). Sungai Prai and Gertak Sanggul were

shown to have the highest anthropogenic input of metals. Results from a similar study

carried out in 1997 showed an improvement in heavy metals detected in the coastal marine

environment, with only Hg and As exceeding the interim standard of 0.0001 g L-1 and

0.1 g L-1 respectively (DOE, 1997). Liong (1986) showed that the concentrations of Cd,

Cu, Pb and Zn in various shellfish species were 0.19-1.32, 0.98-34.84, 0.06-0.42 and

19.28-471.8 g g-1 wet weight, respectively.

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Information on the toxicity of various metals on tropical marine organisms is still

under development. Some studies had been carried out on the Malaysian marine

organisms (Din, 1988; Zain 1990, Hamzah, 1991; Mohd. Noor, 1994; Ong and Din, 1995;

Shazili, 1995; Noorzah and Eliza, 1996; Almah and Mazlin, 1996; Phang et al., 1996).

Toxicity data from tropical marine organisms are required for environmental criteria

formulation. The ASEAN-Canada Marine Science Programme Phase II (1992-1998) is a

concerted effort by ASEAN member countries to develop criteria and standards for marine

organisms and selected parameters including trace metals.

Threats from Pollution

The number of health-related cases associated with the consumption of

contaminated seafood, have been on the increase over recent years. A more recent event

has been the first ever report of a Paralytic Shellfish poisoning occurrence in Peninsular

Malaysia linked to consumption of contaminated mussels (Ismail et al., 1994). In April

1995, there was an incident of cyanide poisoning from contaminated fish reported on the

Island of Pangkor (Department of Fisheries, 1995). The source of cyanide was from

leaking containers, which were illegally dumped into a garbage disposal area fringing the

coastline. The number of case of fish mortality have also been on the increase both for

wild and cultured species (Choo et al., 1992).

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Decreasing fish production over recent years has been attributed to

overexploitation from the use of increasingly efficient fishing-gear technology. But

anthropogenic factors such as xenobiotic contaminants that impact marine ecosystems

cannot be ruled out as contributing to this state of health of the fisheries. Studies by Law et

al. (1993) have shown that juvenile shrimps are susceptible to very low concentrations of

phenolics. However, more research needs to be carried out to ascertain the impacts of

xenobiotics on the different life-stages of marine organisms. The water-quality monitoring

programme of the Department of Environment has been providing information on the state

of the environment. However, the physico-chemical data at best provide only information

on contaminants in the water body, and little information on the effects of these

contaminants on the marine biota. Biomonitoring for residues is being carried out at a

minimal level. This is perhaps due to insufficient expertise and/or funding currently

available.

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The Strait of Malacca, and the marine flora and fauna associated with it, are facing

a major threat from many chemical and non-chemical pollutants. Although some

initiatives have been taken to monitor for some of these pollutants, very little information

is available on the biological effects of exposure to xenobiotic chemicals. Development of

toxicity tests for tropical species is being pursued in the ASEAN region under the action

plan for conservation of nature. The 7-year ASEAN-Canada Cooperative Programme on

Marine Science (Phase II) focuses on establishing criteria for management of living

marine resources (Watson et al., 1992). More research needs to be carried out to ascertain

whether these xenobiotics, singly or in a mixture, can elicit detrimental physiological

changes and/or mortality in marine organisms. To this end, the recently developed

concept of utilising biomarkers for environmental risk assessment is being critically

evaluated on a worldwide basis.

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1.2. Relevance of Biomarkers to Environmental Risk Assessment

Since the days of the Minamoto tragedy in 1956 (Clark, 1986), environmental risk

assessments have become increasingly important in many countries of the world.

Governments have set up elaborate water quality monitoring programmes and regulatory

frameworks to protect the aquatic environment, and, in particular, human health.

Conventional risk assessment involves monitoring for certain classes of

xenobiotics and other physico-chemical indicators of pollution. These are normally very

costly, involving sophisticated analytical instruments and elaborate field sampling

strategies (NRC, 1992). Decisions on the permitted release of xenobiotics into the

environment are dependent on a process of risk assessment. Assessment of risks has

always been based upon a comparison between laboratory toxicity data of surrogate

species and the expected exposure in the environment.

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Furthermore, in a dynamic system such as a moving water body, contaminants are

present in high concentrations for very short periods of time before dilution takes effect.

This poses a real problem in acquiring a representative sample in the time window

available. Other complicating factors when attempting accurate and representative

sampling are the sediment type, bottom topography, water currents and water chemistry

(see McCarthy and Shugart, 1990). Toxicological data derived from laboratory

experiments have provided clues to the relationship between contaminants and their

possible negative effects on aquatic flora and fauna. These organisms are continuously

exposed to their surrounding matrix and therefore should integrate the contaminants

present, even if the xenobiotics are short-lived in the environment, e.g. organophosphate

insecticides and polyaromatic hydrocarbons. Exposure in itself is not reflective of the

bioavailability of contaminants. When this problem is combined with the wide diversity of

potential routes, species-specific differences and the pharmacodynamics of the

contaminants themselves, an assessment becomes much more difficult (McCarthy and

Shugart, 1990). For example, many toxicants do not bioaccumulate but are metabolised in

many cases into more toxic metabolites. One example of such a compound is the

organophosphate insecticide, malathion which upon accumulation is converted to the more

toxic metabolite, malaoxon (Murphy et al., 1968). Because of these problems the main

concern for environmentalists and regulators in recent times has been the level of

uncertainty in determining the ecological effects that result from exposure to toxic

chemicals released into the environment.

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Physico-chemical analyses, although valuable and necessary, do not provide all the

pertinent information required for pollution assessment (Gray et al., 1992). Biological

studies and ecotoxicological approaches can provide a more meaningful and realistic

assessment of the state of the marine environment. This is so because biological studies

provide direct information on biological effects.

Over the last decade, there has been an increasing emphasis on the use of

biochemical, physiological, and histological changes as well as other aberrations in

organisms to estimate either exposure to chemicals or the resultant effects (Huggert et al.,

1992). These changes in biological response have been termed ‘biomarkers’.

Biomarkers in aquatic environmental monitoring

Biomarkers are defined as “biological responses that can be related to an exposure

to, or toxic effect of, an environmental chemical or chemicals”. Biomarkers can be

considered as intermediates between sources of contamination and higher-level effects

(Suter, 1992). It is therefore necessary to show the presence of a dose-response

relationship, either a predictable increase or decrease of the biomarker with increasing

exposure, and that higher-level effects are predictable from such. As such, biomarkers can

act as early warning of environmental effects of xenobiotics before there are serious effects

upon individuals or population.

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Depledge (1993) suggested that biochemical and cellular biomarker responses on

their own are of unknown ecological significance. Increased MFO induction in livers of

fish from highly polluted water may signify pollution exposure, but the fish may continue

to grow and reproduce normally and the MFO response may be viewed as an

acclimatisation process to altered environmental conditions rather than a manifestation of

an injury (Jimenez et al., 1990). Therefore, the use of a hierarchy of biomarkers has been

suggested to be the most appropriate method to assess environmental risks from xenobiotic

insult (Depledge, 1993).

Fish in Biomarker Research

Fish are suitable organisms with which to monitor aquatic contamination, as they

are located at the top of the aquatic food chain. Fish are known to accumulate toxicants

and they are in direct contact with polluted water via their gills and body surface. As such

fish have been used in toxicological investigations of the aquatic environment.

Biomarkers have been applied to the field in assessing fish health populations

experiencing contaminant stress (Shugart and Southworth, 1990; Lower and Kendall,

1990) and in ecological risk assessment (Suter II, 1990). When working with biomarkers

the genetic differences between individuals taken from different populations including the

occurrence of genetic resistance to xenobiotics as shown in the mussel, Mytilus edulis,

should always be considered (Minier and Galgani, 1995). There is also the need to

consider, and therefore, plan out accordingly, differences that might ensue from species,

size, age, sex, season, habitat, and geographical origin (McCarthy and Shuggart, 1990).

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Biomarkers can also be categorised according to their specificity to an exposure.

One category which can be divided into two groups comprising the stressor independent

proteins; the heat shock proteins which are induced by a range of compounds and metals

(Sanders, 1990); and glucose-regulated proteins, which are induced when there is

deprivation of oxygen or glucose. The other group is the stressor-dependent biomarkers:

the mixed-function oxygenases (MFOs), Glutathione S-transferases (GSTs), the haem-

oxygenases (HO) and the metallothioneins (MT), where proteins are induced by exposure

to specific contaminants or physical conditions. The second category of biomarkers is

based on their specific responses to a class of contaminants. These are the biochemical,

cellular and whole animal. Among the biochemical biomarkers are the occurrence of DNA

adducts (Ericson et al., 1995), micronuclei in blood cells (Spies, 1990), and liver

oncogenes (McMahon, 1990). Cellular biomarkers would include induction of MFOs by

organophosphates (OPs), polynuclear aromatic hydrocarbons (PAHs), polychlorinated

biphenyls (PCBs) (Forlin and Anderrson, 1995); acetylcholinesterase inhibition

(Bocquene, 1993), metallothionein induction (Kille, 1995), and vitellogenesis induction by

oestrogenic-like xenobiotics (Matthiessen, 1995). Whole-animal biomarkers, which

involve measurement of physiological end points such as scope for growth (sfg) and

adaptive behaviour, are strictly not biomarkers if the NRC definition is applied.

1.3. Problem Statement

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Conventional water quality monitoring programmes have been expensive and

manpower intensive. The resulting data were of questionable quality (National Research

Council, 1992). A great deal of work on environmental monitoring has been focused on

the determination of mainly physico-chemical parameters and residue levels, especially the

organochlorines and heavy metals (Peakall and Shugart, 1991). There is currently a move

in the United States of America and Canada to undertake an integrated approach to

environmental monitoring viz-a-viz water quality monitoring, benthos community studies

and toxicity studies - the TRIAD study (Chapman et al., 1995). Though the approach is

deemed to produce better quality information on the state of the environment, it will no

doubt be much more expensive to implement.

Malaysia has instituted, for the past two decades, a monitoring programme for

water quality in both riverine and marine environments. As with the North American

experience, environmental monitoring has been expensive while the data generated have

been useful only to a limited degree for management decisions. Very limited information

is available for residue levels in marine organisms, and less still for toxicological effects on

these. It is not surprising that marine water-quality-criteria development in the tropics is

still being developed. The ASEAN-Canada Cooperative Programme on Marine Science -

Phase II (ending 1996) has been assigned the task to generate toxicity data on marine

organisms for criteria development in ASEAN (ACMSP-II, 1991-1998).

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Molecular biomarkers may provide a relatively inexpensive and precise alternative

or/and complementary technique to environmental monitoring of exposure and effects.

Research efforts into the benefits and viability of molecular biomarkers as a monitoring

and diagnostic tool should be immediately initiated for the tropical marine environment so

as to address immediately the increasing problems posed by Man’s economic activities. It

is in this context that the present study is proposed: the initiation of a research programme

to evaluate the responses of acetylcholinesterase, the biomarker of neurotoxic effect, in the

tropical seabass, Lates calcarifer (Bloch) when exposed to various man-made

contaminants.

1.4. Approach for this study

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In this study, acetylcholinesterase response, a proven biomarker of effect will be

studied in the tropical seabass, Lates calcarifer (Bloch). Among the major contaminants in

riverine and coastal marine waters, the organophosphate and carbamate insecticides can

cause serious ecotoxicological problems because of their very toxic nature. Although

transfer of these contaminants to the aquatic environment is generally episodic and usually

diffuse and chronic, it has been shown that very low concentrations of these contaminants

could compromise the well being of aquatic animals (Galgani and Bocquene, 1996).

Furthermore, recent evidence has indicated that AChE inhibition is not only caused by the

organophosphates and carbamates alone – the presence of other, yet unidentified,

contaminants may be causing the inhibitory response (Payne et al., 1995; Huang et al.,

1996). Payne et al., (1996) found muscle AChE in the trout, Salmo trutta collected from

two urban rivers receiving kraft pulp mill effluents, was significantly. Similar AChE

inhibition was also observed in the spotted garfish, Lepisosteus oculatus attributed to

exposure to multiple contaminants in the lower Mississippi River basin (Huang et al.,

1997). Therefore, AChE must not be considered as a specific biomarker just for

organophosphorus and carbamate insecticide exposure (Galgani and Bocquene, 1989; Day

and Scott, 1990).

There are three approaches to this study involving:

1. In vitro exposures of brain AChE from unexposed fish to various

contaminants, and assaying for inhibitory effects,

2. In vivo exposures of seabass to various xenobiotics under controlled laboratory

conditions, and

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3. Field-exposures of feral fish at various locations as to validate the laboratory-

based findings.

Seabass will be exposed to sub-lethal levels of the organophosphates, diazinon and

fenitrothion; the carbamate, methyl isopropyl carbamate (isoprocarb); the organochlorine,

lindane; the polyaromatic hydrocarbon, benzo()pyrene, used crankshaft oil (both

acetone- and water-soluble fractions), and the trace metals, cadmium, zinc and copper.

The choice of diazinon, fenitrothion and isoprocarb was based on their ease of availability

and widespread use in Malaysian agriculture and horticulture. Benzo()pyrene is normally

used as an indicator of PAH contamination which might ensue from petroleum and

shipping activities; while lindane, a member of the OCP family of insecticide is still being

legally used in Malaysia. Used crankshaft oil and industrial effluents have been identified

as significant pollutants. The addition of the trace metals is to validate their influence on

AChE activity in the presence of xenobiotics.

Recent developments in AChE determinations have moved from the

spectrophotometric (BMC) procedure based on Ellman’s microplate-based technique

(Moores, 1988). It was shown that the microplate technique is much more sensitive than

the spectrophotometric method, with a coefficient of variation of 6.9 - 25.5% in the former

and 34.6 - 55.4% in the latter (Day and Scott, 1988). The assay technique will initially

follow that of Galgani and Bocquene (1991) with the protocol being optimised for the

tropical seabass, as the technique is being developed.

Ultimately, it is hoped that the seabass will be utilised as a sentinel for the

monitoring of environmental contamination in riverine and coastal marine waters.

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1.5. Specific Objectives

The objectives of this study are:

1. To evaluate and optimise the methodology for AChE activity for the tropical

seabass, L. calcarifer.

2. To ascertain the 96-hr LC50 values of different xenobiotics and trace metals in

relation to AChE activity in the tropical seabass, L. calcarifer.

3. To determine in vivo dose-response relationships between xenobiotics and the

activity of AChE in the tropical seabass, L. calcarifer of different size,

nutritional conditions and exposure to different salinity regimes.

4. To undertake in vitro studies on the effect of xenobiotics and trace metals,

singly and in combination, on the AChE activity of the seabass.

5. To evaluate the response of AChE in caged feral seabass, L. calcarifer,

exposed to waters from various coastal and estuarine locations in Penang.

1.6. References

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_______________. (1989). Heavy metals in Malaysian shellfish. Paper presented at

Fisheries Research Seminar, Department of Fisheries Malaysia, June 25-29,

1989, Melaka.

Annual Fisheries Statistics, 1970-1990. Department of Fisheries, Ministry of

Agriculture, Kuala Lumpur, Malaysia.

Choo, P.S., I. Ismail and H. Rosly. (1994). The West Coast of Peninsular

Malaysia. In: An Environmental Assessment of the Bay of Bengal Region.

Staffan Holmgren (Edt). Bay of Bengal Programme. Madras, India. pp 33-51.

Department of Environment. (1990). Environmental Quality Report. Annual Report :

Ministry of Science Technology and the Environment Malaysia.

Department of Fisheries Malaysia. (1989). Deep-sea Fisheries Resources within the

Malaysian Exclusive Economic Zone – Survey Report of Demersal and

Pelagic Resources. Department of Fisheries, Ministry of Agriculture, Kuala

Lumpur, Malaysia.

Galgani, F. and G. Bocquene. (1991). Semi-automated colorimetric and enzymatic

assays for aquatic organisms using microplate readers. Water Res. 25 (2):

147-150.

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Jimenez, B.D., A. Oikari, S.M. Adams, D.E. Hinton, and J.F. MaCarthy.

(1990). Hepatic Enzymes as Biomarkers: Interpreting the Effects of

Environemntal, Physiological and Toxicological Variables. In: Biomarkers of

Environmental Contamination (Editors J.F. MacCarthy and L.R. Shugart).

Lewis Publishers

Jothy, A. A. (1973). Coral Reefs in the Coastal Waters of West Malaysia, their

Utilisation and Conservation. Department of Fisheries, Ministry of Agriculture

Malaysia.

Lui, Y.P. (1992). Multispecies Fish Resources and Multispecies Fisheries in the

Coastal Waters of Peninsular Malaysia, with special reference to the West

Coast. Paper presented at the 6th Session of the Standing Committee on

Research and Development, Colombo, Sri Lanka May 18-21 1990, FAO

Fisheries Report No. 463 Suppl. Rome, Italy, 215 pp.

Phang, S. M. (1990). Seagrass - a neglected natural resource in Malaysia. Proc.

12th Ann. Sem. Malay. Soc. Mar. Sci., Kuala Lumpur.

Rohani, I., S.M. Chan and I. Ismail. (1992). Organochlorine pesticides and PCB

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