Bioaccumulation of Metals in Selected Fish Species and the Effect of Ph on Aluminium Toxicity in a Cichlid Oreochromis Mossambicus

BIOACCUMULATION OF METALS IN SELECTED FISH SPECIES AND THE EFFECT OF PH ON ALUMINIUM TOXICITY IN A CICHLID OREOCHROMIS MOSSAMBICUS

BY LIZET COETZEE

THESIS

SUBMITTED IN FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE MASTER OF SCIENCES

IN ZOOLOGY

IN THE

FACULTY OF NATURAL SCIENCES

AT THE RAND AFRIKAANS UNIVERSITY

SUPERVISOR: PROF H.H. DU PREEZ CO-SUPERVISOR: PROF J.H.J. VAN VUREN

MAY 1996 - ACKNOWLEDGEMENTS -

Special thanks to:

My Lord for the wisdom, strength and insight He gave me:

" By wisdom the Lord laid the earth's foundation; By understanding He set the heavens in place" "Wisdom is supreme, therefore get wisdom, though it cost all you have; get understanding" Proverbs 3:19; 4:7

My supervisor, Prof H.H. du Preez and co-supervisor, Prof J.H.J. van Vuren, for their support, motivation and guidance throughout the project

Prof J.H. Swanepoel, the head of Department Zoology, for the use of the facilities and the opportunity to perform this study

My parents for their patience, love and financial support during my studies

My husband for his love, support and motivation; without him it wouldn't have been possible

The Rand Afrikaans University, Water Research Board and the Foundation for Research Development for their financial support

The Institute for Water Quality Studies, Department of Water Affairs and Forestry for the chemical analysis of the water samples

Dirk Erlank Gabriel Motlhabane, Solomon Kwapa (Switch) and Solomon Tshabala (Sony) for their assistance with the experimental systems and my work in the aquarium

Naas and Reinette Ferreira for the use of their computer and all their support

Irma Naude for linguistically attending to my thesis

Gail Nussey for all her support and advise -ABSTRACT-

The Upper catchment of the Olifants River, from its origin near Bethal, to its confluence with the Wilge River, north of Witbank, as well as it tributaries, are being subjected to increasing afforestation, mining, power generation, irrigation, domestic and industrial activities. These activities have a profound effect on the water quality and the major point sources of pollution in this area include mines, industries and very importantly, combined sewage purification works, located alongside the river, which, in addition to oxidizable material contains detergents, nutrients, and metals. It was therefore necessary to determine the extent to which these activities affect the water quality of the system. The impact of these activities was therefore addressed by a Water Research Commision Project namely "Lethal and sublethal effects of metals on the physiology of fish" of which the present study investigated effects at two localities, namely in the Olifants River (locality OR1) before its confluence with the Klein Olifants River and a locality in the Klein Olifants River (locality KOR1). Apart from the field study, toxicity tests were also performed in a laboratory, in order to determine the effects of low pH and elevated aluminium concentrations on the haematology, osmoregulation and carbohydrate metabolism of the Mozambique Tilapia, Oreochromis mossambicus as the acidification of soil systems may cause the transfer of aluminium into aqueous solutions, where it may be present in different forms. During the field study, the chemical and physical characteristics of the river water were evaluated, with special attention to the concentrations of certain metals (manganese, copper, chromium, lead, nickel, zinc, iron and aluminium) in the water and sediment, as well as in fish, which are known to accumulate the elements supra and are therefore valuable as indicators of these pollutants.

ii The two fish species used for the investigations were the African sharptooth catfish, Clarias gariepinus and the moggel, Labeo umbratus. Four tissue types were dissected, namely the muscle, liver, skin and gill tissues. The metal concentrations in these organs/tissues, as well as in the water and sediment, were determined in a laboratory with an atomic absorption spectrophotometer. Statistical analyses were performed on the results obtained from this study and the order and extent of bioaccumulation of these metals in the water and sediment were determined, as well as in the fish organs/tissues. Its dependence on the size, sex and species of the fish and the localities and seasons were investigated.

The selected water quality variables were mostly well within the guideline limits proposed for the protection of aquatic life, except for the phosphate concentrations, which were much higher than the permitted level of 1 mg/I for effluent water. The high phosphate concentrations in the water were due to effluent received from combined sewage purification works in the area of localities KOR1 and OR1, informal settlements situated alongside the river and agricultural runoff. With the exception of manganese, copper and zinc, the metal concentrations were higher than the recommended guideline limits. All the selected metals were, however, present in much higher concentrations in the sediments, especially iron and aluminium, with the highest concentrations found in the smallest particle sizes of the sediments. The high metal concentrations in the water and sediment indicated some degree of metal pollution, but due to the hardness of the water, these metals were not necessarily acutely toxic to the aquatic organisms. The general order of bioaccumulation of the selected metals was, in decreasing concentrations: Fe, Cr, Al, Ni, Pb, Zn, Mn and Cu in the water and: Al, Fe, Mn, Cu, Cr, Ni, Zn and Pb for the sediment, with exceptionally high concentrations of aluminium and iron. The different organs/tissues of the two species bioaccumulated different levels of the metals and it was clear which organs/tissues accumulated higher concentrations of a certain metal. The gills and liver tissues bioaccumulated the highest metal concentrations, with

iii the highest concentrations of copper and iron found in the liver.

These tissues should therefore be used for determining of the bioaccumulation of these metals in the fish. Although the lowest metal concentrations were found in the muscle and skin tissues, these tissues should always be included in general biomonitoring programmes, as it is consumed by humans and especially in this case, by the people from informal settlements located alongside the river, as well as anglers at locality OR1. High metal concentrations were found in the fish tissue, but concentrations in the muscle and skin tissue were still fit for human consumption. These concentrations should, however, always be monitored, as locality KOR1 receives effluent from combined sewage purification works located upstream from it, as well as raw sewage from informal settlements, different industries and urban and agricultural runoff. In addition to these sources, locality OR1 also receives effluent from mainly coal mines, situated in the catchment area. The results showed dependence of the bioaccumulation of most of the metals in the tissues/organs of both species, on the species and lengths of the fish, as well as the localities where the fish were collected. It is widely accepted that the water characteristics may influence the chemical form, availability and toxicity of various metals to aquatic organisms. The establishment of a general biomonitoring programme is therefore needed, where the hydrological and geomorphological characteristics, the chemical and physical water quality and the riparian vegetation are also considered, because these all affect the aquatic ecosystem.

Identification of the lethal as well as sublethal effects of chemicals on living organisms is critical in evaluating and predicting the impacts of metal pollution. As previously stated, sublethal exposure of fish was performed to investigate the effects of aluminium and low pH, as these have caused fish kills in the catchment. Aluminium has different effects at different pH values and maximum toxicity occurs over the pH range 5.0 - 5.5. The fish were therefore exposed to pH 5.2, as well as combinations of pH 5.2 and

iv different concentrations of aluminium. These concentrations were chosen with respect to concentrations found in the Olifants River and its tributaries during the present study. The results indicated sublethal effects (especially on the carbohydrate metabolism) of the fish at these concentrations and pH value. The pH values at the two localities used for description in this study were much higher, thus no serious aluminium toxicity to aquatic organisms is proposed. There are, however, problems with low pH at other localities of the upper catchment. Decreases in the pH values at localities KOR1 and OR1 due to acid mine drainage or low flow of the river, could cause serious problems, as the aluminium concentrations at these two localities in the sediments are very high and can lead to fish kills.

Nutritionally and recreationally, fish constitute the most important segment of the aquatic ecosystem and therefore metal concentrations in fish are obviously of great concern. High concentrations of metals are toxic to the ecosystem as a hole and to humans in particular, since they are at the end of a variety of food chains by virtue of a varied diet. Further research is therefore needed as high concentrations of metals were found in the tissues and organs of the fish, as well as the sediment and water. It is suggested that the sources of pollution in the study area should be further investigated to fully understand the impact they have on the deteriorating environment. A detailed water quality biomonitoring programme of the study area should also be established by the Department of Water Affairs and Forestry and conducted as frequently as possible. The laboratory investigations on the sublethal and lethal effects of metals on indigenous fish species should be continued and expanded as this would enable scientists to update the Water quality index, Water 2, currently being assessed as a management tool to protect aquatic freshwater fish species.

v -UITTREKSEL-

Die bolope van die Olifantsrivier, vanaf die oorsprong naby Bethal tot die samevloei met die Wilgerivier noord van Witbank, asook die sytakke, word beinvloed deur bedrywighede soos boomaanplantings, mynbou, kragopwekking, besproeiing, verstedeliking en nywerheidsontwikkeling. Hierdie aktiwiteite het beslis 'n effek op die waterkwaliteit van die rivier. Die belangrikste puntbronne van besoedeling in die studie gebied sluit myne en industries in, maar die belangrikste is naasliggende gekombineerde riool suiweringswerke. Die behandelde rioolwater bevat oksideringstowwe, wasmiddels, nutriente en metale. Die mate waartoe hierdie aktiwiteite die waterkwaliteit van die rivier beinvloed, moes daarom vasgestel word. Hierdie doel is bereik deur die chemiese en fisiese eienskappe van die rivierwater te bepaal en te evalueer teenoor bestaande waterkwaliteitsriglyne. Spesiale aandag is aan die konsentrasies van geselekteerde metale in die water en sediment, asook in sekere vis spesies gegee, wat daarvoor bekend is om hierdie elemente te akkumuleer.

Die bioakkumulering van mangaan, koper, lood, chroom, yster, sink, nikkel en aluminium in die water en sediment is ook ondersoek. Vis is by die twee lokaliteite gedissekteer waar monitering plaasgevind het. Die lokaliteite is voor die samevloei van die Olifants- rivier (lokaliteit OR1) en die Klein Olifantsrivier (lokalitiet KOR1) gegee. Die skerptand baber, Clarias gariepinus en die moggel Labeo umbratus is vir analise versamel. Vier weefsel tipes naamlik spier-, lewer-, vel- en kieuweefsel is uitgedissekteer. Die konsentrasies van mangaan, koper, lood, chroom, nikkel, aluminium, yster en sink in hierdie organe en weefsels, sowel as in die water en sediment, is in die laboratorium met behulp van 'n Atoomabsorpsiespektrofotometer bepaal. Die resultate verkry is statisties ge-analiseer om die graad van bioakkumulering van hierdie metale in die water en sediment vas te stel. Die vlakke van bioakkumulering in die visweefsel en organe en

vi die afhanldikheid daarvan van die grootte, geslag en spesie van die vis, asook die lokaliteite en seisoene waartydens die vis versamel is, is ook bestudeer.

Daar is bevind dat die waterkwaliteitsveranderlikes meestal binne die voorgestelde grense van die waterkwaliteitsriglyne was, behalwe die fosfaatkonsentrasies, wat aansienlik hoer was as die toelaatbare 1 mg/1 vir afloop water. Hierdie hoe fosfaatkonsentrasies in die water, kan toegeskryf word aan afvloei vanaf gekombineerde rioolsuiweringswerke in the omgewing van die twee lokaliteite. Rou riool vanaf informele huisvestings en afvloei vanaf landerye het ook hiertoe bygedra. Met die uitsondering van mangaan, koper en sink, was die metaalkonsentrasies in die water hoer as die toelaatbare perke. Hierdie metale was egter teenwoordig in baie hoer konsentrasies in die sediment, veral yster en aluminium, met die hoogste konsentrasies in die kleinste partikel groottes van die sediment. Die hoe metaalkonsentrasies in die water en sediment dui op 'n bepaalde vlak van metaalbesoedeling, maar omdat die water van die Olifants Rivier redelik hard is, is die toestande nie noodwendig toksies vir die akwatiese lewe nie. Die algemene vlak van bioakkumulering van die geselekteerde metale is, in afnemende konsentrasies, soos volg: Fe, Cr, Al, Ni, Pb, Zn, Mn en Cu in die water, en: Al, Fe, Mn, Cu, Cr, Ni, Zn en Pb in die sediment met uitsonderlike hoe konsentrasies van aluminium en yster. Die verskillende organe en weefseltipes van die twee spesies, het die metale op verskillende vlakke bevat en daar kon duidelik afgelei word watter organe of weefseltipes die hoogste konsentrasies van 'n bepaalde metaal ge-akkumuleer het. Die kieu- en lewerweefsel het die hoogste konsentrasies van die metale getoon, met koper- en ysterkonsentrasies, die hoogste in die lewer. Hierdie organe kan dus altyd nuttig gebruik word vir die bepaling van die bioakkumulering van die genoemde metale in vis. Alhoewel die laagste metaal konsentrasies in die spier- en velweefsel teenwoordig is, is dit noodsaaklik dat hierdie weefsels altyd ingesluit moet word in algemene biomoniteringsprogramme, aangesien dit die eetbare deel van die vis is. Die bevolking van informele behuising benut vis uit die natuurlike strome as voedselbron. Die geskiktheid van die vis as

vii proteIen bron moet daarom vasgestel word. Ten spyte van die hoe metaalkonsentrasies in organe soos die kieue en lewer is die metaalvlakke in die vel- en spierweefsel binne perke. Hierdie vis kan daarom as voedselbron benut word. Metaal konsentrasies moet daarom altyd gemoniteer word, aangesien lokaliteit KOR1 afvloei ontvang vanaf gekombineerde rioolsuiweringswerke (stroom-op), sowel as rou riool vanaf die informele huisvestings. Afloopwater van verskeie kragsentrales, industries en stedelike gebiede en landelike omgewing hou ook 'n besoedelings gevaar in. Bykomend tot hierdie bronne, ontvang lokaliteit OR1 nog afloopwater vanaf steenkoolmyne in die area. Die resultate het 'n athanldikheid- van die bioakkumulering van die meeste metale in die organe en weefsel van die twee spesies getoon, met die spesifieke spesie en lengtes van die visse, sowel as die lokaliteite waar hierdie visse gevang is. Waterkwaliteitseienskappe beinvloed die chemiese vorme, beskikbaarheid en toksisiteit van verskeie metale vir akwatiese organismes. Daarom is die daarstelling van 'n algemene biomoniteringsprogram van groot belang, waar die hidrologiese en geomorfologiese eienskappe, die chemiese en fisiese water kwaliteit en die oewer plantegroei ook in ag geneem word. Hierdie faktore bepaal die eienskappe van akwatiese ekosisteme.

Identifisering van die effekte van chemikaliee op lewende organismes is belangrik om sodoende die metaalbesoedeling te kan beheer. Bykomend tot die veldstudie, is toksisiteitstoetse ook in die laboratorium uitgevoer, om sodoende die effekte van laer pH vlakke en verhoogde konsentrasies van aluminium op die hematologie, osmoregulering en koolhidraatmetabolisme van die bloukurpur, Oreochromis mossambicus, te bepaal. Die pH waarde van 5.2 is gebruik aangesien die maksimum toksisiteit van aluminium tussen pH 5.0 en 5.5 aangetref word. Aluminium het verskillende effekte by verskillende pH waardes en blootstelling aan 'n lae pH is gedoen, aangesien versuring die oplossing van aluminium in akwatiese sisteme veroorsaak. Die aluminiumkonsentrasies gebruik in hierdie studie, is gekies volgens die konsentrasies teenwoordig in die Olifantsrivier en die sytakke soos gedurende die huidige studie vasgestel. Die resultate het definitiewe

viii subletale effekte getoon, veral op koolhidraatmetabolisme. Die pH-vlakke by die twee lokaliteite wat bestudeer is, is aansienlik hoer as die pH waarby die eksperimente uitgevoer is. Geen ernstige aluminiumtoksisiteit vir akwatiese organismes kan daarom by die betrokke twee lokaliteite verwag word nie. Die lae pH by ander lokaliteite in die opvangsgebied van die Olifantsrivier en verlaging van die pH vlakke by lokaliteite KOR1 en OR1, as gevolg van suur afloop vanaf myne of laagvloei van die rivier, kan die voortbestaan van die vis bedreig. Die rede hiervoor is die hoe aluminiumkonsentrasies by hierdie twee lokaliteite in die sediment en oplossing daarvan in die water deur verlaagde pH wat tot visvrektes kan lei.

Uit die oogpunt van voeding en ontspanning, beslaan vis die belangrikste posisie in die akwatiese ekosisteem en daarom is die vlak van metaalkonsentrasies in visweefsel vanselfsprekend van groot belang. Hoe konsentrasies van metale is toksies vir die ekosisteem as 'n geheel en veral vir die mens wat die eindverbruiker in 'n verskeidenheid voedselkettings is. Verdere navorsing is nodig, aangesien hoe konsentrasies van metale in die weefsels en organe van die vis, asook in die sediment en water voorkom. Navorsing behoort op die bronne van besoedeling in die studie area toegespits te wees sodat die impak daarvan op die agteruitgang van die omgewing vasgestel kan word. Vanuit die bevindinge van hierdie studie is dit duidelik dat 'n gereelde moniteringsprogram wat waterkwaliteit en bioakkumulering insluit, deur die Departement van Waterwese en Bosbou oorweeg behoort te word.

ix - TABLE OF CONTENTS -

INTRODUCTION 1-1

1.1 References 1 - 7

THE OLIFANTS RIVER BASIN AND LOCALITY DESCRIPTION 2 - 1

2.1 General description 2-1 2.2 Sources of water 2-3 2.2.1 Groundwater 2-3 2.2.2 Surface water 2-4 2.2.3 Reuse of effluent 2-7 2.3 Water user sectors 2-7 2.3.1 Afforestation 2-7 2.3.2 Power generation 2-10 2.3.3 Mining 2-10 2.3.4 Irrigation 2-11 2.3.5 Stock watering 2-11 2.3.6 Domestic and industrial 2-11 2.4 Water quality 2-12 2.5 The study area 2-17 2.6 References 2-20

WATER AND SEDIMENT 3 - 1

3.1 Introduction 3-1 3.2 Materials and methods 3-4 3.2.1 Water 3-4 3.2.2 Sediment 3-5 3.2.3 Data processing 3-7 3.3 Results 3-7 3.3.1 Water 3-7 3.3.2 Sediment 3-23 3.4 Discussion 3-35 3.4.1 Physical and chemical characteristics of the river water 3-35 3.4.2 Metal concentrations in the water and sediment 3-43 3.5 Conclusion 3-46 3.6 References 3-49 3.7 Appendixes 3-56

1 BIOACCUMULATION OF ZINC AND COPPER IN THE TISSUES AND ORGANS OF GLARUS GARIEPINUS AND LABEO UMBRATUS 4-1

4.1 Introduction 4-1 4.2 Materials and methods 4-4 4.2.1 Field sampling 4-4 4.2.2 Laboratory procedures 4-4 4.2.3 Statistical procedures 4-8 4.3 Results 4-9 4.3.1 Fish size 4-9 4.3.2 Differences in bioaccumulation of zinc and copper in the different tissues/organs 4-9 4.3.3 Species differences 4-16 4.3.4 Relationship between lengths and zinc and copper concentrations 4-20 4.3.5 Differences between males and females 4-20 4.3.6 Seasonal differences 4-26 4.3.7 Localities differences 4-26 4.4 Discussion 4-30 4.5 Conclusion 4-40 4.6 References 4-42 4.7 Appendixes 4-48

BIOACCUMULATION OF MANGANESE AND LEAD IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS AND LABEO UMBRATUS 5-1

5.1 Introduction 5-1 5.2 Materials and methods 5-3 5.3 Results 5-3 5.3.1 Differences in bioaccumulation of manganese and lead in the different tissues/organs 5-3 5.3.2 Species differences .5-7 5.3.3 Relationship between lengths and manganese and lead concentrations 5-11 5.3.4 Differences between males and females 5-11 5.3.5 Seasonal differences 5-11 5.3.6 Localities differences 5-15 5.4 Discussion 5-19 5.5 Conclusion 5-29 5.6 References 5-30 5.7 Appendixes 5-33

2 BIOACCUMULATION OF CHROMIUM AND NICKEL IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS AND LABEO UMBRATUS 6-1

6.1 Introduction 6-1 6.2 Materials and methods 6-3 6.3 Results 6-3 6.3.1 Bioaccumulation of chromium and nickel in the different tissues/organs 6-3 6.3.2 Species differences 6-11 6.3.3 Relationship between lengths and chromium and nickel concentrations 6-11 6.3.4 Differences between males and females 6-11 6.3.5 Seasonal differences .6-14 6.3.6 Localities differences 6-14 6.4 Discussion 6-21 6.5 Conclusion .6-27 6.6 References 6-29 6.7 Appendixes 6-34

BIOACCUMULATION OF ALUMINIUM AND IRON IN THE TISSUES AND ORGANS OF CLAMS GARIEPINUS AND LABEO UMBRATUS 7-1

7.1 Introduction 7-1 7.2 Materials and methods 7-3 7.3 Results 7-3 7.3.1 Differences in bioaccumulation of aluminium and iron in the different tissues/organs 7-3 7.3.2 Species differences 7-11 7.3.3 Relationship between lengths and aluminium and iron concentrations 7-11 7.3.4 Differences between males and females 7-16 7.3.5 Seasonal differences 7-16 7.3.6 Localities differences 7-16 7.4 Discussion 7-20 7.5 Conclusion 7-27 7.6 References 7-28 7.7 Appendixes 7-32

THE EFFECTS OF LOW PH AND ALUMINIUM ON THE HAEMATOLOGY, OMOREGULATION AND CARBOHYDRATE METABOLISM OF OREOCHROMIS MOSSAMBICUS 8-1

8.1 Introduction 8-1

3 8.2 Materials and methods 8-3 8.2.1 Experimental procedure 8-6 8.2.1.1 Exposure of the test organism 8-6 8.2.1.2 Controls 8-16 8.2.1.3 Blood sampling 8-16 8.2.1.4 Measurement of variables .8-16 8.2.1.5 Data processing 8-20 8.3 Results 8-20 8.4 Discussion 8-26 8.5 Conclusion 8-38 8.6 References 8-40 8.7 Appendixes .8-47

9. SUMMARY AND RECOMMENDATIONS 9 - 1

9.1 General summary 9-1 9.1.1 Water and sediment .9-1 9.1.2 Bioaccumulation of the selected metals in the tissues and organs of C. gariepinus and L.umbratus 9-3 9.1.3 Toxicity tests of low pH and elevated aluminium concentrations on the haematology, osmoregulation and carbohydrate metabolism of the Mozambique tilapia, Oreochromis mossambicus 9-6 9.2 Recommendations 9-7

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INTRODUCTION

The use of metals for industrial, mining and agricultural purposes and the subsequent occurrence as trace contaminants, have resulted in increased loadings thereof in the aquatic environment. Diffuse and point source pollution from these activities contribute significantly towards metal and organic contamination, acidification and mineralization of the surface waters. Many potential contaminants, for example metals, ultimately find their way into surface waters which are the natural habitats for a large variety of aquatic species. It is therefore of primary importance to determine the effects of these contaminants on the water and sediment quality, as well as the degree of bioaccumulation of these pollutants by the fish and other aquatic biota.

All metal ions are potentially harmful to organisms at some level of exposure and absorption. The ecological significance of metals stems from their general toxicity and the fact that they are non-biodegradable. Metals vary widely in their physical, chemical and biological properties and can be found in several different forms after entering the water (Fig 1-1). In the aquatic environment, metals may exist in the dissolved form as free hydrated ions; as complex and chelated ions with inorganic ligands, including OH -, C032-, Cl- or as organic ligands including amines, proteins, humic acids and fulvic acids, or as part of organic molecules. Metals can also exist in the particulate form as colloidal complexes or aggregates, adsorbed on different types of particles, precipitated (for example as metal coatings on particles), or incorporated in organic particles such as plankton.

1-1 •

Introduction

Metal form -) Inorganic Soluble -> Ion Complex --+ Chelate -+ Molecule

-> Organic Particulate -■ Colloid Adsorption -> Suspended

-> Carrier particles -+ Mineral detritus Clay minerals Humic substances Lipids --* Residual organics --* Hydrous Fe/Mn oxides Carbonates

Presence of other substances Synergism -> Addition Antagonism

Water system -> Geology --+ Temperature -> Basin morphometry -> pH > Hydrology -> Dissolved oxygen Light Transparency Salinity -> Alkalinity Level of bioproduction Humic level Sediment trapping Biological system -> Age Size -> Food supply Activity -> System of protection Adoption to metals

Fig 1 - 1 Flow diagram of the factors affecting the potential uptake of metals in fish

1-2 Introduction

The toxicity of metals is linked to the specific forms of these metals in the aquatic medium, which are influenced by interacting factors, such as temperature, pH and water hardness. These environmental factors determine the chemical speciation of the metals and consequently the bioavailability of these metals to the aquatic organisms (Abel, 1989). Interaction between pollutants, the developmental stage of the organism and interspecific variations in susceptibility to metal, are other factors which also influence metal toxicity to aquatic organisms (Hellawell, 1986).

The uptake and retention of chemicals in the body of an organism, defined as bioaccumulation, may be via absorption through the gills (bioconcentration) and/or through ingestion of contaminating food (biomagnification) (EPA, 1985). The concentration of the metals in aquatic organisms differ, as it reflects the nett effect of two competing processes, namely that of uptake, and that of elimination. The balance between these two will depend on the ambient water concentration and the relative rate of these two processes. Bioaccumulation can be influenced by factors relating to the organism itself, e.g. species, physiological condition, growth, age, sex, pollutant interactions and the physical/chemical properties of the environment (Mance, 1987).

Metal accumulating strategies range from the strong accumulation of all the metals acquired, to the regulation of the body metal concentration to an almost constant level, by means of the balancing of the metal uptake and excretion. Both the toxicity and bioaccumulation potential of a chemical, are greatly affected by the rate of elimination. If an unchanged chemical can be eliminated rapidly, accumulation of residues will not take place. In vertebrates, elimination can take place via several routes, including transport over an integument or respiratory surface, secretion in bile, the skin (mucus), the gonads, and through excretion from the kidney in the urine (Heath, 1987).

The tendency of many aquatic organisms to accumulate metals from their environment, has led to the use of some of these organisms for the assessment of metal pollution.

1-3 Introduction

Fish are useful in biological monitoring, because they are known to accumulate metals in their organs and tissues, they are readily identified and available, they can be sampled easily and quantitatively and they have a cosmopolitan distribution (Hellawell, 1986). Fish can therefore provide valuable information in addition to water and sediment data. Furthermore, the monitoring of bioaccumulation of metals and organic compounds is essential, because fish and other organisms are used as food. This aids in the protection against the consumption of contaminated food and the detected levels can be judged against standards set for food in general (Mance, 1987). The accumulate levels in aquatic organisms may also pose a risk to animals such as birds and mammals that feed on them, resulting in a decline in population numbers (Lloyd, 1992). Monitoring can be both spatial and temporal (Mance, 1987), where spatial monitoring of bioaccumulation may produce data that would identify potential unknown areas with high concentrations while at known discharges it will provide some information regarding the area being effected. Temporal changes in bioaccumulation will provide information regarding the trend of bioaccumulation, which will in turn be used to identify stability, improvement or deterioration (Mance, 1987).

Despite the fact that South Africa has various mining, industrial, domestic and agricultural activities, the possible metal contamination from these activities has only recently received attention. These studies include research on the various metal concentrations in the water and sediments of South African rivers (Roux, Badenhorst, Du Preez & Steyn, 1994; Seymore, Du Preez, Van Vuren, Deacon & Strydom, 1994) and dams (Coetzee, 1993) as well as metal concentrations in fish tissues and organs (Du Preez & Steyn, 1992; Van den Heever & Frey, 1994; Seymore, Du Preez & Van Vuren, 1995; Van den Heever & Frey, 1996) and metal concentrations in freshwater crabs from industrial and mine-polluted freshwater ecosystems (Steenkamp, Du Preez & Schoonbee, 1994; Steenkamp, Du Preez, Schoonbee & Van Eeden, 1994). Studies were also performed on the effects of metals on the physiology of freshwater fish in South Africa (Bezuidenhout, Schoonbee & De Wet, 1990; Wepener, Van Vuren & Du Preez, 1992;

1-4 Introduction

Van Vuren, Van der Merwe & Du Preez, 1994; Nussey, Van Vuren & Du Preez, 1995a, b, c).

The present study focuses on the concentrations of selected water quality variables and metal concentrations in sediment, water and fish from the Olifants River and Klein Olifants River. Some of the tributaries of the Olifants River have a very low pH and in combination with metals like aluminium, these were thought to be responsible for fish kills (Dr P Kempster, personal communication, 1995). The effects of aluminium and low pH on the haematology and osmoregulation of the Mozambique tilapia, Oreochromis mossambicus were also investigated.

The specific objectives were: to assess the overall water quality, by determining the selected water quality variables in the Olifants River before and after the confluence with the Klein Olifants River to investigate possible correlations between the metal concentrations in the water and the sediment to determine the bioaccumulation order of selected metals in the different tissues and organs of two freshwater fish species, namely the catfish, Clarias gariepinus and the moggel, Labeo umbratus. This data was used to evaluate the difference between the localities, seasons, species and different tissues and organs, as well as the sex of the fish to perform sublethal toxicity tests, under controlled laboratory conditions, exposing the Mozambique tilapia, Oreochromis mossambicus to low pH and aluminium levels.

This study forms part of a much larger project, based on the potential pollution of the Olifants River and its tributaries and can be of aid in evaluating the water quality of the Olifants River, as little information is available concerning the metal concentrations in

1-5 Introduction

the upper Olifants River catchment. Eighteen sampling sites have been selected along the river, of which two of these sites, one in the Olifants River, before its confluence with the Klein Olifants River and the other in the Klein Olifants River, will receive attention in this study.

1-6 Introduction

REFERENCES

ABEL, P.D. (1989). Water Pollution Biology. Ellis Horwood Limited Publishers, Chichester. 231 pp.

BEZUIDENHOUT, L.M., SCHOONBEE, H.J. & DE WET, L.P.D. (1990). Heavy metal content in organs of the African sharptooth catfish, Clarias gariepinus (Burchell), from a Transvaal lake affected by mine and industrial effluents. Part 1. Zinc and copper. Water SA. 16(2):125 - 129.

COETZEE, P.P. (1993). Determination and speciation of heavy metals in sediments of the Hartbeespoort Dam by sequential chemical extraction. Water SA. 19(4):291.

DU PREEZ, H.H. & STEYN, G.J. (1992). A preliminary investigation of the concentration of selected metals in the tissues and organs of the tigerfish (Hydrocynus vittatus) from the Olifants River, Kruger National Park, South Africa. Water SA. 18(2):131 - 136.

EPA (1985). Technical support document of water quality-based toxic control. Office of Water Regulation and Standards. U.S. Environmental Protection Agency, Washington, D.C.

HEATH, A.G. (1987). Water Pollution and Fish Physiology. CRC Press, Inc., Florida. 245 pp.

1-7 Introduction

HELLAWELL, J.M. (1986). Biological Indicators of Freshwater Pollution and Environmental Management. Elsevier Applied Science Publishers Ltd., London. 546 pp.

LLOYD, R. (1992). Pollution and freshwater fish. Fishing News Books Publishers. London. pp. 176.

MANCE, G. (1987). Pollution threat of heavy metals in aquatic environment. Pollution Monitoring Series. Elsevier Applied Science Publishers, London. pp. 372.

NUSSEY, G., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1995a). Effect of copper on blood coagulation of Oreochromis mossambicus (Cichlidae). Comp. Biochem. Physiol. 111C(3):359 - 367.

NUSSEY, G., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1995b). Effect of copper on the haematology and osmoregulation of the Mozambique tilapia, Oreochromis mossambicus (Cichlidae). Comp. Biochem. Physiol. 111C(3):369 - 380.

NUSSEY, G., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1995c). Effect of copper on the differential white blood cell counts of the Mozambique tilapia (Oreochromis mossambicus). Comp. Biochem. Physiol. 111C(3):381 - 388.

ROUX, D.J., BADENHORST, J.E.E, DU PREEZ, H.H. & STEYN, G.J. (1994). Note on the occurrence of selected trace metals and organic compounds in water, sediment and biota of the Crocodile River, Eastern Transvaal, South Africa. Water SA. 20(4):333 - 340.

1-8 Introduction

SEYMORE, T., DU PREEZ, H.H., VAN VUREN J.H.J., DEACON, A. & STRYDOM, G. (1994). Variations in selected water quality variables and metal concentrations in the sediment of the lower Olifants and Selati rivers, South Africa. KOEDOE. 37(2):1 - 18.

SEYMORE, T., DU PREEZ, H.H. & VAN VUREN, J.H.J. (1995). Manganese, lead and strontium bioaccumulation in the tissues of the yellowfish, Barbus marequensis from the lower Olifants River, Eastern Transvaal. Water SA. 21(2):159 - 172.

STEENKAMP, V.E., DU PREEZ, H.H. & SCHOONBEE, H.J. (1994). Bioaccumulation of copper in the tissues of Potamonautes warreni (Calman) (Crustacea, Decapoda, Branchiura), from industrial, mine and sewage-polluted freshwater ecosystems. S. Afr. J. Zool. 29(2):152 - 161.

STEENKAMP, V.E., DU PREEZ, H.H., SCHOONBEE, H.J. & VAN EEDEN, P.H. (1994). Bioaccumulation of manganese in selected tissues of the freshwater crab, Potamonautes warreni (Calman), from industrial and mine-polluted freshwater ecosystems. Hydrobiologia. 288:137 - 150.

VAN DEN HEEVER, D.J. & FREY, B.J. (1994). Human health aspects of the metals zinc and copper in the tissue of the African sharptooth catfish, Clarias gariepinus, kept in treated sewage effluent and in the Krugersdrift Dam. Water SA. 20(3):205 - 212.

VAN DEN HEEVER, D.J. & FREY, B.J. (1996). Human health aspects of certain metals in tissue of the African sharptooth catfish, Clarias gariepinus, kept in

1-9 Introduction

treated sewage effluent and in the Krugersdrift Dam: Iron and Manganese. Water SA. 22(1):67 - 72.

VAN DEN HEEVER, D.J. & FREY, B.J. (1996). Human health aspects of certain metals in tissue of the African Sharptooth catfish, Clarias gariepinus, kept in treated sewage effluent and in the Krugersdrift Dam: Chromium and Mercury. Water SA. 22(1):73 - 78.

VAN VUREN, J.H.J., VAN DER MERWE, M. & DU PREEZ, H.H. (1994).The effect of copper on the blood chemistry of Clarias gariepinus (Clariidae). Exotoxicology and Environmental Safety. 29:187 - 199.

WEPENER, V., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1992). The effect of hexavalent chromium at different pH values on the haematology of Tilapia spamnanii (Cichlidae). Comp. Biochem. Physiol. 101C(2):375 - 381.

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• CHAPTER 2

THE OLIFANTS RIVER BASIN AND LOCALITY DESCRIPTION

2.1 GENERAL DESCRIPTION

The Olifants River originates in the Bethal-Trichardt area and flows in an easterly direction before crossing the Kruger National Park into Mozambique (Fig 2-1). The Olifants River basin in the Transvaal drains a large area of over 54 575 km' (Theron, Prinsloo, Grimsehl & Pullen, 1991). The main tributaries of the Olifants River in the upper catchment of the river system, from the source near Bethal, to its confluence with the Wilge River, north of Witbank, are the Klein Olifants River (with its tributaries being the Keerom Spruit and Rietkuil Spruit), the Steenkool Spruit, Trichardt Spruit and Riet Spruit (which flow into the Olifants River before it joins with the Klein Olifants River) and the Klip Spruit, which flows into the Olifants River before its confluence with the Wilge River.

Topographically, the catchment falls mainly on the undulating Highveld between 1 200m to 1 800 m above mean sea level. The climate is warm and the average temperatures show moderate fluctuation, with average summer temperatures varying between 18 °C and 26 °C, while the winter averages are between 0 °C and 13 °C. Frost may occur from May to September with seasonal rainfall occurring predominantly in summer from October to March (Theron et ed., 1991). The mean rainfall for the area is in the range of 650 - 800 mm. Evaporation is in the range of 75 - 190 mm per month with June the month with the lowest and December the highest evaporation.

2-1 The Upper Olifants River Catchment

Broo khorsisp ru it •

A UPPER OUFANTS RIVER CATCHMENT WILGE RIVER CATCHMENT C ELANDS RIVER CATCHMENT MIDDLE OLIFANTS RIVER CATCHMENT LOWER OLIFANTS RIVER CATCHMENT F BLYDE RIVER CATCHMENT STEELPOORT CATCHMENT OLIFANTS RIVER - LOSKOP CATCHMENT

FIG 2-1 THE OLIFANTS RIVER CATCHMENT, INDICATING TIIE DIFFERENT SUBCATCHMENTS, IN PARTICULAR THE UPPER OLIFANTS RIVER CATCHMENT (STUDY AREA)

2-2 The Upper Olifants River Catchment

The vegetation can be described as a grassland biome which can be seen as an ecological unit which represents large, natural and homogeneous areas of biotic and abiotic features. The biotic component is closely related to physical factors, particularly soil type and climate (Steffen, Robertson & Kirsten, 1991). Two veld types, namely Turf Highveld type, which is a pure grassveld that occurs in the upper reaches of the Olifants River tributaries, Steenkool Spruit and Trichardt Spruit and the false grassveld type, Bankenveld, which dominates the rest of the area (from the perspective of Acock's veld types). The vegetation has a uniform physical appearance and is dominated by hemicryptophytes of Poaceae and the number of threatened plant species are less than ten (Theron et aL, 1991). Soil erosion is also limited because of the high vegetation cover.

The main geological outcrops are the bushveld Complex, the Waterberg Group, the Karoo Supergroup and the major Dyke/Sill intrusives. Structurally the most significant feature within the catchment is the intrusion of the Bushveld Complex into older Transvaal sediments. Dolerite, mostly in the form of sills, has intruded the Karoo rocks in the catchment. The chemical make-up and geometry of these sill intrusives are not conducive to groundwater development, except along the contact zones where sedimentary strata are often fractured (Theron et aL, 1991).

2.2 SOURCES OF WATER

2.2.1 GROUNDWATER Two hydrological regions can be found, namely the Karoo Sequence Region and the Granitic Region. The Karoo Sequence Region in the Middelburg and Witbank area, is made up of interbedded shale, sandstone and coal strata, with intrusive dolerite and

2-3 The Upper Olifants River Catchment

ultramafic dykes, which will provide good borehole sites, and sills. The Granitic Region is confined to all areas where granite, gneiss and granitic type rocks are found in the area. Boreholes are at present an important source of water supply to rural, domestic and stock watering and is expected to yield in the region of 1.5 1/s to 5 1/s. Although roughly 70 % of the Upper Olifants River Catchment has high to moderate groundwater potential, dry boreholes are becoming more plentiful. The chemistry of groundwater is greatly influenced by the media with which it has been in contact, as well as the duration of the contact. The dissolution of minerals from a rock produces water which contains the same minerals. Two or more types of water from different geological origins can also combine to produce a composite groundwater assemblage. Possible groundwater types associated with various geological terrains are thus subject to migration influences (Theron et al., 1991). Sodium,' calcium, silica, iron and magnesium concentrations will dominate groundwater originating from the granitic type rocks and the degree of mineralization is expected to be moderately high. The rocks of the Bushveld complex are made up of ferro-magnesian silicates and the resulting groundwater will be mineralized with iron, magnesium, calcium, sodium and phosphate dominating (Theron et al., 1991)

2.2.2 SURFACE WATER Dams play an important role in supplying water at a high level of assurance. In the Olifants River catchment in South Africa, thirty major dams, with a total percent storage capacity of 1 064.87 million m3 are on record. Over 2 000 dams can be found with a surface area of less than 1 ha and a volume of less than 20 000 m 3 (Theron et al., 1991). The dams in the Upper Olifants River catchment can be divided into three categories (Fig 2-2):

2-4 The Upper Olifants River Catchment

FIG 2-2 PERCENTAGE OF WATER STORAGE (DAMS) IN THE UPPER OLIFANTS RIVER CATCHMENT

2-5 The Upper Olifants River Catchment

Major dams which have a great impact on the runoff regime in a river and can be defined as those that fulfill a major function in the community and have a capacity greater than 2 million m 3. In the upper catchment, the Witbank, Middelburg, Doornspoort, Riet Spruit and Trichardsfontein Dams fall into this category. Small dams, which are defined as those individually insignificant with regards to the hydrology of a basin. Small dams have a capacity of less than 100 000 m 3. In the upper catchment, small dams are mainly used for stock-watering and soil conservation (Theron et al., 1991). Minor dams which are defined as those dams which are significant only on a local level. There are over 200 minor dams in the Olifants River basin, each with a capacity of between 100 000 m 3 and 2 million m3 and generally supply only one institution with water. In the Upper Olifants River Catchment, minor dams are used for irrigation, mining, stock watering, recreation and power generation (Theron et al., 1991).

Rainfall is the most important determinant of runoff and occurs predominantly in summer in the form of showers and thunder storms. Runoff at the major impoundments, Witbank Dam and Middelburg Dam decreased from 125 million m 3/annum at Witbank Dam and almost 44 million m3/annum at Middelburg Dam to 107.4 and 37 million m3/annum respectively. This is due to the effect of afforestation, irrigation and evaporation from minor and small dams on the natural runoff at the dams. It is expected that the runoff will decrease further to 104 and 36 million m 3/annum respectively by the year 2010, due to possible increase in irrigation and minor and small dams.

2-6 The Upper Olifants River Catchment

2.23 REUSE OF EFFLUENT The reuse of effluent as a source of water is important, particularly in areas of urban and industrial growth. Effluent should therefore be considered as a supplementary source that can be purified to different standards for various categories of use, e.g. industrial consumers which are presently supplied with raw water from Witbank or Middelburg Dams. The reuse of effluent would however involve the cost of additional purification and also the installation of pumping systems, therefore there is at present no significant merit in adopting the reuse of effluent (Theron et al., 1991).

2.3 WATER USER SECTORS

Most of the water in the Upper Olifants River catchment is used for afforestation, mining and power generation, irrigation, domestic and industrial purposes, as well as for maintaining the ecological systems. The locations of all the major abstractions are showed in Figure 2-3 (See Table 2-1).

2.3.1 AFFORESTATION Although conditions are not ideal for afforestation due to low rainfall, approximately 17 680 ha of exotic afforestation occurs in the Upper Olifants River. Catchment, which is concentrated around Witbank and Middelburg in the catchments of the Olifants and Klein Olifants Rivers respectively. No significant indigenous forests occur and only 7 300 ha are managed as commercial forests. The balance comprises stands of unattended wattles, pines and gums. The timber produced is used for mining poles, pulping and charcoal. No new afforestation is expected, as natural climatic conditions limit the suitability of the area for afforestation. Present exotic plantations decrease the mean annual runoff by ± 5% (Theron et aL, 1991).

2-7 The Upper Olifants River Catchment

BETIIAL LEGEND O MINES Trkbanistroatela Dam MUNICIPALITIES TRICDARDT A TOWNS & VILLAGES

POWER STATIONS

FIG 2-3 TUE MAJOR ABSTRACTION POINTS IN THE UPPER OLIFANTS RIVER CATCHMENT, WITH THE MINES INDICATED AS A - M, THE TOWNS AS N & 0 AND THE POWER STATION AS P. FOR A DESCRIPTION, SEE TABLE 2-1

2-8 The Upper Olifants River Catchment

TABLE 2-1 MAJOR ABSTRACTIONS IN THE STUDY AREA

A Albion colliery Steenkool Spruit B Douglas colliery (Douglas Olifants River (Douglas Dam) section) Imported C Douglas colliery (Wolwekrans Olifants River section) D Douglas colliery (Van Dyk Drifts Olifants River section) E New Clydesdale Olifants River F Tavistock colliery Olifants River/Phoenix Dam G Transvaal Navigation colliery Olifants River/Underground H Witbank Consolidated colliery Zaaiwater Spruit I Arnot colliery Klein Olifants River/Imported Kleinkopje colliery Olifants River K Riet Spruit Opencast Services Riet Spruit Dam L Duhva Opencast Services Witbank Dam/Imported M Zwakfontein Sand Trichardt Spruit/Underground N Witbank Witbank Dam 0 Middelburg Middelburg Dam P Duhva Power Station Witbank Dam

2-9 The Upper Olifants River Catchment

2.3.2 POWER GENERATION The upper catchment of the Olifants River system, with its tributaries, drains a portion of the Mpumalanga highveld, where most of the thermal power stations in the country are located, as six of the eight power stations are situated in the upper catchment. These power stations receive imported water, with the Komati- , Hendrina- , Arnot- and Duhva power stations receiving water from the Komati system and Kriel- and Matla power stations from the Usutu system. Each power station is designed to use water of a specific quality, with the TDS concentrations being one of the most important variables in determining the necessity for pre-treatment of the water, used for purposes like the cooling of the circuits, for steam generating circuits utilizing demineralised water and for the transportation of ash from boilers to ash dams. Power stations are one of the most important sources of heat pollution, which can have serious effects on the aquatic environment and its organisms, since it may change the natural temperature range of the water, while a decrease in dissolved oxygen is also experienced with increasing temperatures.

23.3 MINING The operating mines in this region consist of 37 coal, 6 brick, 17 sand, 4 felsite and 7 clay mines. These mines used both surface and underground water, but now consume mainly imported water from neighbouring catchments. Mining consumption presently totals about 22 million m3 water per year of which 10.7 million m 3 are imported and 7.2 million m3 are from surface water sources (Theron et aL, 1991). With the development of new power stations in the area, the coal mining sector also expanded, with the subsequent increased demand for water.

2-10 The Upper Olifants River Catchment

2.3.4 IRRIGATION Irrigation is a major water use sector in the Upper Olifants River catchment and at present 4 760 ha of land is irrigated. Large areas under irrigation carry fodder crops and posture to support the extensive livestock farming activities. Maize and potatoes, followed by vegetables, lucerne, groundnuts and dried beans are irrigated in this area. Water used for irrigation comes from the Olifants River and its tributaries. Development for irrigation is estimated to be about 1 800 ha which will increase the water demands by ± 9.9 million m3 per year, making the total water required for irrigation almost 35 million m3 per year. Water availability is however the main restriction on irrigation use, but if not, 25 million m 3 per year would be abstracted in the subcatchment (Theron et aL, 1991).

Above the Witbank Dam however, difficulty is experienced with the water quality, e.g. runoff in the Koring Spruit is unsuitable for irrigation, while in a number of other areas low flows, in particular, can be of such poor quality that crops and soils cannot be irrigated. Runoff in the Klein Olifants River, the Wilge River and Kranspoort Spruit, is however of good quality for irrigation purposes.

2.33 STOCK WATERING The highveld is traditionally a cattle farming district, with the large stock population in the Upper Olifants River catchment ± 457 000 large stock units (LSU) which use an estimated amount of 8.7 million m3 per year. Stock watering relies on surface water, springs and boreholes for water supply.

2.3.6 DOMESTIC AND INDUSTRIAL Seven towns are situated in the Upper Olifants River Catchment. The Witbank/Middelburg complex is a major industrial area with smaller towns Davel,

2-11 The Upper Olifants River Catchment

Kinross, Trichardt, Hendrina and Ogies. Several magisterial districts fall partly inside the catchment boundaries, namely Bethal, Belfast, Ermelo, Highveld Ridge, Middelburg and Witbank. The total domestic and industrial water demand is expected to be 65.8 million m3 for the year 2010, based on the probable population size (Theron et al., 1991). Of this projected demand, almost 67% will be by Witbank and the township, Kwa Guqu, from the Witbank Dam and 26% by Middelburg and the township, Mhluzi, from the Middelburg Dam.

A total of 36.6 million m3 water per year is presently used for domestic purposes in the Upper Olifants River Catchment, which can be divided into: 33.6 million m3 as surface water from the Bronkhorst Spruit Dam, the Middelburg Dam and the Witbank Dam, serving Ogies, Middelburg and Witbank respectively 2.3 million m3 from boreholes, used by the majority of the rural population. 0.9 million m3 is imported water supplied by Rand Water to Kinross and Trichardt, the Usutu-Vaal Scheme to Davel and by the Nooitgedacht-Komati pipeline to Hendrina (Theron et al., 1991).

The industrial towns of Middelburg and Witbank is expected to grow rapidly and due to the enormous population growth, accelerated water demand is therefore also expected.

2.4 WATER QUALITY

It is important to take into account the different sources of pollution, which can be divided into two categories:

2-12 The Upper Olifants River Catchment

1) POINT SOURCES OF POLLUTION Mines: Much of the pollution in this catchment is due to the extensive mining activities. Generally, coal, brick, sand, felsite and clay mines are found in the Olifants River catchment, with coal mines most commonly found in the upper catchment. Coal mines generally have an influence on the water pH, turbidity and total dissolved salts concentrations (Fig 2-4; See Table 2-2). Industries: Different industries have different effluent discharges, for example the vanadium industry contributes to the sulphate load in the water. The industries do not have permits to discharge effluents into any river and all their effluent is disposed off through the municipal sewage systems. Sewage treatment works: There are a number of sewage treatment works in the area: The Hendrina sewage treatment works, which is at present overloaded and it is necessary to double the size of the works to cope with the expansion of the Jonkerville township; Two sewage works at Kinross which discharge into tributaries of the Vaalbank Spruit; Newer works that are situated on the farm Zondagskraal and serve the western part of Kinross and the rehabilitation school; Middelburg or Boschkraans sewage works which is an activated sludge system; The Trichardt sewage works where the effluent is discharged into the Trichardt Spruit; The Riverview works that discharges into the Olifants River; The Ferrobank works which discharge into the Brug Spruit, a tributary of the Klip Spniit; The Naauwpoort sewage works where the effluent is discharged into a

2-13 The Upper Olifants River Catchment

BETHAL LEGEND

(--) MINES Trichardaroatela Data TRICHARDT MUNICIPALITIES A TOWNS & VILLAGES

0 POWER STATIONS

FIG 2-4 MAJOR EFFLUENT DISCHARGE POINTS IN THE UPPER 0 LIFANTS RIVER CATCHMENT, WITH THE MINES INDICATED AS A - I, THE MUNICIPALITIES AND TOWNS AS J - N AND THE POWER STATIONS AS 0 - R. FOR A DESCRIPTION, SEE TABLE 2-2

2-14 The Upper Olifants River Catchment

TABLE 2-2 MAJOR EFFLUENT DISCHARGE POINTS

A Albion colliery Steenkool Spruit B Tavistock collieries Zaaiwater Spruit C Transvaal Navigation Olifants River D Douglas colliery (Douglas section) Douglas Dam E Goedehoop colliery Olifants River F Blinkpan colliery Koring Spruit G Matla colliery Riet Spruit H Greenside colliery Witbank Dam I Middelburg Mine Services Niekerk Spruit J Witbank Olifants River/Brug Spruit K Middelburg Klein Olifants River L Hendrina Klein Olifants River Trichardt Trichardt Spruit N Kinross Vaalbank Spruit 0 Komati Power Station Koring Spruit P Kriel Power Station Steenkool Spruit

Q Arnot Power Station Bosman Spruit R Hendrina Power Station Woes-Alleen Spruit

2-15 The Upper Olifants River Catchment

spruit that flows into Witbank Dam; The Davel sewage effluent is processed in an oxidation ponds system, where the effluent is evaporated, thus there is no discharge into the river system.

The origin, volume and concentration of the effluent of these point sources can be quantified (Theron et al., 1991).

2) NON-POINT SOURCES OF POLLUTION Agriculture and forestry where irrigated areas are a diffuse source of nutrients such as nitrates and phosphates, as well as biocides. Surface runoff and high loads of suspended solids may be a result of the physical disturbance of soil and vegetation (Dallas & Day, 1993). Irrigation and subsequent evaporation of water from the land, can result in concentration of dissolved solids and therefore high TDS (Theron et al., 1991). Algal growth can be stimulated by nutrients, from disturbance of land and application of fertilizers and livestock excreta which enter receiving water bodies. Surface runoff from urban areas where pollutants such as soil, garden chemicals and pet wastes are carried to the river system via storm water drains. Potential sources also include road pavement materials, motor vehicles, atmospheric fallout, litter, domestic spraying and unauthorized dumping and washing. Contaminated groundwater can also seep to surface streams.

The volumes and concentrations of effluent from non-point sources of pollution cannot be quantified and the origin is diffuse (Theron et al., 1991).

2-16 The Upper Olifants River Catchment

2.5 THE STUDY AREA

The study area comprises of the upper part of the Olifants River, from Bethal to the confluence of the Olifants River with the Wilge River. The two sampling sites chosen for discussion in this study, is locality 13 (KOR1) in the Klein Olifants River and locality 14 (OR1) in the Olifants River (Fig 2-5). These two sampling sites were chosen to determine the water quality and the effect of the different effluent discharges in the Klein Olifants River before its confluence with the Olifants River (KOR1) as well as in the Olifants River after its confluence with all the streams receiving effluent from the mines, industries, towns and power stations in the area, before it reaches the Klein Olifants River. Fish were sampled at a locality in the Klein Olifants River, situated more upstream before it reaches the nature reserve and closer to the sewage works, in the first two months of the study. At this location, the plant growth on the river bank consists mainly of grass. The river bed consists of rocks with thick deposits of silt and algae on the rocks, due to effluent from the combined sewage works and deep pools occurring occasionally. As a result of small sample sizes and lack of species diversity, another locality was chosen approximately two kilometres downstream in the nature reserve (Botshabelo). At this location, the river is very deep and the river bank also consists mainly of grass and a few trees. The water quality was mainly the same as it was at the first location, but the habitat were much more favourable for the fish. Fish species such as the moggel, Labeo umbratus and the African sharptooth catfish, Clarias gariepinus are mainly found in this area.

This location is dependant on water from the Middelburg Dam and the sewage works located upstream. The point sources of pollution in this area are therefore mainly the combined sewage treatment works and domestic and industrial effluent. The sewage treatment works and a number of informal settlements alongside the river, are responsible for the nutrient load in the river, for example phosphates and

2-17 The Upper Olifants River Catchment

LEGEND

1 WITBANK DAM BEIHAL 2 HOESMAN SPRUIT 3 ROESMAN SPRUIT 4 OLIPANTS RIVER AT DUN VA Trichardtsfontein Dam S KORINO SPRUIT IRICHARDT 6 STEENKOOL SPRUIT 7 DAVEL 8 WOES- ALLEEN SPRUIT 9 MIDDELBURO DAM 10 SUURSTROOM 11 OLIPANTS RIVER 12 SPOOK SPRUIT 13 AASVOELKRANS 14 OLIPANTS RIVER LODGE FIG 2-5 LOCATIONS OF THE SAMPLING SITES IN THE. STUDY AREA

2-18 The Upper Olifants River Catchment

nitrates, which can lead to eutrophication. The industries in the area are point sources of pollution containing constituents such as calcium, sulphates, potassium, sodium and metals. Industries in the Middelburg area include steel works and engineering firms providing for the steel works and two large paint factories, which may contribute to surface runoff. The non-point sources of pollution in this area are mainly atmospheric deposition, rural and urban runoff and agriculture. Polluted ground water seeping to surface streams can also be a source of pollution.

Locality OR1 in the Olifants River is located at the Olifants River Lodge holiday resort. The river bed consists of deposits of silt and sand with occasional rocks. The vegetation on the river bank is very dense, consisting of grass, trees and reeds. Aquatic plants (Salvelinus and Potamageton spp.) and algae are common in this area. The point sources of pollution above this sampling point, are mainly the sewage works, mines and industries containing constituents such as fluorides, metals, phosphates, sulphates, calcium, magnesium, sodium and potassium. This locality receives water from the Witbank Dam and streams such as Suurstroom and Spook Spruit, receiving effluent from mines such as the Middelburg Mine Services, that flows into the Olifants River. Organic pollution may also occur at this point, which are mainly due to a number of geese living on the river bank. Industries located alongside the river in the Witbank area include steel works like Highveld Steel, Ferro Metals and Trans Alloys as well as service industries for the mines located in the area, a petrol depot and a brewery. Non-point sources of pollution include runoff from towns and agricultural activities along the river. Fish species found in this area are the moggel, L. umbratus, the catfish, C. gariepinus and the yellow fish Barbus marequensis.

2-19 The Upper Olifants River Catchment

2.6 REFERENCES

DALLAS, H.F. & DAY J.A. (1993). The effects of water quality variables on riverine ecosystems. A review (WRC Project number 351). ix pp.

STEFFEN, ROBERTSON & KIRSTEN INC (1991). Water Resources planning of the Olifants River Basin. Study of Development Potential and Management of the Water Resources. Annexure 1: Topography, climate, vegetation, wildlife and archaeology. DWA Report No. P.B000/00/1191. 54 pp.

THERON, PRINSLOO, GRIMSEHL & PULLEN CONSULTING ENGINEERS (1991). • Water Resources planning of the Olifants River Basin. Situation assessment. Volume 3 Part 1. Subcatchment B100. 158 pp. Water Resources planning of the Olifants River Basin. Water infrastructure. Annexure 4 Part 1. dams in subcatchments B100, B200, B310 and B320. 46 pp. Water Resources planning of the Olifants River Basin. Physical infrastructure and economic activities. Annexure 5. 45 pp. Water Resources planning of the Olifants River Basin. Demography of primary water users. Annexure 6. 5 pp. Water Resources planning of the Olifants River Basin. Water use for domestic, industrial, mining and power generating purposes. Annexure 8. 67 pp. Water Resources planning of the Olifants River Basin. Irrigation. Annexure 9 Part 1. Subcatchments B100, B200 and B320 upstream of Loskop Dam. 7 pp. Water Resources planning of the Olifants River Basin. Water availability from major dams. Annexure 16 Part 1. Basin upstream of Loskop Dam. 52 pp.

2-20 The Upper Olifants River Catchment

- Water Resources planning of the Olifants River Basin. Water quality. Annexure 19. 86 pp.

2-21 . •

. • !„, ... • ..... 6,- '.,;-`,, (,•if .,. ',' ‘.4‘..; - j'1,>, - ' .,-,', .-''

.k -...

" •

; • •.

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• .• CHAPTER 3

WATER AND SEDIMENT

3.1 INTRODUCTION

Water, which covers 71 % of the earth's surface, is a fundamental feature thereof, with the estimated volumes of water in storage being the oceans (97.6%), groundwater down to 1 km (2.07%), lakes and reservoirs (0.28%), saline lakes (0.007%), soil moisture (0.005%), biological moisture in plants and animals (0.005%), atmospheric moisture (0.001%), swamps and marshes (0.003%) and rivers and streams (0.0001%) (Wetzel, 1983). According to Wetzel, (1983), the freshwater supply is, however, in reality much less, because of several factors: Rainfall is not evenly distributed over the land surfaces, nor has man distributed himself proportionally in relation to water availability. The total consumption has increased exponentially with demophoric growth. The severe degradation of the water quality, stemming from demophoric growth and anthropogenic activities.

Water pollution can be defined as the presence of significant amounts of unnatural substances or abnormally high concentrations of natural constituents at a level that cause undesirable effects. It is important to note that even substances which are essential to life, for example metals such as copper and zinc, can be highly toxic when present in large amounts. Growing industrial and mining activities can cause severe effects on the environment. Domestic sewage, for example, may contain in addition to oxidizable material, detergents, nutrients, pathogens and metals (Mason, 1991). Metal ions in aquatic systems, are partitioned among different compartments of these

3-1 Water and Sediment

ecosystems, specifically among the water and sediment, where these metallic elements are important causes of environmental pollution. Metal ions are ubiquitous, readily dissolved in and transported by water, concentrated in sediment and taken up by aquatic organisms and strongly bound by sulthydryl groups of proteins. Metal contamination is usually associated with mines and metal smelters, urban and industrial development, atmospheric deposition near major highways, cities and industries and with regions receiving acid deposition (here metals may be leached by acid from the soil and bedrock). It is usually released in liquid wastes, or leached from solid waste disposed on land. Metals associated with fossil fuels, either as contaminants in coal and oil, or as fuel additives (for example, lead compounds) enter aquatic ecosystems via rain or dustfall.

Many contaminants of concern to water quality, for example metals, are found associated with particulate matter in the environment, for metals that do not remain soluble in the water are adsorbed and accumulated by bottom sediments, acting as a sink (Gardiner, 1974). Natural water systems possess a variety of pathways for metal deposition to the sediment. Two of the most important processes for the transfer of metals from the solution phase to the sediments are precipitation and adsorption. The relative balance between the two is dependant upon the degree of supersaturation and available solid surface area. The concentration of metals in the sediment depends on its pH, CaCO 3 and organic matter content, as changes in the pH of the water could have a direct bearing on the water solubility as well as the depositing capacity of such metals in the substrata of standing and flowing water ecosystems (Forstner & Prosi, 1979). The distribution of these metals in the bottom sediments, is also affected by several factors, including deposition, sorption, enrichment in organisms and diagenetic redistribution of trace elements (Forstner & Muller, 1975). It has been shown that iron and manganese oxides, sediment size distribution and organic material content of sediments influence the distribution of trace metals in aqueous environments (Chao & Theobald, 1976). Fine- grained sediments rich in clay and organic materials and high in iron and manganese oxides tend to sorb metals, in contrast to coarse textured sediments, poor in clay and

3-2 Water and Sediment

organic material and deplete of or low in iron and manganese oxides. Sediments can act both as carriers and potential sources of contaminants in aquatic systems which may even affect groundwater quality. Contaminants are present in the sediments and interstitial water in a variety of physical and chemical forms, of which some can, if released into the water column in sufficient concentrations, cause adverse effects on aquatic organisms and other beneficial uses of the waters. The composition of interstitial waters in sediment is the most sensitive indicator of the types and the extent of reactions that take place between pollutant-loaded sediment particles and the aqueous phase that contacts them. Metals are released from the sediments into the water phase when conditions like the pH change in the water, having a direct bearing on the solubility and/or precipitation of the metals in a freshwater aquatic environment (Wood, 1974; Campbell & Stokes, 1985). The remobilization processes in which natural waters provide the main pathways, can reintroduce these metals into the ecosystem in a bio-available form. Near urban and industrial areas, for example, where substantial amounts of dredging takes place, waterway sediments typically contain large amounts of chemicals which would be of potential concern if they were released from the sediment or if the contaminated sediments were ingested by organisms. The management of sediments is therefore critical to the successful and cost-effective remediation of polluted rivers, because they may also serve as sources of contaminants to aquatic organisms.

Because of the point, as well as possible non-point sources identified in the Upper Olifants River Catchment, it was necessary to determine the effects that these activities have on the water and sediment quality of the Upper Olifants River. In this section of the study, the physical and chemical characteristics of the water were investigated, as well as the metal concentrations (Mn, Cu, Pb, Cr, Zn, Fe, Ni & Al) in the water and different particle sizes of the sediment.

3-3 Water and Sediment

3.2 MATERIALS AND METHODS

Water and sediment were sampled every three months from February 1994 to May 1995 at fourteen sampling sites (Fig 2-10) in the Olifants and Klein Olifants River. Only two of the fourteen sites will be referred to in this study concerning metal accumulation in the water and sediments, named KOR1 (locality 13 in Klein Olifants River) and OR1 (locality 14 in the Olifants River), as fish were also collected at these two localities.

3.2.1 WATER Two surface water samples were collected at each sampling site, using pre-washed polythene bottles, well-rinsed with river water, approximately 10 cm below the surface. The one sample was frozen until the metal concentration could be analyzed in the laboratory. The other sample was preserved with mercuric chloride (HgC1 2) and was refrigerated until further analysis by the Institute for Water Quality Studies, Department of Water Affairs and Forestry. These analyses included the following: pH ♦ Ammonia (NH4-N) (mg/I) ♦ Nitrate and nitrite (NO3- +NO2 -N) (mg/I) ♦ Fluoride (F) (mg/I) ♦ Total alkalinity (CaCO3) (mel) ♦ Sodium (Na) (mg/1) ♦ Magnesium (Mg) (mei) ♦ Silicon (Si) (mg/1) ♦ Phosphates (P043--P) (Inel) ♦ Sulphates (S042-) (mg/1) ♦ Chloride (C1-) (mg/1)

3-4 Water and Sediment

♦ Potassium (K) (mg/l) ♦ Calcium (Ca) (mg/1) ♦ Conductivity (EC) (mS/m) ♦ Total dissolved salts (TDS) (mg/1)

The frozen water samples were thawed in the laboratory, after which 50 ml of well-mixed river water was measured into a 100 ml Erlenmeyer flask. Nitric acid (55 %) and perchloric acid (70 %) were added to the water in a 2:1 ratio and the mixture was concentrated to 25 ml on a hot plate (Standard methods, 1992). Each sample was then made up to 50 ml with distilled water and stored in clean glass bottles. The total metal concentrations in the river water, were determined by means of a Varian atomic absorption spectrophotometer (Spectra AA-10), for which the analytical standards for Mn, Cu, Pb, Cr, Zn, Fe, Ni and Al were prepared from Holpro stock solutions. Potassium chloride (Ka, 200g/1 distilled water) was added to the 50 ml samples, for analysis of aluminium, in order to suppress ionization of aluminium (Varian, 1989). The metal concentrations in the water were determined as follows:

Metal concentration(mg/1) = AAS reading (mg/1)

3.2.2 SEDIMENT One sediment sample (top 5 cm), was taken at each of the fourteen sampling sites with a stainless steel core sampler, fitted with a perspex liner. The core samples were sliced into five centimetre sections, after which the samples were frozen until further analysis in the laboratory. The samples were thawed in the laboratory and dried in an oven at 50 °C for 4 days until dry. This drying temperature ensures that no chemical or

3-5 Water and Sediment

textural characteristics of clay particles which may occur in the sediment, will be altered. After cooling, a 70 g subsample of each sample was placed in an Endecott-mechanical siever, with a sieve rack consisting of sieves at 0.5 0 (phi) intervals, with phi being a factor of the grain sizes in micrometer on a logarithmic basis. The following mesh sizes were used for the purpose of this study: 3 phi (fine sand) 4 phi (course silt) <4.75 phi (fine silt and fine clay)

All the glassware used, was prewashed by being soaked in a 2% Contrad soap solution (Merck chemicals) for 24 hours, rinsed in distilled water, washed in acid (1M HC1) for 24 hours and rinsed again in distilled water (Giesy & Wiener, 1977).

One gram of each of the dried samples was weighed into a 50 ml Erlenmeyer flask, after which concentrated nitric acid (55 %) and concentrated perchloric acid (70 %), in a 2:1 ratio were added. Digestion was performed on a hot plate (±200 °C) for at least six hours, until clear. After cooling, each solution was filtered using an acid resistant 0.45 Am paper filter under vacuum. The filter system was rinsed with distilled water after each filtration and each sample was made up to 50 ml with distilled water and stored in clean glass bottles until the further analysis of the metal concentration.

The metal analysis for the sediment was done in the same way as the water metal analysis. The accurate and precise determination of trace elements in freshwater biological sediment samples is an important aspect of freshwater pollution studies, especially regarding studies of human exposure to toxic elements through freshwater fish consumption. Therefore, for the purpose of this study, a standard sediment sample was used to ensure accuracy (IAEA/R1/64). Care was taken to avoid trace metal contamination from laboratory ware and a triplicate acid blank was also analyzed. The

3-6 Water and Sediment

metal concentration in each sediment sample was calculated as follows:

Metal concentration (p,g/g) = AAS reading (ug/ml) x Sample volume (ml) Sample mass (g)

3.23 DATA PROCESSING Data was processed on an IBM compatible computer utilising the STATISTICA for Windows programme. Differences in mean values were accepted as being statistically significant if P:50.05. Simple linear correlations were also performed between the water and sediment samples, where the linear relationship between two variables is considered, but neither is assumed to be functionally dependant upon the other.

3.3 RESULTS

33.1 WATER The river water showed to be alkaline for both localities (Table 3-1 Appendix, Fig 3-1), with pH values fluctuating between a low of 7.6 at locality KOR1 in February 1995 and a high of 9.1 at locality OR1 in May 1994 (Table 3-1). The water temperature ranged between 8.6 °C in the winter to 27.3 °C in summer (Table 3-1 Appendix, Fig 3-1). The river water seemed well oxygenated at the two localities, with values between 5.4 mg/1 and 12.5 mg/1 and oxygen saturation between 70 and 146 % (Table 3-1 Appendix, Fig 3-2). The conductivity ranged between 35.4 mS/m and 76 mS/m with the highest value

3-7 Water and Sediment

1 0

= = G 25

I a.

• 10

1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 '7 o/ o/ -7 Loca ity KOR1 Locality OR1 Loca ity KOR1 Locality OR1

FIG 3-1 THE PH AND TEMPERATURE (°C) OF THE WATER AT LOCALITIES KOR1 AND OR1 (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-8 Water and Sediment

O O 0 O Nr)

0\ 4) A O 00 0

24. O 1 0 O 0 6 - O 0 0 0 O 01 0 O O O N-1 e-o VI

0. A ton O 00 0 0 0 00 00 00 00 0 00 01. 5. 9. 4. 00 00

00 O 00 O O 0 6. 5

N O 56. O O 0. 31. 8 2.

37.

O 20

4 11 11 O O 4 csi 11. 0 -

1 28. 0 0 - 00 - 00 - 00 - 00 - 00 - 00 - cr co O 00 -

00 - O 00 - 00 - O O 5. O 5. 1 247. 23. 11. 59. 7.

6 24.0 3. O 6. 86. CO 9. O O O O 4 0 2 0 40 20 1 80

60 O 06.

CO 0 00, 1 4.

O O 61. O 2.

00 9. 0. 0,0 O 14. 1

1 00 vl 1 oR 1 "c?

0 0 +1 0± 5 +I +I +I N +I +I 80± +I 90±

80± 20± O 30±

70± 141 O oo O 6. 8 O O 93. 3 ri 14 41 54. 21. 00 86. 00 00 97. O O O O 00 00 00 00 00 00 8 00, 8 17. 00 00 00 O O 76. 09.00 46. N

N 5 1 76. 51. 4.

28. 8 8 0 0. 3. O ‘r

1 1 1 O O 5

cc, 1 00 - 0 - 00 - 00 - 0 - 00 - 00 - 8 00 - 8 0 0 - 5. 00 - 4.

O 4. 8 O "c? 9. 251. 3 2 1 81.0 5. 2. 3 0 70. N 5. O 0 0 00 0 0 3 2 70 60

80 50 O 04. 60

O 1

er 4.60 O 0 0 31. 2. 3. 4. 0. 31. 1 00 O 1 1 2 4. (.1 2

O 0±

O 0 0±

+1 0± 7 0± 8

0± + +I O +I 8 41 +1 0± I O 8 20± O r-- 80± 2 50± O V) 0 ON 8 30. O 0. N 07. 1 51. 363. 23.

2 00 1 33. 22. N 00 co 793

t; 0:11111'. .;"" VI

3-9 Water and Sediment

at locality OR1 in November 1994 (Table 3-1 Appendix, Fig 3-2). The lowest turbidity (2 NTU) was found at locality KOR1 in May 1994 and the highest (50 NTU) during May 1995 at both localities (Table 3-1 Appendix, Fig 3-3). The NO 3- + NO2 concentration at locality KOR1, ranging from 0.04 - 4.95 mg/1, was low in February 1994, increasing rapidly, with the highest value found in November 1994 and decreasing afterwards. At locality OR1, the NO 3- + NO2- concentration ranged between 0.04 - 3.49 mg/1 and was the highest during May 1994 with very low values in November 1994 and February and May 1995 (Table 3-1 Appendix, Fig 3-3). The sodium and chloride concentrations were the highest during November 1994, with values decreasing in February and May 1995 at locality KOR1. At locality OR1, the sodium and chloride concentrations were low in February 1994, with values increasing during May 1994, decreasing in August 1994 and rising again in November 1994. The highest values were however obtained in February 1995 , with much lower values found during May 1995 (Table 3-1 Appendix, Fig 3-4).

The magnesium and sulphate concentrations, ranging between 11 and 31 mg/I for magnesium and 59 and 205 mg/1 for sulphate, were the highest during November 1994 at both localities and decreased rapidly during February and May 1995 (Table 3-1 Appendix, Fig 3-5). The highest potassium and P0 43--P concentrations were found at locality KOR1 in November 1994 and at locality OR1 in February 1994, with very low PO4 values during February and May 1994 at locality KOR1 and during August and November 1994 at locality OR1 (Table 3-1 Appendix, Fig 3-6). The NH 4-s- concentration was mostly in the same range at both localities, fluctuating between 0.04 and 0.06 mg/I, except during August 1994, when a higher concentration of 0.19 mg/I was found at locality KOR1 (Table 3-1 Appendix, Fig 3-7). The silica concentrations showed high variation, with values ranging between 0.4 and 4.6 mg/l. The highest silica concentrations were found in November 1994 and May 1995 at locality KOR1, while very high concentrations were found in February 1994 at locality OR1, with much lower, constant values at this locality during August and November 1994 and February 1995

3-10 Water and Sediment

C) E E C 0) 3 40 C) ' 6 7 0 C 0

1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6

0 / 0 / Loca ity KOR1 Locality OR1 Loca ity KOR1 Locality OR1

FIG 3-2 THE DISSOLVED OXYGEN CONCENTRATIONS (mg/1) AND CONDUCTIVITY (mS/m) OF THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-11

Water and Sediment

ioN 30 3 0 z

I C'J

20 0 1 23 4 5 6 Z 2

10 1 2 3 4 5 6 12 3 4 5 6 1 2 3 4 5 6

o Locality KOR1 Locality OR1 Locality KOR1 Locality OR1

FIG 3-3 THE TURBIDITY (NTU) AND NITRITE AND NITRATE CONCENTRATIONS (mg/l) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACII BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 =MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-12 Water and Sediment

60 50

50 40

30 E

20 20

10 10

1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6

Loca ity KOR1 Locality OR1 Loca ity KOR1 Locality OR1

FIG 3-4 THE SODIUM AND CHLORIDE CONCENTRATIONS (mg/1) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-13 Water and Sediment

5 250

30 7 200

7 25

...... t; 150 -..z.- = cr) 20 E

100 7 /

1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 / Loca ity KOR1 Locality OR1 Locality KOR1 Locality OR1

FIG 3-5 THE MAGNESIUM AND SULPHATE CONCENTRATIONS (mg/1) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-14

Water and Sediment

16 3.5

12 F 2.5

10 aA is A 8 0

1.5

Tn 1 2 3 4 5 6

0.5 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6

Loca ity KOR1 Locality OR1 Locality KOR1 Locality OR1

FIG 3-6 THE POTASSIUM AND PHOSPHATE CONCENTRATIONS (mg/I) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-15 Water and Sediment

0.2 5

4 0.15

r /

bA 3 C Ts) 0.1 E z 2

0.05

.=:E= 1 2 3 4 6 1 2 3 4 1 2 3 4 5 6 1 2 3 4 5 6 0 Locality KOR1 Locality OR1 Locality KOR1 Locality OR1

FIG 3-7 THE AMMONIUM AND SILICA CONCENTRATIONS (mg/1) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-16 Water and Sediment

(Table 3-1 Appendix, Fig 3-7). The total dissolved salts and calcium concentrations were the highest during November 1994 and decreasing in February and May 1995 at locality KOR1. At locality OR1, the TDS concentrations ranged between 247 and 501 mg/1, with the lowest values recorded in February 1994 and May 1995 and highest in November 1994 and February 1995, while the calcium concentrations, ranging between 23 and 56 mg/1, showed the highest values in November 1994 (Table 3-1 Appendix, Fig 3-8).

The highest values for CaCO3 and fluoride at locality KOR1, were found during February 1995 (Table 3-1 Appendix, Fig 3-9). The fluoride concentrations were constant at locality KOR1 at 0.06 mg/1, except in November 1994, when a value of 0.07 mg/1 was recorded. The fluoride concentrations were also stable during February and May 1994 at locality OR1, but higher concentrations occurred in November 1994. The CaCO3 concentrations, ranging between 39 and 109 mg/1 at locality KOR1 and 65 and 119 mg/1 at locality OR1, showed the highest values in February 1995.

The metal concentrations for the river water (Fig 3-10) are summarized in Table 3-2 and the mean values compared to guideline values set by Kempster, Hattingh & van Vliet (1982), Kiihn (1991) and Canadian guidelines (1987), for the protection of aquatic life (Table 3-3). At locality KOR1, the mean zinc and aluminium concentrations were much higher than at locality OR1, while the mean copper, chromium and iron concentrations were higher at locality OR1 and the lead, manganese and nickel concentrations were in the same range at both localities. In the first year, at locality KOR1, the copper and zinc concentrations were higher than in the second year, while the lead and iron concentrations were higher in the second year. At locality OR1, the lead and aluminium concentrations were higher in the first year, while the manganese, chromium, iron and nickel concentrations were higher in the second year.

3-17 Water and Sediment

2=7 ,0

..—. C) .=" G E 400 co el N C) 13 a) 300 E > 0 C N 0 N _ Ts. 200 20 0 1—

100 0 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 1 2 3 4 5 6 -7 Loca ity KOR1 Locality OR1 Loca ity KOR1 Locality OR1

FIG 3-8 THE TOTAL DISSOLVED SALTS AND CALCIUM CONCENTRATIONS (mg/1) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 =MAY 1995)

3-18 Water and Sediment

140 / 0.8 /

0.7 120

4=== 0.6 100

E 0.5 O 80 E G) 0.4

60 0 LL 0.3

40 0.2

20 0.1 1 2 3 4 5 6 1 2 3 4 5 6 '7 1 2 3 4 5 6 1 2 3 4 5 6 0/ Locality KOR1 Locality OR1 Loca ity KOR1 Locality OR1

FIG 3-9 THE ALKALINITY AND FLUORIDE CONCENTRATIONS (mg/1) IN THE WATER AT LOCALITIES KOR1 AND OR1. (EACH BAR REPRESENTS A MONTH OF SURVEY - 1 = FEBRUARY 1994, 2 = MAY 1994, 3 = AUGUST 1994, 4 = NOVEMBER 1994, 5 = FEBRUARY 1995 AND 6 = MAY 1995)

3-19

Water and Sediment

2.5

0.8 1.1

1.5

1.2

C O Ti

4E I a) C UO nl 0.6 a) 0.4

0.5 0.3 0.5 0.1 /-7 0.08 0.06 0.3 0.06 0.1 0.2 0.01 0.1 0.05

/--7 0 ; / / / Mn Cu Pb Cr Zn Fe Ni Al Mn Cu Pb Cr Zn Fe Ni Al

FIG 3-10 THE MEAN METAL CONCENTRATIONS IN THE WATER (mg/1) AS FOUND DURING THE PERIOD FEBRUARY 1994 - MAY 1995 AT LOCALITIES KOR1 (A) AND OR1 (B). THE STANDARD DEVIATION FOR EACH METAL, IS PRESENTED ABOVE EACH BAR

3-20 Water and Sediment

TABLE 3-2 THE METAL CONCENTRATIONS IN THE WATER OF THE OLIFANTS RIVER AT LOCALITIES KOR1 AND OR1, DURING THE PERIOD FEBRUARY 1994 - MAY 1995

KOR1 Feb 94 0.02 0.03 0.13 0.27 0.19 0.99 0.21 0.39 May 94 0.02 0.70 0.13 0.26 0.16 1.08 0.19 3.0 Aug 94 0.02 0.01 0.23 0.29 0.05 1.27 0.10 0.35 Nov 94 0.01 0.01 0.24 0.21 0.15 1.35 0.03 0.72 . Feb 95 0.06 0.03 0.24 0.18 0.12 1.40 0.23 1.14

May 95 0.11 0.19 0.21 . 0.29 0.14 3.21 0.23 3.73 OR1 Feb 94 0.02 0.03 0.20 0.33 0.05 1.66 0.22 0.12 May 94 0.02 0.04 0.16 0.3 0.05 1.73 0.20 0.38 Aug 94 0.02 0.01 0.21 0.28 0.04 1.28 0.10 0.37 Nov 94 0.01 0.01 0.26 0.233 0.10 1.14 0.04 0.75 Feb 95 0.02 0.02 0.02 1.62 0.07 3.33 0.37 0.21 May 95 0.31 0.02 0.11 1.02 0.06 3.51 0.25 0.12

3-21 •• ▪

Water and Sediment RED

0 COMPA

r 't -•

,4 0 0 1 rl3

00 3

> 3

0 CO /1 CO CO 1 Ca

ag / Ca Ca mg On '4, 8 /I 0 /I /I 4.

g 0 mg <

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5:[

z t 6. 5:1 dness dne tec dness 6. ro u li< p .4 p Har = <4 a Har Har I: I - 1: pH> /I: / / cn kj•• / /1 to /I - Ls. 8 1 mg 001 mg 002 mg 025 mg 005 mg 2 a 02 mg 0. 0. 0. 0. o 0. 0. z E z cr.1 0 n 0 v-) ,•-•• z q z 0

a, /I o„, /1 a

z 2 mg 05 mg 1 mg

0 0. 0. f-* ,=4 0. - 6 o 005 - 02 - Z07- 025 0. 0. 0. A A A A 8r" 4,

o 0 47

CV 81

,-• c7.1 9458 4 •

L,a coe„, 0 r.s • 5000 - 35820 - 0 co 2 2 7

z s o z 0734 1 P 00 24886 00 fi2 +1 +1 +1 z 0 ON ON 646± e•-• 512± 53 z 39 0 o ,1•11.11- .... 0 a., 00

00 54

4.1 (+) 9 5822 CV CN

g • • 0 - 00 00 - Z q 00 .—■ 2204 298 CtS' +1 U 5 75

Z ••• CO 3 1 CV <1. 24241 d Z co 0 +1 +I +1 +I 84± 5 .7

On•

S.--F•11

3 - 2 2 Water and Sediment

3.3.2 SEDIMENT The percentage of the different sediment fractions in the top 5 cm of the samples of both localities consists mainly of fine sand, followed by smaller percentages of course silt and the least fine silt and fine clay (Fig 3-11). Metal concentrations in the different grain- size fractions, showed high variation. Comparison of the concentrations of the eight selected metals in the three different particle sizes, showed that for all the metals, the highest concentrations occurred predominantly in the fine silt and fine clay particles (Table 3-2 Appendix and Table 3-3 Appendix).

The highest manganese concentrations were found during May and November 1994 at locality KOR1 in the fine silt and fine clay particles and also in February 1994 at locality OR1, with high concentrations in the other particles during this time as well (Fig 3-12). The copper concentrations were mostly in the same range at both localities, except for an extremely high value found in November 1994 at locality KOR1. At locality OR1, the copper concentrations in the fine silt and fine clay particles were very high during February, May and November 1994, with the highest concentration in the fine sand particles in November 1994 (Fig 3-13). At locality KOR1, the lead concentrations were higher in the fine silt and fine clay particle, while at locality OR1, the coarse silt particles showed higher concentrations during February and August 1994 (Fig 3-14). The chromium concentrations were mostly in the same range at both localities, except for higher concentrations in the fine silt and fine clay particles at locality KOR1 during February 1994 and in the fine sand particles in February 1995 and much higher concentrations in the fine silt and fine clay particles during November 1994 at locality OR1 (Fig 3-15). High concentrations of zinc were found during February 1994 at locality KOR1, while the concentrations at locality OR1 increased from February - May 1994, with a lower value in the fine silt and fine clay particles in August 1994, but higher values in the coarse silt particles at that time, after which the highest value was found in

3-23 FIG 3-11 / TO MAY1995 SAMPLES COLLECTEDATLOCALITIESKOR1 AND OR1FROMFEBRUARY1994 PERCENTAGE OFDIFFERENTSEDIMENTFRACTIONS (TOP5CM)OFTHE

Fine sandEiCoursesilt Persentag e g rain size (%) 100 120 40 20 60 80 0 Locality KOR1 ❑ Fine silt&clay Locality OR1 Water andSediment / / 3-24

Water and Sediment

2,500 / 2,500

2,000 2,000

01 p.2 CD co S 3 c 1,500 c 1,500 0 0 V_ .7711 t" z a) ad U c o 1,000 . 1,000 U U C C M 2

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

Fine sand Ei Course silt ❑ Fine silt & Fine clay /

FIG 3-12 MEAN MANGANESE CONCENTRATIONS (µg/g) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-25 Water and Sediment

700

250

...... 7:i 0 / 0 0) 200 a a c 0 400 as E a) c.) c 0 0 U o = = 100 ° 200 0

100

0 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

N Fine sand ❑ Course silt ❑ Fine silt & Fine clay /

FIG 3-13 MEAN COPPER CONCENTRATIONS (Ag/g) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-26 FIG 3-14 Pb concentration (pg/g) 100 20 40 60 Feb 94 May 94Aug THE SEDIMENTCOLLECTEDATLOCALITIES KOR1(A)ANDOR1(B)FROM FEBRUARY 1994TOMAY1995 MEAN LEADCONCENTRATIONS r4 Fine sand Nov 94 Feb 95 El Course silt May 95 (tig/g) ❑ "t3 a 0- -0 20 CM CI) C 0 C 0 Fine silt&clay 40 50 3 IN 10 0 THE DIFFERENTPARTICLESIZESOF Feb 94 ■ ► 0 10 .0 .0 .0 0 0 0 • 0 0 • 0

May 94 0; Oo $1 $1 4

Aug 94 Water andSediment hO NO AO NO AO ♦ ■ &I P40 NO NO NO NO

• • Nov 94 Feb 95 No No 3-27 May 95 .o Water and Sediment

350

400

0) -3-1 ) 250 0) 0) 3 3 c 300 0 0 200

U 0 150 Oc 200 O U_ U 100

100

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 Mey 95

Fine sand D Course silt ❑ Fine silt & Fine clay

FIG 3-15 MEAN CHROMIUM CONCENTRATIONS (grig) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-28 Water and Sediment

November 1994 in the fine silt and fine clay particles (Fig 3-16).

The iron concentrations were very high with values found during May 1994 in the fine silt and fine clay particles being the highest at locality KOR1, while the coarse silt particles showed the highest concentrations during this period at locality OR1 (Fig 3-17). The highest nickel concentrations were found in the fine silt and fine clay particles during May 1994 at locality KOR1 and in the fine sand particles during February 1995. At locality OR1, the highest concentrations were found in the fine silt and fine clay particles during November 1994 (Fig 3-18). Aluminium concentrations reached a high of 95 400 pg/g dry mass during May 1994 in the fine silt and fine clay particles at locality KOR1, while the concentrations in these particles were the highest at locality OR1 during February and November 1994 (Fig 3-19).

The correlations between water and sediment metal concentrations, were predominantly negative and not statistically significant (P>0.05) (Table 3-4). The highest negative correlation was obtained for chromium at locality OR1 followed by zinc at locality KOR1, manganese and iron at locality OR1, chromium and manganese at locality KOR1, copper at locality OR1, iron, aluminium and nickel at locality KOR1 and nickel and lead at locality OR1. A few positive correlations were found with the highest for zinc at locality OR1, followed by lead and copper at locality KOR1 and aluminium at locality OR1. The inverse relationship between the metal concentration in the water and sediment, was only significant (P5_0.05) for chromium and zinc at locality OR1.

3-29 •

Water and Sediment

400 700

350 600

300 00 0) 5 0) 250 C 0 400 tC Tt3 4E. 200 O 0 C 300 0 0 0 150 C C N N 2 00 100

100 50

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

Fine sand El Course silt ❑ Fine silt & Fine clay

FIG 3-16 MEAN ZINC CONCENTRATIONS (ug/g) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-30 Water and Sediment

200

150 rn .53 100 0) 0)

C C 0 0 as ca 100 C C" a) a) U 0 C C O 0 2 z

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

Fine sand 13 Course silt ❑ Fine silt & Fine clay

FIG 3-17 MEAN NICKEL CONCENTRATIONS (Agjg) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-31

Water and Sediment

100 70

ds) 80 't3 C

50 0 Thousan ( ) 0 CS) 40 (pg/g C ion 0 t -413. 30 40 tra

+6) C.) ncen 20 C.) co 0 a) Fe U_ 10

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

■ Fine sand El Course silt ❑ Fine silt & Fine clay

FIG 3-18 MEAN IRON CONCENTRATIONS (p,g/g) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-32 Water and Sediment

100 /

--- 50 cn 0 ds) 0 C Sra 0 .c 40

Thousan I- (

) 60 0) /g 01 S 30 (Ng

n C 0 io t 40 tii tra t 0) 20 0 C oncen 0 0 c 0 Al 471C- 10

Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

Fine sand Efl Course silt ❑ Fine silt & Fine clay

FIG 3 -19 MEAN ALUMINIUM CONCENTRATIONS (i.tg/g) IN THE DIFFERENT PARTICLE SIZES OF THE SEDIMENT COLLECTED AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

3-33 Water and Sediment

E-■

C:4 z z

C.4

3-34 Water and Sediment

3.4 DISCUSSION

3.4.1 PHYSICAL AND CHEMICAL CHARACTERISTICS OF THE RIVER WATER Increasing acidity in a river or stream, shown by decreasing pH levels (i.e. acidification), can occur as a result of acid input to water from various sources, including the atmosphere (acidic precipitation or "acid rain"), thermal power stations, vehicles, industrial effluents or internal (microbial and chemical)/natural processes or acid mine drainage (Ellis, 1989). Coal mines, which are commonly found in the study area, are the single most important cause of acid mine drainage, although the mining of other ores also causes local water quality problems. This process of acidification affects major water sources and has a great impact on the survival, growth and physiological aspects of all the fauna and flora residing in these affected water environments, as aqueous pH can greatly influence the bioavailability and toxicity of cationic metals to fish. The pH at localities KOR1 and OR1 was alkaline, but much lower pH values (as low as 3.6 at locality 10) were indicated in this study and should there be a period of low water flow, there could be a significant decrease in the pH of the river water.

The pH is a measure of the hydrogen ion activity in a water sample, according to the expression: pH = -logio[H1, where [H+] is the hydrogen ion activity. In most instances, however, the change in pH in natural waters, is associated with other changes in water quality. For example, a decline in pH, may be accompanied by an increase in the metal concentrations, especially that of aluminium, which is toxic to most organisms in excessive concentrations and which is found in very high concentrations at localities KOR1 and OR1. The effect of pH change on the remobilisation of aluminium has been associated with fish kills in the Olifants River Catchment (Dr Kempster, personal communication). Rapid increases in pH can also cause increased NH 3 concentrations which are also toxic to aquatic life. Ammonia has been shown to be ten times as toxic

3-35 Water and Sediment

at pH 8.0 as at pH 7.0 (EIFAC, 1969). This should be taken into consideration, as the pH at locality OR1 in May 1994, was as high as 9.1. However, a pH more alkaline than 9 for long periods, can diminish reproduction and growth of the fish (Mount, 1963). The pH for localities KOR1 and OR1, was generally within the recommended range of guideline values for aquatic ecosystems (Table 3-1).

The property of alkalinity is usually imparted by the presence of bicarbonates, carbonates and hydroxides and sometimes in inland waters by borate, silicate and phosphates. The CO2-HCO3--0O3= equilibrium system is the major buffering mechanism in fresh waters (Wetzel, 1983). Alkalinity is important for fish and other aquatic life in freshwater systems, as it buffers pH changes that occur naturally as a result of photosynthetic activity of the chlorophyll-bearing vegetation (Train, 1979). The buffering capacity of the study area seemed to be fairly moderate, as the alkalinity ranged between 39 mg/1 and 119 mg/I, compared to alkalinity of 140 - 235 mg/1 CaCO 3 found in the lower Olifants River (Seymore, du Preez, van Vuren, Deacon & Sfrydom, 1994). Components of alkalinity, such as carbonate and bicarbonate may also complex some toxic metals and thereby reduce their toxicity (Train, 1979).

All organisms have an optimum temperature for normal life functions. By changing the natural water temperature, the aquatic biota is exposed to potentially lethal or sublethal temperatures. Temperature changes can affect the dissolved oxygen in the water, as higher temperatures reduce the solubility of dissolved oxygen in water, decreasing its concentration and therefore also its availability to aquatic life. This has been observed at localities KOR1 and OR1, with higher temperatures associated with lower dissolved oxygen and lower temperature with higher dissolved oxygen in the river water. With the high range of temperature at localities KOR1 and OR1 during the period February 1994 to May 1995 this is important to take into consideration, as the body temperature of most aquatic organisms is almost equal to that of the ambient water and their temperature tolerance is low. Therefore, the direct effect of temperature increase or

3-36 Water and Sediment

decrease can be detrimental to these organisms, as an increase of 7 °C is adequate to double the chemical reactions in plant and animal cells. A change in water temperature can influence the metabolic activity and behaviour of the organism which may affect their exposure to a pollutant and it may also alter the physical and chemical state of the pollutant (Cairns, Heath & Parker, 1975), as for example, in the case of metals, where an increase in temperature causes an increase in toxicity of these metals (Felts & Heath, 1984). In this study, the temperature changes are an important factor, especially with the release of cold bottom water from, for example the Witbank Dam, as well as the low flow of the water during drought periods, which can have a negative effect on the aquatic organisms. It is also important to note that dams of moderate depth (> 10 m) can exhibit thermal stratification. With respect to this, metals like aluminium increase slightly with depth during the thermal stratification period, with higher ranges during the mixing period. The dynamics of iron also evolves from low concentrations of total iron in the entire water column during the thermal mixing period, when the water is well oxygenated, to higher levels in deep, poorly oxygenated water (especially Fe') produced through electrochemical of biological reduction of Fe' precipitates present in sediments (Galvin, 1996). The high summer water temperatures found at the two localities, together with the high pH values and high nutrient load from the sewage works in the area of the two localities, also favour the growth of aquatic weeds and algae.

For sufficient aerobic conditions to occur, water should contain sufficient dissolved oxygen. Dissolved oxygen in the water prevents the chemical reduction and subsequent leaching of iron and manganese, mainly from sediments (EPA, 1973). The dissolved oxygen concentration at both localities was however within the Canadian guideline limits (>5 mg/1), but should the dissolved oxygen levels decrease rapidly, it can adversely affect aquatic insects and other animals upon which the fish feed (Train, 1979). It is, however, important to take into consideration that fish vary in their oxygen requirements according to species, age, activity, temperature and nutritional state. Fish are found to be able to survive for a period of time at oxygen concentrations considerably below that

3-37 Water and Sediment

which are considered suitable for thriving populations (Train, 1979). Even under optimum conditions, it is rare to find dissolved oxygen concentrations of more than 8 - 10 mg/I (Ellis, 1989), although dissolved oxygen concentrations of 11.5 mg/I and 12.5 mg/1 were observed in August 1994 at both localities. This may be due to photosynthetic oxygen, produced under the influence of sunlight by aquatic plants and algae, as excessive growth of these plants was observed at that time.

Turbidity can be caused by suspended matter such as silt, clay, finely divided organic and inorganic matter, plankton and microscopic organisms, as well as the absorption of soluble coloured compounds (Day, 1990). Factors such as rainfall can also have an affect on the turbidity as the flow of the river changes and more of the sediment particles are found in the river water. An increase in turbidity can have the following effects of the aquatic environment: Reduced light penetration, which causes a reduction in photosynthesis, resulting in reduced plant biomass and available food Reduced visibility of pelagic food Reduced availability of benthic food Temperature decrease, which can affect temperature sensitive species Physiological impairment, as suspended material can cause damage to fish gills, leading to impaired respiration.

The results obtained from the study showed high values of turbidity at locality KOR1 in November 1994, when rainfall occurred and February and May 1995, which can be ascribed to excessive sediments, possibly from the informal settlements and dirt roads in the area. At locality OR1 the turbidity increased in August 1994 with the highest value in May 1995, when excessive growth of aquatic plants and algae was also observed.

Ammonia may enter surface and ground water through discharge of industrial wastes

3-38 Water and Sediment

containing ammonia as a by product, or through wastes from industrial processes using "ammonia water" (Train, 1979). High levels of ammonia may also enter the water as fertilizer components, or most importantly, through effluents from industries and sewage works, which may be the case at locality KOR1. Ammonia is toxic to fish and this toxicity varies with the pH of the water (Boyd, 1982). At lower pH values, the less toxic ammonium ion (NH4+) exists in the water, while the more toxic ammonia (NH 3) is present in more alkaline conditions, as is the case at localities KOR1 and OR1. The pH range in most natural waters is such that the NH4+ fraction of ammonia predominates, although the NH3 fraction can reach toxic levels in highly alkaline waters. Even a small increase in pH (together with an increase in temperature), for example from pH 7 - 8, will increase the toxicity of ammonia tenfold. The highest ammonia levels were found at locality KOR1, which is situated downstream of a combined sewage works, where industrial and urban runoff are routed through the sewage works into the river. It also receives effluent from informal settlements in the area, but the NH 4+ values obtained for this study are still within the guideline limits (Kempster et al., 1982).

Ammonium is oxidized to nitrite (NO 2-) by bacteria known as Nitrosomas, found in water, soil, sewage and the digestive tract. Nitrite is in turn rapidly oxidized to nitrate (NO3-) in oxygenated natural waters by bacteria known as Nitrobacter. Nitrite and nitrate are two forms of total oxidized nitrogen (TON). Ellis (1989) suggested that the TON concentrations in drinking waters should not be more than 11 mg/I (as nitrogen). The major point sources of nitrogen entry into water are industrial waste waters, septic tanks and municipal waste waters. Diffuse sources include fertilizer, animal wastes, atmospheric fallout, nitric oxide and nitrite discharges from automobile exhaust and other combustion processes and losses from natural sources such as mineralization of soil organic matter (NAS, 1972). The TON concentrations at locality KOR1 were high in August and November 1994 and February 1995, which may be ascribed to the combined sewage purification works and informal settlements in the area. At locality OR1, high values were obtained in May 1994, but the TON levels are still within the guideline limits

3-39 Water and Sediment

(as nitrogen) and coincide with the high NH 4-N concentrations in August 1994. Denitrification on the other hand, takes place when nitrate-containing soils become anaerobic and the conversion to nitrite, molecular nitrogen, or nitrous oxide occurs. Ammonium ions may also be produced in some circumstances.

Phosphorus, which usually gives rise to excessive algal growth, is normally the limiting element, because the amount of biologically available phosphorus is usually small in relation with the required amount. An increase in phosphorus, as detected at locality KOR1 from August 1994 - May 1995, will therefore result in an increase in productivity, also observed at that time. The values obtained during this study for phosphorus, are much higher than the guideline values (Kempster et al., 1982) as higher concentrations of phosphorus are likely to occur in waters receiving treated sewage (localities KOR1 and OR1) and leaching or runoff from cultivated land (Dallas & Day, 1993). The lower flow of the water at locality KOR1 in the summer, could also influence the concentrations, as dilution is minimal.

Among trace elements, fluoride plays a major role in deteriorating aquatic ecosystems. Fluoride concentrations are influenced by water hardness, as high calcium concentrations suppress fluoride concentrations through precipitation of insoluble calcium fluoride (Smith, Holsen, Ibay, Block & de Leon, 1985). The fluoride concentrations for localities KOR1 and OR1 ranged between 0.5 mg/1 and 0.7 mg/1 and fall within the target values set by Kempster et al. (1982).

Elemental chlorine is a greenish-yellow gas, that is highly soluble in water and reacts readily with many inorganic substances and all animal and plant tissue. Chlorine in the free available form, reacts with nitrogenous organic material to form chloramines, which are toxic to fish (Train, 1979). Chlorine causes damage to the epithelium of fish gills and subsequent production of mucus and eventual clogging of the gill lamellae (Cairns et al. 1975). Chlorine can be used for disinfection and removing of unwanted tastes and

3-40 Water and Sediment

odours from drinking water and destruction of pathogenic bacteria, but an excess can make water unsuitable for maintaining living aquatic conditions. Chlorination is also used in sewage treatment to reduce odour or the number of bacteria in effluents discharged to surface waters and it is also added to cooling waters and other industrial waste waters (Train, 1979). The chlorine concentrations at localities KOR1 and OR1 fluctuated between 10 mg/I and 42 mg/I, which are well below the guideline limits of 50 - 400 mg/I (Kempster et al., 1982).

Sodium and potassium are ubiquitous in natural waters, and are the major cations involved in ionic, osmotic and water balance. Sodium and potassium also play a role in the transmission of nervous impulses and thereby muscle contraction. According to Hellawell (1986), as mentioned above, sodium is the least toxic metal cation, although fish kills in the Olifants River have previously been associated with high levels of potassium and sodium, with elevated potassium levels thought to have been responsible for these mortalities (Moore, 1990). The values obtained in this study in the Upper Olifants River (16 - 56 mg/I Na; 6.3 - 14.1 mg/I K), with the highest values obtained in November 1994 at locality KOR1 fall, however, within the set guideline values (Kempster et al., 1982) and should therefore not be considered variables of concern.

Magnesium is an essential element, found in chlorophyll and a variety of enzymes involved in the transmission of nerve impulses and muscle contraction. Magnesium compounds are much more soluble than those of calcium which is one of the major elements essential for living organisms. Because of magnesium's solubility characteristics and its minor biotic demand, the concentrations of magnesium are relatively conservative and constant (Wetzel, 1983), an observation also supported in this study, with values ranging between 11 mg/I and 3 mg/I, well below the guideline values of 1 500 mg/I (Kempster et al., 1982).

The predominant form of dissolved sulphur in water, is the sulphate ion (S0 42.). The

3-41 Water and Sediment

sulphate itself is not toxic, but in excess, it forms sulphuric acid, a very strong acid which reduces the pH of the water, having deleterious effects on aquatic life. The major sources of sulphate to surface waters, are surface runoff (S0 42-), precipitation and dry fallout (S042- and SO2), most importantly mines (acid mine drainage) and groundwater (S042-). In the Karoo Sequence Region, as found in the Witbank area, the shale and coal strata contribute to the sulphate load in the water, as they contain pyrites, which, in contact with water and air, oxidize and form copper and iron sulphates. Drainage water from mines causes severe pollution problems, with coal mines being very important. When coal is mined, high levels of the mineral pyrite, a crystal composed of reduced iron and sulphur (FeS2), are often oxidized as a result of exposure to air, water and chemosynthetic bacteria and these utilize the energy obtained from the conversion of the inorganic sulphur to sulphate and sulphuric acid. The S0 42- concentrations at localities KOR1 and OR1 were, however, well within the guideline limits (Kempster et al., 1982), ranging between 59 mg/1 and 205 mg/l.

The major sources of total dissolved salts are firstly, weathering of the rocks over which the water flows or from which it drains. Secondly human activities have increased the TDS concentrations of waters worldwide, particularly in arid regions. Increased levels of TDS (known as salinization of mineralization) may be caused by discharging of saline industrial effluents into rivers, irrigation and return of large quantities of sewage effluent to inland waters. The TDS concentrations at localities KOR1 and OR1 are within the guideline limits (Kuhn, 1991), with values ranging from 247 mg/1 - 517 mg/l. It is, however, suggested that monitoring of the TDS should take place constantly during the year, in order to establish a definite trend. The conductivity (EC) is exerted by the dissolved salts and therefore the following relationship exists between the TDS and EC: EC (mS/m) x 6.5 = TDS (mg/1), depending, however, on the composition of the water (pH and bicarbonate content) (Kempster et al., 1982).

3-42 Water and Sediment

3.4.2 METAL CONCENTRATIONS IN THE WATER AND SEDIMENT Metal toxicity and availability in aquatic ecosystems are influenced by factors such as particulate, dissolved organic and inorganic constituents, which affect both form and bioavailability of these metals (Giesy, Briese & Leversee, 1978) and are dependant on the hardness of the water (Sprague, 1969). For example, copper is most toxic in the divalent cationic form and copper toxicity is inversely related to the water hardness (Brown, Shaw & Shurben, 1974). The water at localities KOR1 and OR1 is moderately hard and therefore would reduce the bioavailability of metals, compared to soft water systems. But, longterm exposure to levels of available forms of these metals may pose a potential threat to aquatic life (Steenkamp, du Preez, Schoonbee & van Eeden, 1994). According to Hellawell (1986), the approximate order of metal toxicity is: Hg> Cu > Cd & Zn > Au? & Sn >Ag?, Al & Ni >Pt?, Fe' & Fe' >Ba>Mn & Co>Li &K>Ca >Sr & Mg>Na.

The lead, chromium, iron, nickel and aluminium concentrations in the water at localities KOR1 and OR1, were much higher than the guideline limits for the protection of aquatic life (Kempster et al., 1982; Environment Canada, 1987; Kuhn, 1991). The lead and nickel concentrations, ranging between 0.02 - 0.26 mg/1 and 0.03 - 0.37 mg/1 respectively, were much lower than values obtained by Du Preez and Steyn, (1992), found in the water of the Olifants River in the Kruger National Park, ranging between 0.23 - 0.44 mg/1 and 0.13 - 2.4 mg/I respectively, which were also much higher than the values obtained by Seymore et al. (1994) in a study also performed in the lower catchment of the Olifants River. The iron concentrations at localities KOR1 and OR1 were also lower than concentrations found in the Elsburg Spruit in Gauteng, receiving mining and industrial effluents (Van der Merwe, Schoonbee & Pretorius, 1990).

The mean copper and zinc concentrations in the water at locality KOR1 (0.16 mg/1 and 0.14 mg/I respectively), located downstream from a combined sewage purification works, were higher than concentrations found in a maturation dam, as well as in the

3-43 Water and Sediment

Krugersdrift Dam (Van den Heever & Frey, 1994), the Vaal River (Rand Water Board, 1982) and Transkei Rivers (Du Preez, 1985). High copper concentrations can be expected as a result of the sewage works and industries alongside the river, which is also subjected to domestic, agricultural and industrial effluents (Steenkamp, Du Preez & Schoonbee, 1994). The copper and zinc concentrations found at this locality are, however, well below the recommended value of 1 mg/1 and 5 mg/1 respectively, for drinking water (USEPA, 1976). With the normal water intake of man, of ± 2 litres per day, it poses no health risk to the consumer. The copper and zinc concentrations at locality OR1 were, however, in the same range than the values obtained in these rivers and dams. The mean chromium concentrations at locality KOR1 (0.25 mg/1) and locality OR1 (0.63 mg/1) were also higher than chromium concentrations found in treated sewage effluent from the Bloemspruit treatment plant (0.012 mg/1) and the Krugersdrift Dam (0.009 mg/1) (Van den Heever & Frey, 1996), with the standard for chromium in the water, set by the USEPA (1976), being 0.05 mg/1. These concentrations were also higher than the concentration found in the Elsburg Spruit (0.15 mg/1), due to electroplating industries in the area (Schoonbee & Van der Merwe, 1989).

The mean manganese concentrations at localities KOR1 and OR1 (0.04 mg/I and 0.07 mg/I respectively) were however lower than concentrations found in the Elsburg Spruit and a tributary, receiving effluent from disused ash dams, foundry waste dumps and mines, ranging between 4 - 8 mg/I (Schoonbee & Van der Merwe, 1989). It should, however, be noted that measurements were taken only once every three months and that an accurate indication of the metal concentrations in the water could therefore not be obtained. More frequent measurements, for example at least weekly over an extended period are therefore suggested. The lead, aluminium, copper and zinc concentrations were, however, higher at locality KOR1, while the manganese, chromium, iron and nickel concentrations were higher at locality OR1, with very high concentrations of iron, aluminium and chromium at both localities. This coincides with the high iron and aluminium concentrations in the sediment.

3-44 Water and Sediment

In this study, the metal concentrations were higher in the sediments than in the water, due to natural background levels and adsorption of metals to the sediment particles, especially the fine silt and fine clay particles. The sediment at localities KOR1 and OR1 can be regarded as heavily polluted, if compared to the elemental composition of average earth sediments, and the iron and aluminium concentrations were extremely high at both localities (Table 3-2). Granite rocks are found in the area between Witbank and Middelburg, and the metal concentrations obtained during this study, were very high in comparison to typical background concentrations in granite (Freedman, 1989) (Table 3-2). The major problem, however, with evaluating these concentrations is that no background levels of the specific catchment is available. Manganese and iron concentrations were much higher in the sediment, at both localities, than concentrations found in the lower catchment of the Olifants River (Seymore et al., 1994). Concentrations found during this study, were much lower than values obtained for manganese and nickel in the Natal Spruit and Bronkhorst Spruit Rivers (Gauteng) and the Nooitgedacht Dam (North West Province), receiving effluents from mine dumps, mine tailing dams, industries and sewage purification works (Steenkamp, Du Preez, Schoonbee & Van Eeden, 1994; Steenkamp, Du Preez, Schoonbee & Shafir, 1995). The chromium and zinc concentrations found at localities KOR1 and OR1 were however much lower than values obtained in the lower catchment of the Olifants River (Seymore et al., 1994). More intensive research is however needed concerning the metal concentrations in the different particle sizes of the sediment, as well as regarding the remobilisation of the different metals and form of metals from the sediments, as aquatic sediments tend to become contaminated with both inorganic and organic chemicals, which are sorbed to particulate matter or in solution in sediment pore water (Knight, 1984). These contaminants can affect other aquatic organisms by becoming bioavailable through resuspension of leaching, as changes in the pH of the water could have a direct bearing on the solubility of the metals as well as the depositing capacity of such metals in the sediments (FOrstner & Prosi, 1979). It is generally accepted that the compositions of top sediment layers reflect the current water quality of the system. In

3-45 Water and Sediment

determining the extent to which the sediment texture can play a role in increasing the sediment element concentrations, it was shown that although inorganic substances can also concentrate on large substrata, fine-grained sediments bind metals more efficiently than do coarse-grained sediments, with the lowest metal concentrations found in the fine sand particles and the highest concentrations in the fine silt and fine clay particles. The concentration of sediment-bound metals, therefore appears to be dependant on the surface area of the sedimentary particles.

It was found that a predominant negative correlation exists between the concentration of metals in the water and the metal concentrations in the sediments. A negative correlation indicates that an increase in the value of one of the variables (water) is accompanied by a decrease in value of the other variable (sediment). The highest negative correlation between metal concentration in the water and in the sediment, was obtained for zinc, which is a common pollutant of surface fresh waters in many industrialised areas, at locality KOR1. However, a significant positive correlation was obtained for zinc at locality OR1. A positive correlation implies that for an increase in the value of one of the variables, the other variable also increases in value. If a zero correlation should be obtained, denoting that there is no linear association between the magnitudes of the two variables, it means that a change in magnitude of one variable, does not imply a change in magnitude of the other. It must be stressed that these comparisons must be treated with caution, as water samples were only collected every three months.

3.5 CONCLUSION

Results obtained on the water chemistry of the river in the upper catchment area during

3-46 Water and Sediment

the present study, pointed towards polluted conditions, existing due to the effects of mining, industrial and sewage pollution in the system. The selected water quality variables showed values mostly within the guideline limits for protection of aquatic life, but the phosphorus concentrations were much higher than the permitted 0.1 mg/I in the water, as a result of effluent discharges from combined sewage purification works and informal settlements in the study area. As the P0 43- concentration is above the guideline value, the 0.1 mg/I standard for effluent is not sufficient. It is therefore suggested that this standard be revised, possibly as a site specific guideline, as, for example at locality KOR1, the river consists mainly of treated sewage effluent in the dry seasons. Excessive aquatic plant growth was also observed at both localities as a result of high nutrient loads in the water from the sewage purification works, especially in the dry, warm seasons.

Mining, industries, informal settlements and combined sewage purification works also contribute to the metal load in the water and sediment. The most concentrated pool of elements was, however, found in the bottom sediments of the water column, with extremely high concentrations of iron and aluminium. The sediment consists mostly of fine sand, with the highest metal concentrations detected in the fine silt and fine clay particles. The bottom sediment of a water column plays an important role in the pollution of river systems and can reflect the water quality of that system. Changes in water quality, such as a decrease in pH, can cause metals to be reintroduced into the water. The pH at localities KOR1 and OR1 were however more alkaline, but other localities in the upper catchment (for example locality 10) showed pH values as low as 3.6. Research regarding the speciation of different metals and the availability of these metals is, however, suggested as well as more frequent monitoring of the water quality in order to obtain accurate data.

Information essential to the proper conservation and management of a water system, can therefore be gathered by determining the concentration of metals in the sediments. The contamination of polluted sediments can also be assessed from the distribution of the

3-47 Water and Sediment

metals on the different sediment grain sizes. The highest metal concentrations are found in the smallest grain size and may therefore be related to the surface area of the particle. Furthermore, the use of sediment based toxicity tests should be investigated as it would provide some indication of the toxicity of the specific sediment. Metal concentrations in the sediment also suggested that metal contamination may be linked to the concentrated levels recorded in the surface waters, as metals can be remobilized from the sediments when water quality, for example the pH, changes. Water and sediment analyses are therefore of primary importance in the framework of environmental investigation.

3-48 Water and Sediment

3.6 REFERENCES

AMERICAN PUBLIC HEALTH ASSOCIATION. (1989). Standard methods for the examination of water and wastewater. (16th Edition). APHA, AWWA & APCF Joint Publication. Washington D.C. pp. 1269.

BOYD, C.E. (1982). Water Quality Management for Pond Fish Culture. Elsevier Scientific Publishing Company, New York. 318 pp.

BROWN, V.M., SHAW, T.L. & SHURBAN, D.G. (1974). Aspects of water quality and the toxicity of copper to rainbow trout. Water Res. 8:797 - 803.

CAIRNS, J. (Jr), HEATH, A.G. & PARKER, B.C. (1975). The effects of temperature upon the toxicity of chemicals to aquatic organisms. Hydrobiologia. 47(1):135 - 171.

CAMPBELL, P.G.C. & STOKES, P.M. (1985). Acidification and toxicity of metals to aquatic biota. Can. J. Fish. Aquat. Sci. 42:669 - 675.

* CHAO, T.T. & THEOBALD, P.K. (1976). Economic Geology. 71: 1560 - 1569.

DALLAS, H.F. & DAY, J.A. (1993). The Effect of Water Quality Variables on Riverine Ecosystems: A review. WRC Project No 351. pp. 68 - 79.

DAY, J.A. (1990). Pitfalls in the presentation of chemical data. Sthn. Afr. J. Aquat. Sci. 16:2 - 15.

3-49 Water and Sediment

DU PREEZ, A.L. (1985). The chemical composition of Transkei river water. Water SA. 11(1):41 - 47.

DU PREEZ, H.H. & STEYN, G. (1992). A preliminary investigation of the . concentration of selected metals in the tissues and organs of the tigerfish (Hydrocynus vittatus) from the Olifants River, Kruger National Park, South Africa. Water SA. 18(2):131 - 136.

ENVIRONMENT CANADA (1987). Canadian water quality guidelines. Report prepared by the Task Force on the water quality guidelines of the Canadian Council of Resource and Environment Ministers. 407 pp.

ELLIS, M.M. (1937). Detection and measurement of stream pollution. Bull. Bur. Fisheries. 48: 365

ELLIS, M.M. (1989). Surface water pollution and its control. The Macmillan Press Ltd., London. pp. 373.

ENVIRONMENTAL PROTECTION AGENCY. (1973). The control of pollution from hydrographic modifications. EPA 43-0/9-73-017, U.S. Government. Printing Office, Washington, D.C.

EUROPEAN INLAND FISHERIES ADVISORY COMMISSION. (1969). Water quality criteria for European freshwater fish - extreme pH values and inland fisheries. Prepared by EIFAC working Party on Water Quality Criteria for European freshwater fish. Water Research. 3: 593.

3-50 Water and Sediment

FELTS, P.A. & HEATH, A.G. (1984). Interactions of temperature and sublethal environmental copper exposure on the energy metabolism of bluegill, Lepomis macrochirus. J. Fish. Biol. 25:445 - 453.

FORSTNER, U. & PROSI, F. (1979) Heavy metal pollution in freshwater ecosystems. In: Biological aspects of freshwater pollution. 0. Ravera [ed.] Pergamon Press, Oxford. pp. 161.

FRANSON MARY ANN, H. (Man. Ed.) (1989). Standard methods for the Examination of water and wastewater (17th edn.). American Public Health Association. Port City Press, Maryland, USA.

FREEDMAN, B. (1989). Environmental ecology. The impacts of pollution and stresses on ecosystem structure and function. Academic Press, Inc. New York. 424 pp.

GALVIN, R.M. (1996). Occurrence of metals in waters: An Overview. Water SA. 22(1):7 - 18.

GARDINER, J. (1974). The chemistry of cadmium in natural water. I: A study of cadmium complex formation using the cadmium specific-ion electrode. Water Res. 8:22 - 38.

GIESY, J.P. & WIENER, J.G. (1977). Frequency distributions of trace metal concentrations in five freshwater fishes. Trans. Am. Fish. Soc. 106:393 - 403.

GIESY, J.P., BRIESE, L.A. & LEVERSEE, G.L. (1978). Metal binding capacity of selected maize surface waters. Environ. Geol. 2:257 - 803.

3-51 Water and Sediment

HELLAWELL, J.M. (1986). Biological indicators of Freshwater Pollution and Environmental Management. Elsevier Applied Science Publishers Ltd., London. pp. 546.

KEMPSTER, P.L., HATTINGH, W.A.J. & VAN VLIET, H.R. (1982). Summarized water quality criteria. Department of Water Affairs, South Africa. Technical Report No. TR108. pp. 45.

* KNIGHT, A.W. (1984). The evaluation of contaminated sediment employing selected benthic freshwater invertebrates. Final Report of the USEPA Cooperative Agreement No Cr-808424, University of California, Davis CA. USEPA Environmental Research laboratory, Corvallis, OR.

KOHN (1991). Sensitiewe visspesie werkswinkel 1991. Kruger National Park Rivers Research Programme. 37 pp.

MASON, C.F. (1991). Biology of Freshwater Pollution, Second addition. Longman Group UK Ltd., England. 351 pp.

MOORE, C.A. (1990). Water Quality Requirements of the Biota of the Kruger National Park Rivers. Report presented at the Workshop on the Preliminary Water Quality Guidelines for the Kruger National Park Rivers. Held in Pretoria from 23 to 24 October 1990. iv + 27 pp.

MOUNT, D.I. (1973). Chronic effect of low pH on fathead minnow survival, growth and reproduction. Water Research. 7:987 - 993.

3-52 Water and Sediment

NATIONAL ACADEMY OF SCIENCES. (1972). Accumulation of nitrate. National academy of Sciences, Washington, D.C.

RAND WATER BOARD (1982). Annual Report (1981/1982). PO Box 1127, Randburg, Republic of South Africa.

SCHOONBEE, H.J. & VAN DER MERWE, C.G. (1989). Investigations into the effects of sewage, industrial and gold mine effluents on the water quality and faunal conditions of a stream in the Transvaal, South Africa. Environmental Quality and Ecosystem Stability. Vol IV-A, Environmental quality. ISEEQS Pub. Eds.: H. Luria, Y. Stelaberger and E. Spanier: Jerusalem, Israel. pp. 401 - 418.

SEYMORE, T., DU PREEZ, H.H., VAN VUREN, J.H.J., DEACON, A. & STRYDOM, G. (1994). Variations in selected water quality variables and metal concentrations in the sediment of the lower Olifants and Selati rivers, South Africa. Koedoe. 37(2):1 - 18. Pretoria. ISSN 0075-6458.

SMITH, L.R., HOLSEN, T.M., IBAY, N.C., BLOCK, R.M. & DE LEON, A.B. (1985). Studies on the acute toxicity of Fluoride ion to Stickleback, Fathead minnow, and Rainbow trout. Chemosphere. 14(9): 1383 - 1389.

SPRAGUE, J.B. (1969). Measurement of pollutant toxicity to fish. Part 1. Bioassay methods for acute toxicity. Water Res. 3:793 - 821.

STANDARD METHODS (1992). Standard Methods for the Examination of Water and Wastewater (18th edn.) American Public Health Association.

3-53 Water and Sediment

STEENKAMP, V.E., DU PREEZ, H.H & SCHOONBEE, H.J. (1994). Bioaccumu- lation of copper in the tissues of Potamonautes warreni (Calman) (Crustacea, Decapoda, Branchiura), from industrial, mine and sewage-polluted freshwater ecosystems. Suid Afrikaanse Tydskr. Dierk. 29(2):152 - 161.

STEENKAMP, V.E., DU PREEZ, H.H., SCHOONBEE, H.J. & SHAFIR, A. (1995). Accumulation of nickel in three freshwater crab populations. S. Afr. J. Wildl. Res. 25(3):67 - 76.

THERON, PRINSLOO, GRIMSEHL & PULLEN CONSULTING ENGINEERS (1991). Water Resources Planning of the Olifants River Basin. Annexure 19: Water quality. pp. 86.

TRAIN, R.E. (1979). Quality criteria for water. US Environmental Protection Agency, Washington D.C. Castle House Publications. pp. 256.

USEPA (United States Environmental Protection Agency) (1976). Quality Criteria for Water. National Technical Information Service. PB - 263 943. Office of Water and Hazardous materials, Washington D.C. EPA, 26 July 1976).

3-54 Water and Sediment

Water SA. 20(3):205 - 212.

VAN DEN HEEVER, D.J. & FREY, B.J. (1996). Human health aspects of certain metals in tissue of the African sharptooth catfish, Clarias gariepinus, kept in treated sewage effluent and the Krugersdrift Dam: Chromium and mercury. Water SA. 22(1):73 - 78.

VAN DER MERWE, C.G., SCHOONBEE, H.J. & PRETORIUS, J. (1990). Observations on concentrations of the heavy metals zinc, manganese, nickel and iron in the water, in the sediments and in two aquatic macrophytes, Typha capensis (Rohrb.) N.E. Br. and Arundo donax L., of a stream affected by goldmine and industrial effluents. Water SA. 16(2):119 - 124.

VARIAN (1989). Flame Atomic Absorption Spectrometry. Analytical methods. Varian Techtron Pty limited, Australia. 146 pp.

WETZEL, R.G. (1983). Limnology. Saunders College Publishing. Philadelphia, New York, Chicago. 767 pp.

WOOD, J.M. (1974). Biological cycles for toxic elements in the environment. Science. 183:1049 - 1053.

* These articles were not reviewed by the author

3-55 Water and Sediment

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3-56 Water and Sediment

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3-57 Water and Sediment

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BIOACCUMULATION OF ZINC AND COPPER IN THE TISSUES AND ORGANS OF CLAMS GARIEPINUS AND LABEO UMBRATUS

4.1 INTRODUCTION

Metals such as zinc (Zn) and copper (Cu), in trace concentrations, are required for normal physiological functions in aquatic organisms. These metals are normally found in low concentrations in nature, but the rapid industrial development and extensive use of metals for manufacture, led to the rising of metal concentrations in the environment. Above certain threshold levels, metals become toxic and it is therefore important that the toxicity of these elements to aquatic organisms is known. Fish, an important food source to humans, is known to accumulate many substances in their tissues. Accumulation patterns are related to the interdependency of uptake and elimination rates of elements (Cross, Hardy, Jones & Barber, 1973). Uptake is considered to be passive and involves diffusion down gradients created by adsorption or binding of metals to the tissue and cell surfaces (Bryan, 1976). A study by Abdullah, Banks, Miles & O'Grady (1976), suggested that the relationship between the fish and the environment is a direct one, but the concentrations of metals in food have also been shown to play an important role in determining the concentrations of metals in fish tissue. Metals will be accumulated to significant levels by organs such as the liver and kidney which secrete specific metal binding proteins (detoxification mechanism) and organs which are the targets of toxicant action (Stagg & Shuttleworth, 1982). Low molecular weight proteins in the liver and kidney, known as metallothioneins, bind metals and thereby reduce their toxicity. Metallothionein-like proteins capable of binding copper, have been described in the liver of Cyprinus carpio by Yamamoto, Ishii & Ikeda, (1978) and also zinc in the liver and kidney of a variety of teleosts (Olson, Squibb & Cousins, 1978).

4-1 Bioaccumulation of Zinc and Copper

Essential and metabolically useful trace elements, such as zinc, are concentrated in the tissues and organs of fish, in excess of requirements (Bowen, 1972), and may represent a store, particularly in liver and skeletal material. Zinc is a common trace metal, essential in minute quantities for fish, but larger amounts are known to be toxic to fish and other aquatic organisms. Zinc toxicity to aquatic organisms is dependant on the water quality, for instance, zinc salts in solution are less toxic to the three-spined stickleback (Gasterosteus aculeatus) in hard water than in soft water (Jones, 1938). In natural fresh waters, the solubility of zinc (Zn'), is essentially controlled by the solubility of zinc carbonate. This is a function of the concentration of carbonate ion and is dependant upon the pH value and concentration of bicarbonate ion in the solution (Soli* 1973). Many factors can, however, also affect the degree of sensitivity shown by an organism to different toxicants, for example the specific diet, the season of the year and the water quality variables such as temperature, pH and hardness indicated as CaCO3.

Zinc causes a variety of deleterious physiological disturbances in salmonids under varying conditions, for example, impaired branchial ion regulation and tissue melano- macrophage induction, tissue inflammatory response (Hughes & Gray, 1972), inhibition of branchial ATPases (Shepard & Simkiss, 1978), impaired antibody production (O'Neill, 1981) and erythrocyte haemolysis (Kodama, Ogata & Yamamoto, 1982). Acutely toxic concentrations of dissolved zinc probably kill fish by altering the gill structure (Skidmore & Tovell, 1972) and subsequent impairment of gaseous exchange (Skidmore, 1970).

Copper, not only an essential element for animals for the synthesis of haemoglobin, formation of bone, maintenance of myelin within the nervous system and as an essential component of key metalloenzymes, is also extensively used in sections of the nonferrous metal industry. The effluent from these industries contains traces of copper, and when

4-2 Bioaccumulation of Zinc and Copper

discharged into water bodies, become available to aquatic organisms, resulting in lethal and sublethal effects. The toxicity of copper is largely attributable to Cu", which forms complexes with chloride, amine, sulphocyanide, bromide, iodide, nitrate, oxalate and pirophosphate, and soluble salts with the exception of ferrocyanide, sulphide and carbonate (Pourbaix, 1966). The toxicity of Cu' is increased by a reduction in water hardness, temperature, dissolved oxygen, chelating agents such as EDTA and NTA, humic acids, amino acids and suspended solids. In the aquatic environment, copper can occur in three physical states, namely soluble, which can pass through a 0.45 Am membrane filter and includes copper as a free cupric ion and as soluble complexes. Soluble copper concentration is the sum of the concentrations of cupric ion and the copper carbonate complex. Secondly particulate, which includes oxide, sulphide and malachite (Cu2(OH)2CO3) precipitates as well as insoluble organic complexes and copper adsorbed on clays and other mineral solids and thirdly colloidal which include polypeptide material and some clays and metallic hydroxide precipitates.

In this section of the study, the levels of zinc and copper were determined in the different tissues and organs of Clarias gariepinus and Labeo umbratus (Fig 4-1) from the Klein Olifants River (Locality KOR1) and the Olifants River (Locality OR1) (See Chapter 2, Fig 2-5). This data were required to assess the possibility of elevated levels of these metals in fish due to possible point and diffuse sources of pollution in the upper catchment of the Olifants River. The species, size, sex, localities and seasonal dependence of the bioaccumulation of these two metals in the tissues and organs of the two species, were specifically addressed.

4-3 Bioaccumulation of Zinc and Copper

4.2 MATERIALS AND METHODS

4.2.1 FIELD SAMPLING Two freshwater species of fish (C. gariepinus and L. umbratus), were collected during the period February 1994 to May 1995, at localities KOR1 and OR1 (Fig 4-1), by means of gill nets (70 - 120 mm stretched mesh size) in the Olifants River and Klein Olifants River (Chapter 2, Fig 2-5) respectively. For C. gariepinus, the mass and total lengths of the fish were recorded after capture, and for L. umbratus the mass, fork length, and total length of the fish were recorded. The fish were then dissected on a polythene dissection board, using clean, stainless steel tools, wearing surgical gloves. The skin and muscle (Fig 4-2) were removed, placed into clean, pre-washed glass bottles and frozen until further analysis in the laboratory. The liver and gills (Fig 4-2) were removed and the total mass was determined, after which the gill filaments were removed, placed in clean, pre-washed glass bottles and frozen until further metal concentration analysis.

4.2.2 LABORATORY PROCEDURES All glassware was soaked in a 2 % Contrad soap solution (Merck chemicals) for 24 hours, rinsed in distilled water, acid-washed in 1M Ha for 24 hours and rinsed again in distilled water (Giesy & Wiener, 1977), prior to use.

Prior to sample preparation, the tissues were thawed and rinsed in distilled water to remove excess mucus coating, or other foreign particles that could have adsorbed metals. Approximately 5 grams of each sample was dried in an oven at 60 °C for a period of 48 hours. In order to determine the percentage of moist of each sample, the wet and dry mass of the samples were recorded. The samples were digested by adding concentrated nitric acid (55 %) and perchloric acid (70 %) to one gram of dry tissue, in a 2:1 ratio in

4-4 Bioaccumulation of Zinc and Copper

FIG 4-1 THE AFRICAN SHARPTOOTH CATFISH, CLAMS GARIEPINUS (A) AND THE MOGGEL, LABEO UMBRATUS (B)

4-5 Bioaccumulation of Zinc and Copper

FIG 4-2 THE DIFFERENT TISSUES AND ORGANS ANALYSED: A - THE SKIN (Al), (A2), MUSCLE (A3) LIVER AND GILLS (A4) OF CLARIAS GARIEPINUS. B - (B1), LIVER (B2), THE SKIN MUSCLE (B3) AND GILLS (B4) OF LABE() UMBRATUS

4-6 Bioaccumulation of Zinc and Copper

a 50 ml Erlenmeyer flask and was performed on a hot plate at 200 - 250°C for ± four hours until the solutions were clear (Van Loon, 1980). After digestion, each solution was filtered through an acid resistant 0.45 um filter paper under vacuum. After each sample had been filtered, the filtering system was rinsed with distilled water and each sample was made up to 50 ml with distilled water and stored in pre-washed glass bottles, until determination of metal concentrations.

To determine the metal concentrations in the tissue samples of the fish, a Varian Atomic Absorption Spectrophotometer (Spectra AA-10) was used (Varian, 1989). Analytical standards for copper and zinc were prepared from Holpro stock solutions. To assure accurate and precise determination of trace elements in freshwater biological samples, a standard tissue sample (IAEA/R1/64) was used. Metal contamination from the laboratory was avoided and a triplicate acid blank was also analysed. The metal concentrations of the tissue samples were calculated as follows:

Metal concentration (µg/ g) AAS reading (pg/l) x Sample volume (50 ml) Sample mass (g) 1

Bioconcentration factors between the fish tissues and the water (BF,„) and the sediment (BFO were also calculated by using the following formula (Wiener & Giesy, 1979):

BF,„ / BFs = metal] in organs\tissues (ug/g dry mass) metal] in water (mg/1) or sediment (ug/g)

4-7 Bioaccumulation of Zinc and Copper

Only the mean metal concentrations (zinc & copper) in each organ/tissue were used for the calculations of the bioconcentration factors.

4.2.3 STATISTICAL PROCEDURES Statistical analyses were performed by using the STATISTICA for Windows programme. The different statistical analyses were only performed on certain data, as sample sizes did not allow the analyses of all the data for all the cases. The capturing success varied and it was not possible to obtain large numbers (preferably >20 individuals, Seymore, 1994) of each fish species at both localities during a specific survey. The low capturing success and the capturing of fish which varied in mass and length, did not enable the selection of fish of a specific size and sex for analyses. Because the sample sizes in this study were small, nonparametric methods were used as these methods are most appropriate in such cases. These methods do not make any assumptions about the distribution (e.g. the normality) of the samples and are sometimes referred to as distribution-free tests. For the comparison of two groups, for example, male & female, localities, seasons and species, the Kolmogorov-Smirnov two-sample test was used. This test is sensitive to differences in the general shapes of the distributions in the two groups, i.e., to differences in, for example, dispersion and skewness. For the comparison of multiple groups, in this case the bioaccumulation of zinc and copper in the skin, muscle, liver and gills, the Kruskal-Wallis analysis of ranks was used. The interpretation of the Kruskal-Wallis test is basically identical to that of the parametric one-way ANOVA, except that it is based on ranks rather than means. To express the relationship between two groups, a correlation coefficient was determined. A nonparametric equivalent to the standard correlation coefficient is the Spearman R test, used in this study to determine the relationship between the metal concentrations in the different tissues/organs and the lengths of the fish.

4-8 Bioaccumulation of Zinc and Copper

4.3 RESULTS

4.3.1 FISH SIZE The mean total lengths of the fish caught at localities KOR1 and OR1 during the period February 1994 to May 1995, ranged between 40 - 68.5 cm for C. gariepinus and 43 - 77 cm for L. umbratus, while the mass of the fish was in the range of 0.2 - 2.8 kg for C. gariepinus and 0.8 - 3.5 kg for L. umbratus, as indicated by the data in Table 4-1.

4.3.2 DIFFERENCES IN BIOACCUMULATION OF ZINC AND COPPER IN THE DIFFERENT TISSUES/ORGANS The moisture content of the individual tissues differed from one another, and to some extent also from one individual to the next, with the mean percentage of moisture being 69 ±2%, 73 -± 2%, 76 ± 3% and 80 ±2 % for the skin, liver muscle and gills of L. umbratus respectively and 72 ± 3%, 77 ±1%, 80 ± 2% and 80 ±2 % in the skin, liver muscle and gills respectively of C. gariepinus. The samples were therefore handled on a dry mass basis instead of a wet mass basis, in order to minimize variation. Tissue/organ differences in the zinc bioaccumulation (Fig 4-3; Table 4-1 Appendix) for C. gariepinus were not clear to distinguish. Nevertheless, the order of bioaccumulation ranged between the highest levels in the gills (G) and liver (L), followed by lower levels in the skin (S) and muscle (M). No significant difference (P>0.05) was found between the liver and skin and the gills and the skin of C. gariepinus at localities KOR1 and OR1 respectively (Table 4-2). For L. umbratus, the order of bioaccumulation for Zn was more apparent, being G>L>S>M (Fig 4-4; Table 4-1 Appendix). No significant difference (P > 0.05) was found between the liver and skin and the gills and skin of L. umbratus at locality KOR1 (Table 4-3).

The general order of bioaccumulation for Cu by both species was: L>G>M=S with

4-9 Bioaccumulation of Zinc and Copper

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4-10 Bioaccumulation of Zinc and Copper

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4-11 Bioaccuniulation of Zinc and Copper

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4-12 Bioaccumulation of Zinc and Copper

300 200

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FIG 4-3 THE MEAN ZINC CONCENTRATIONS (ug/g DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF CLARIAS GARIEPLVUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

4-13 Bioaccumulation of Zinc and Copper

200 200

150

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4-14 Bioaccumulation of Zinc and Copper

TABLE 4-2 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P:50.05; NON-SIGNIFICANT = P>0.05)) BETWEEN THE ZINC CONCENTRATION IN THE TISSUES AND ORGANS OF CL4RIAS GARIEPINUS DURING THE PERIOD FEBRUARY 1994 (F1), MAY 1994 (M 1), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F 2) AND MAY 1995 (M2)

TABLE 4-3 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P . 0.05; NON-SIGNIFICANT = P>0.05) BETWEEN THE ZINC CONCENTRATION IN THE TISSUES AND ORGANS OF LABE() UMBRATUS IN AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

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N N, F2 M2 A, N, F2

A, N, F2 , M2 N, F2

A, N, F2 M2

4-15 Bioaccumulation of Zinc and Copper

very high concentrations in the liver (Fig 4-5; Table 4-2 Appendix). No significant differences between the muscle and skin tissues were found at both localities and between the gills and liver at locality OR1, for C. gariepinus (Table 4-4). The bioaccumulation of copper in the gills and muscle for L. umbratus showed no significant difference at both localities (Fig 4-6; Table 4-5).

The BF,„ for zinc (Table 4-1 Appendix) ranged between 0.2 (calculated in the muscle of L. umbratus in November 1994 at locality OR1) and 4.1 (calculated in the gills of L. umbratus in August 1994 at locality OR1), while the BF, for zinc fluctuated between 0.05 (in the muscle of C. gariepinus in February 1994 at locality OR1) and 1.72 (in the gills of C. gariepinus in February 1994 at locality KOR1).

Bioconcentration factors between the copper concentration (Table 4-2 Appendix) in the organs/tissues of C. gariepinus and L. umbratus and the copper concentration in the water (BF„), ranged from 20 (calculated in the muscle of C. gariepinus in February 1994 at locality KOR1 and in the muscle and skin of C. gariepinus in May 1994 at locality OR1) to 52 700 (calculated in the liver of L. umbratus in February 1995 at locality OR1). The bioconcentration factors between the Cu concentrations in the organs/tissues of the two species and the Cu concentrations in the sediment (BF,), fluctuated between 0.01 (in the skin and muscle of C. gariepinus in May 1994 at locality OR1 and in the skin and muscle of both species in November 1994 at localities KOR1 and OR1 respectively) and 5 (in the liver of L. umbratus in February 1995 at locality KOR1).

433 SPECIES DIFFERENCES Statistical analyses were performed only on data from August 1994 at locality OR1 onward, as sample sizes for the remaining were too small. Zinc and copper bioaccumulation in the different tissues and organs of the two species, varied significantly in most cases. In May 1994 at locality OR1, significant differences (P .0.05) in the zinc

4-16 Bioaccumulation of Zinc and Copper

0 70 /

40 .0

.0'

.0,

.0'

10 10

0 0 Feb 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

/ / El Skin ❑ Liver 0 Muscle •Gills /

FIG 4-5 THE MEAN COPPER CONCENTRATIONS (ug/g DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF CLAMS GARIEPINUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

4-17 Bioaccumulation of Zinc and Copper

400 600

500

,-, 300 to 400 C C 0 ...O ;$' I- 'tillI. .a. 200 4.) ..a. 300 U C U

CO ta.

E 100

100

Al------f---qT-7,--/P- 0 7--7 Aug 94 Nov 94 oA Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

/ / 3 Skin ❑ Liver 0 Muscle iiGills /

FIG 4-6 THE MEAN COPPER CONCENTRATIONS (ug/g DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF LABEO UMBRATUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM AUGUST 1994 TO MAY 1995

4-18 Bioaccumulation of Zinc and Copper

TABLE 4-4 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON-SIGNIFICANT = P>0.05) BETWEEN THE COPPER CONCENTRATION IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS IN FEBRUARY 1994 (F,), MAY 1994 (M1), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

. .. Loc.Aury KOR1

, N, F2 r M2 NS. , N, F2

F, , N, M2 NS.

F, , N ..

LocAury OR1

=7:7 ...... F1, M„ A, F2, M2 ns. M„ A, M2

F„ M, , A, F2 , M2 K. M2

MUSC arr.: ..... F1 MI F2 r M2 ..... •

TABLE 4-5 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON-SIGNIFICANT = P>0.05) BETWEEN THE COPPER CONCENTRATION IN THE TISSUES AND ORGANS OF LABE° UMBRATUS IN AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

rrr +ar

LocAury KOR1

F, , M2 A A, F2

' 111111:1: A, F2 , M2 A, F2 , M2 giniP ■■ swir- ns.

LocAurt OR1

A, N, F2 M2 A, N N, M2

N, F1, M2 A, N, F2

n.s.

......

4-19 Bioaccumulation of Zinc and Copper

concentrations for the two species, were found in the skin and gills (Fig 4-7), with C. gariepinus having higher levels of zinc in the skin, while L. umbratus showed higher levels in the gills. Significant differences were also found in the muscle and liver tissue, with higher levels of zinc in both tissues of C. gariepinus in November 1994 at localities KOR1 and OR1, as well as in February and May 1995 (Fig 4-8). In May and November 1994 and February and May 1995, the Cu liver and muscle concentrations were significantly higher in L. umbratus than in C. gariepinus (Fig 4-9). The Cu gill concentrations were, however, higher in C. gariepinus in November 1994 at locality KOR1 (Fig 4-9).

4.3.4 RELATIONSHIP BETWEEN LENGTHS AND ZINC AND COPPER CONCENTRATIONS Because the sample sizes were too small for individual analysis of the different species at the different localities for each month, these variables had to be grouped together in order to obtain statistically verifiable results. The zinc muscle and skin concentrations showed negative correlations with the lengths of the fish, thus, the larger the fish, the lower the zinc concentrations in these tissues/organs. The zinc concentrations in the liver and gills showed a positive correlation, where an increase in one variable, is associated with an increase in the other. The copper concentrations in the skin, muscle and liver also showed significant negative correlations with the lengths of the fish analysed.

4.3.5 DIFFERENCES BETWEEN MALES AND FEMALES Generally, few significant (P...50.05) differences in the accumulation of zinc and copper in the different tissues/organs were found for C. gariepinus and L. umbratus between males and females. The zinc concentrations were, however, higher in the skin tissues of the males of C. gariepinus in August 1994 and February & May 1995 (Fig 4-10), while the zinc concentrations in the skin of L. umbratus, were higher for the females (Fig 4-11). The zinc concentrations in the gills of C. gariepinus were higher for the males in May

4-20 Bioaccumulationn of Zinc and Copper

A B 200 300

250

.'.'..'.'.

0 May 94 (OR1) Nov 94 (KOR1) Nov 94 (OR1) 0 May 94 (OR1) Feb 95 (OR1) May 95 (OR1)

0 Clarias gariepinus OLabeo umbratus

FIG 4-7 SIGNIFICANT DIFFERENCES (PS 0.05) (MEAN ± S 0), BETWEEN C. GARIEPINUS AND L. UMBRATUS REGARDING THE BIOACCUMULATION OF ZINC IN THE SKIN (A) AND GILLS (B) FROM MAY 1994 TO MAY 1995

4-21 •

Bioaccumulation of Zinc and Copper

A 100 200

In 80

68 150

7) V • O 60 E

0 0 • 100 ' ...... '.• ei

a> c.) 40 U 8 8 N .". C. co ct 50 20

• • • • • • • • • • • • • • • • • .

0 " • ' • Nov 94 (KOR1) Nov 94 (OR1) Feb 95 (OR1) May 95 OR1)

0Clarias gariepinus D Labe° umbratus

)

FIG 4-8 SIGNIFICANT DIFFERENCES (P s 0.05) (MEAN ± SD), BETWEEN C. GARIEPLVUS AND L. UMBRATUS REGARDING THE BIOACCUMULATION OF ZINC IN THE MUSCLE (A) AND LIVER (B) TISSUE FROM MAY 1994 TO MAY 1995

4-22 Bioaccumulation of Zinc and Copper

A 1,000 B 30

(OR1)

'jse j) 25 ti) 800 (OR1)

(OR1) 7.7)

E 20 (OR1) (OR1) 600

0 0 -,e7; 15 0.) w. 8 400 U z (OR1) 8 a. 6- 10 a. 8 8 (OR1) Cl cl) 200 Cl Q 5

(KOR1) (OR1) I 1] May 94 Nov 94 Nov 94 Feb 95 May 95 May 94 Nov 94 Feb 95 2v ay 95

ZOs

30 ti

■ 0 . • • • • . • • . • . • . • . • . • . • . • . • . • • • . • . • ❑ Labeo umbratus . • • • • • • • . • • . • . • Clarias gariepinus 2,1 20

• • . . • . • • • • • • • . • • • • • • • • . . • • • . . • • • • . . • 8 • • • • • • • • . • . . • . • . • • • . • • . • . • . • . • • • . • . • . • . • . • • . • • • • • • • • • • • • • • . • • • • • • • • • • • • • • • • • • • • - • 8 . • . . • . • . • . • . • . • . • . • • . • . • . • • . • . • . • • . • . • . • • . • . • • . • • • . . • . • . • . • . • • . • . . •

• • • • • • • • • • • • . • • . • • • • • • • • . . • • • .

8 ...... ...... • • • • • • • .... • • • • • • • • • 10 . • • • • • • • • . . • • • • • • • . • • • • • • • • • • • • • • • . . • • . • . • . • • • • • • • • • • • • • • • • • • • • •

0 Nov 94 (KOR1)

FIG 4-9 SIGNIFICANT DIkr ERENCES (PS 0.05) (MEAN ± SD), BETWEEN C. GARIEPINUS AND L. UMBRATUS REGARDING THE BIOACCUMULATION OF COPPER IN THE LIVER (A), MUSCLE (B) AND GILLS (C) FROM MAY 1994 TO MAY 1995

4-23

Bioaccumulation of Zinc and Copper

180 A B 140

Skin 160 120

140

100

cf)

• 0 100 80 5 0 U 1;1 8 80 U 60 f. N 8 60 U Muscle

. F4 40

20 20

250

U 150 ❑ Males E Females

8 U N C

FIG 4-10 SIGNIFICANT DIFFERENCES BETWEEN THE MALES AND FEMALES OF C. GARIEPINUS REGARDING THE BIOACCUMULATION OF ZINC IN THE SKIN AND MUSCLE TISSUE IN AUGUST 1994 (A), IN THE SKIN IN FEBRUARY 1995 (B) AND IN MAY 1995 (C)

4-24

Bioaccumulation of Zinc and Copper

A 250 B 140

120

.11). 200 00

rn 100

t3.0 0) 150 80 0 • .—0 -C.1 174

60 100 0 8 8 0 0

N 40 aS 0 50

C 35

C. 20 0 111 Males • Females 173*

15 8 a 8 10

FIG 4-11 SIGNIFICANT DIt. hRENCES BETWEEN THE MALES AND FEMALES OF L. UMBRATUS REGARDING THE BIOACCUMULATION OF ZINC IN THE SKIN IN MAY 1995 (A), IN THE GILLS IN MAY 1994 (B) AND THE BIOACCUMULATION OF COPPER IN THE LIVER OF C. GARIENNUS IN FEBRUARY 1994 (C) AT BOTH LOCALITIES

4-25 Bioaccumulation of Zinc and Copper

1994 (Fig 4-11). Significant differences (Ps 0.05) were found between the sexes for copper, with higher concentrations in the liver tissue of the females at locality KOR1 in February 1994 for C. gariepinus (Fig 4-11).

43.6 SEASONAL DIFFERENCES Statistical analyses between the seasons, were mainly carried out for C. gariepinus, except for the comparison of the bioaccumulation of the two metals in the different tissues and organs of the fish in winter and spring 1994, where L. umbratus was used, as a result of small sample sizes in the case of C. gariepinus. Comparisons between the different seasons for zinc and copper, yielded significant differences (P 0.05) in the different tissues/organs of the two species. The significant differences appeared to be more likely with regard to the zinc concentrations in the skin, followed by the liver & muscle and then the gills. Comparisons between autumn and summer 1995, showed no significant differences regarding the bioaccumulation of zinc, while seven of these cases were found for copper comparing the different seasons of 1994 and 1995. Only in comparison between autumn 1994 and summer 1995, all four the tissue types showed significant differences regarding the bioaccumulation of zinc and copper. For copper, the differences were mostly significant in the gills, followed by the muscle, the liver and the skin (Table 4-6; Table 4-7).

4.3.7 LOCALITIES DIFFERENCES The data indicate several significant differences (P50.05) between the two localities with regards to the bioaccumulation of zinc and copper (Fig 4-12) in the different tissues/organs of the two species. In February 1994 the two localities differed significantly from each other with regards to the zinc liver concentration, with the highest concentrations found at locality OR1 (Fig 4-12). In February and May 1995, the zinc

4-26 Bioaccurnulation of Zinc and Copper

tr) o V") t/1 0 0 CD 0 0 CD CD. 0 0 6 6 c=5 c=5 6 6 6 6 (VVV vl V vi V V vi t/ vi V vi V a a. a a. a a.. 0 CI

e-1 111% C) C) 0 C) 0 0 6 6 0 6 6 6 00 V vi vi VVV V V vi vi P. P. a. a a a. a a.

-• 0 v-4 0000 0 6 6 c; V V vi vi P. P. P. c

1,1 0 0 0 0 vi vi V a = d a.

4-4 c) 1.4 CZ) C) CZ) C.) 6 6 6 vi V V a. P..

4.) .5 5 > •Z1 >°.) a'12. • •-• *-• • -■ cn CD V) ,4 V) •-.1 .4 0 ificant n ig s t No = vi

4-27 Bioaccumulation of Zinc and Copper

ficant i n ig s t No = vi

4-28 Bioaccumulation of Zinc and Copper

A 250 1,000

)

""eB 200 /g 800 g (p,

a.) liver

150 t 600 in

ion

0 t a tr

a) cen Q 100 400 .....

8 con

c.) er

ceS copp

50 200 Mean

0 Feb 1994 Feb 1995 May 1995 Aug 1994 May 1995

❑ Locality KOR1 0 Locality OR1

FIG 4-12 SIGNIFICANT DIFFERENCES BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF ZINC IN THE LIVER OF BOTH SPECIES (A) AND THE COPPER BIOACCUMULATION IN THE LIVER OF L. UMBRATUS (B)

4-29 Bioaccumulation of Zinc and Copper

concentrations in the liver were the highest at locality KOR1, as was the copper liver concentrations in May 1995 (Fig 4-12). Locality KOR1 (Klein Olifants River) differed significantly from locality OR1 (Olifants River) in February 1994, with regards to copper muscle and skin concentrations, with the highest concentrations found at locality KOR1 (Fig 4-13). For L. umbratus the two localities differed significantly in August 1994 with respect to copper liver (Fig 4-12) and muscle (Fig 4-13) concentrations, with locality OR1 showing the highest concentrations.

4.4 DISCUSSION

A majority of studies related to metal concentrations in aquatic organisms, emphasize bioaccumulation of several metals in different tissues/organs of different species of fish (Bezuidenhout, Schoonbee & De Wet, 1990; Du Preez & Steyn, 1992; Seymore 1994; Van den Heever & Frey, 1994). A compound or elements such as metals, if bioavailable, can be absorbed from the water, by means of gills or epithelial tissues, and concentrated by the body. Once absorbed, the blood transports the element to a storage point, like fat, muscle and bone, or the kidney and liver for storage and/or transformation (Heath, 1987). Elements transformed by the liver, may be stored there and excreted in the bile or stored in extrahepatic tissue, like fat. It may also be passed back into the blood for possible excretion by the kidney or gills.

As can be seen in this study, the liver is a site of high bioaccumulation of especially copper as reflected in L. umbratus at locality OR1 with values between 50 and 568 p,g/g dry mass. These values were very high compared to values found for Barbus marequensis (Seymore, 1994), which also accumulated the highest copper concentrations in its liver. Copper concentrations were also higher in the liver than the

4-30 Bioaccumulation of Zinc and Copper

30 25

7-,

Feb 1994 Aug 1994 Feb 1994

/ ❑ Locality KOR1 El Locality OR1

)

FIG 4-13 SIGNIFICANT DIFFERENCES BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF COPPER IN THE MUSCLE OF BOTH SPECIES (A) AND THE SKIN OF C. GARIEPINUS (B)

4-31 Bioaccumulation of Zinc and Copper

zinc concentrations. This may be due to the metal binding proteins in the liver. Harrison & Lam (1986) found that proteins in the liver of the bluegill, Lepomis macrochirus appeared to bind copper and zinc. Their data indicate that the copper may have displaced the zinc from the proteins in the low molecular weight fractions of the metaloproteins in the livers from fish collected from a discharge site from a power plant. The ability of fish to regulate essential metals, should therefore also be taken into account. The liver is in an advantageous position for clearing the blood of substances entering the circulation from the gastro-intestinal tract. Since the blood from the gastro- intestinal tract passes through the liver before reaching the systemic circulation, theoretically the liver can remove toxicants from the blood, biotransform them, or excrete them into the bile and thus prevent their distribution to other parts of the body. Metal concentration in the liver therefore reflects its multifunctional role in detoxification and storage and for that reason fish livers were analysed to investigate their possible use as indicators of trace element pollution in the freshwater environment.

Zinc and copper accumulation has also been shown in this study, to occur in the organs concerned with osmotic and ionic regulation in fish exposed to elevated levels of the metals in the water, with zinc values in the gills ranging between 88 - 257 pg/g dry mass and 139 - 184 ktg/g dry mass for C. gariepinus and L. umbratus respectively. If the results for zinc in the gills for C. gariepinus are compared with levels reported by Bezuidenhout et al. (1990) collected from the Germiston lake in the Transvaal, which is affected by mining and industrial effluents, similar levels were detected in the gill tissues collected during this study. Copper concentrations, on the other hand, fluctuated between 5 and 29 pg/g dry mass for C. gariepinus, which were lower than levels recorded by Bezuidenhout el al. (1990) in the Germiston lake. These values were, however, relatively high in comparison to values found for B. marequensis in the Lower Olifants River and Selati River (Seymore, 1994), ranging between 81 - 198 //g/g dry mass for zinc and 5 - 18 Ag/g dry mass for copper. It is important to note the different

4-32 Bioaccumulation of Zinc and Copper

pathways by which the metals can enter the fish. According to literature, the four different routes for pollutants such as metals to enter the fish are via the food items ingested, through drinking water (Eddy, 1981), through absorption via the skin and by means of absorption through the gills (Matthiessen & Brafield, 1977; Heath, 1987). The uptake of metals in the ionic form through the gills is believed to be simple diffusion, possibly through pores, as there is no evidence for active transport of pollutant metals into fish via the gills, although cases of carrier mediation may occur, as is the case with calcium (Heath, 1987). It is clear that the gills are very important, due to their intimate contact with the external environment and importance as an effector of osmotic and ionic regulation and it is clear for L. umbratus that zinc uptake occurs primarily in the gills. Zinc has a twofold influence on the gills, namely bioconcentration of metals and the structural cellular alterations that were noted on the gill's of Tilapia sparmanii in a study performed on the effects of zinc on the gills (Van Rensburg, 1989). Zinc concentrations were higher in the gills and because the gills also serve as a regulator, thus diminishing the toxicity excretion. Therefore less zinc is concentrated in the body than in the gills. The liver and gills can also act as depo tissues, where the uptake of metals significantly exceeds the elimination and these organs will accumulate the metals.

The zinc and copper muscle and skin concentrations on the other hand, are relatively low, in comparison to values obtained from the mine and industrial polluted Germiston lake, which are an indication of good regulation of the metals in these tissues (Bezuidenhout et al., 1990). The concentrations of these metals in these tissues are very important, as zinc and copper are essential elements for human beings, with zinc being present in several amino acids and numerous enzymes such as alcohol dehydrogenase, carbonic anhydrase and carboxy peptidase (Gray & Bertini, 1986), involved respectively in ethanol and protein metabolism. Zinc is also contained in alkaline phosphatase and superoxide dismutase enzymes and therefore has a great

4-33 Bioaccumulation of Zinc and Copper

influence on human growth, encephalic development, reproduction, the immunological system and correct cellular metabolism. Copper, with iron and cobalt is involved in the production of haemoglobin and erythrocytes and therefore, in the bone and nervous system formation.

The bioavailability of these metals should also be taken into account. It is assumed that lower bioconcentration factors for a metal indicate lower bioavailability. The bioconcentration factors for zinc, between the tissues/organs and the water, ranging between 163 and 1400, were mostly very high, compared to those found by Du Preez et al. (1992) for the tigerfish Hydrocynus vittatus, in the Olifants River, Kruger National Park, with values between 62 and 303 (assuming moisture content of 76 % for muscle tissue, 73 % for liver tissue and 80 % for gill tissue). These high bioconcentration factors for zinc for C. gariepinus and L. umbratus in the Olifants River and Klein Olifants River are therefore an indication that these metals are highly available for uptake by these species. The bioconcentration factors between the tissues/organs and the water for copper, ranged between 20 and 52 700, which were much higher than values obtained by Du Preez et al. (1992), ranging between 234 and 937 (dry mass, assuming moisture content of 76 %, 73 % and 80 % for the muscle, liver and gill tissues respectively), and also higher than values found by Seymore (1994) in the Lower Olifants River, the Selati River and Pionier Dam (Kruger National Park), fluctuating between 10 and 18 090. The lowest and highest bioconcentration factors for zinc and copper, do not necessarily coincide with lower and higher zinc and copper concentrations in the water at the specific localities. It is, however, important to note that the total concentration of the metal in the water does not play the major role in availability of that metal, but the metal species, as well as the physico-chemical conditions of the waters are very important, as it determines the toxicity and speciation of the metals. More research is needed regarding the bioavailability of metals and the different forms of these metals in the water and sediment at the specific localities.

4-34 Bioaccumulation of Zinc and Copper

The bioconcentration factors between the tissues/organs and the sediment for zinc and copper, were generally lower than the bioconcentration factors between the tissues/organs and the water, which indicates even a smaller fraction of these metals in the sediments are actually available for accumulation by these species at these localities in the system. The bioconcentration factors between the different tissues/organs of the two species and the sediments found in this study were also lower than bioconcentration factors calculated for the Lower Olifants River Catchment. This lower availability of zinc and copper from the sediments, can be ascribed to the formation of complexes with suspended solids or inorganic complexes (for example, carbonate) in the fairly hard water (mean total alkalinity: 79 mg/I CaCO3 at locality KOR1 & 86 mg/1 CaCO3 at locality OR1, see Table 3-1 Appendix) in this part of the system, therefore reducing their bioavailability. It should also be taken into account that the bioconcentration factor would not give the accurate indication of the relative bioavailability of zinc and copper uptake, if these metals are regulated in the fish (Wiener & Giesy, 1979), as seemed to be the case in the muscle and skin tissues of these species.

Zinc and copper accumulation in the different tissues/organs have also been shown to be a species specific function. A significant difference (P5.0.05) was found in the zinc gill concentration between the two species, with the highest zinc concentrations found in L. umbratus. This can be expected, as L. umbratus is dependant on only its gills for breathing and the gills are an important route for metal uptake and the metals are absorbed in the ionic form by the gills in high concentrations, while C. gariepinus is also air breathing, using branched air breathing organs, situated in the cavities above the gill arches. Another important factor to consider, regarding the lower zinc concentrations in the gills of C. gariepinus, is the ability of the catfish to excrete zinc, as well as copper (Bryan, 1976), with the loss of the metal occurring via the gills, probably involving the mucus layer covering this organ (Varanasi & Marky, 1978). This should be further investigated for C. gariepinus as the copper concentrations were higher in the gills of

4-35 Bioaccumulation of Zinc and Copper

C. gariepinus than in the gills of L. umbratus. This can, however, be explained, as the gill tissue acts as a depo tissue, accumulating copper to a certain maximum concentration, in this case not yet reached, after which it is excreted (Francis, 1994). It was also observed that the concentrations of zinc in the skin, muscle and liver were higher in comparison to that in the gills in C. gariepinus than in L. umbratus. A relationship between metal concentrations in fish and the trophic level of the species studied, is also suggested, indicating the importance of the food pathway. Mathis & Cummins (1973) found that significantly higher concentrations of zinc were present in omnivorous fish. Thus, the feeding patterns of C. gariepinus, which is omnivorous and catches living prey and eats any organic material, including aquatic weeds, detritus, as well as fish, birds, frogs, small mammals, reptiles, snails, crabs, insects and plant material, are important (Van der Waal, 1972). L. umbratus on the other hand, feeds on soft sediment and organic material and this can be the main reason for the significant differences between the two species with higher zinc concentrations in the skin, liver and muscle of C. gariepinus (Skelton, 1993). The higher concentrations of zinc in the skin and muscle of C. gariepinus can be ascribed to the fact that the catfish does not have scales, facilitating the uptake of the metal through the skin and muscle tissues. The highest copper concentrations were however found in the liver and muscle of L. umbratus, while the highest concentrations in the gills were observed for C. gariepinus.

The accumulation of metals in the different tissues/organs of the two species, also differs between the males and females, but no definite trend could be established, although it seemed as if the males of C. gariepinus accumulated higher concentrations of zinc in the skin and muscle tissues, whereas the females of L. umbratus accumulated the highest concentrations in the skin tissues. It is suggested that the gonads of the different sexes should be examined to determine the actual differences between the males and females regarding the bioaccumulation of zinc and copper in the different tissues and organs of

4-36 Bioaccumulation of Zinc and Copper

the two species. This is supported by the findings of Du Preez et al. (1992) who performed a study on the concentrations of metals in the tissues and organs of the tigerfish, H. vitalus from the Olifants River, Kruger National Park and also a study performed on other aquatic animals, including the tissues of the freshwater crab, Potamonautes warreni from industrial, mine and sewage-polluted freshwater ecosystems (Steenkamp, Du Preez & Schoonbee, 1994). Seymore (1994) suggested that female fish require greater amounts of zinc, necessary for gonad development. The concentrations in the gonads therefore increased until the fish were sexually mature. These findings also showed that when the zinc concentrations in the gonads (especially the female gonads) decreased, the zinc concentrations increased in the internal sources (skin, muscle and liver) and also the other way around. The specific role of copper in the development of the gonads is not certain, but it seems as if copper is required for certain stages of the development and zinc for others (Seymore, 1994).

In this study, there was a relationship between the bioaccumulation of zinc and copper concentrations and the size of the fish, with the lengths of the fish, ranging between 34 and 75 cm. The zinc and copper concentrations in the skin and muscle, were lower with increasing fish size, which can be related to new tissues being incorporated at a greater rate than metals can be actively transported into the tissues to establish a steady state concentration dilution by growth (Cross, et al., 1973). Small fish also have a higher metabolic rate per unit body mass than large fish and hence require relatively more oxygen (Winberg,1956; Matthiessen et al., 1977) and this requirement is met via a higher rate of flow over the gills and by having a larger gill surface area/g body mass than larger fish (Hughes, 1970). Thus, increasing metal concentrations can be related to decreasing fish size. Although copper concentrations in the liver showed a negative correlation, the zinc concentration in this organ increased with fish size. This increase can be accounted for by physiological requirements for survival that dictate the maintenance of certain levels of elements. If no relationships would be found, the

4-37 Bioaccumulation of Zinc and Copper

concentration could be due to less rapid growth which allows a steady state of metal uptake and elimination from the tissues as would be the case in adult fish. It must be stressed that the comparison was based on the fish combined over a time period, which may influence the comparison. A more accurate evaluation would be possible if a definite size range of fish was captured at the same time. Nevertheless, the present data and those of previous studies (De Wet, Schoonbee, De Wet & Wiid, 1994) indicate that the size of the fish must be considered, especially when dealing with juveniles compared to adult fish.

SEASONAL DIFFERENCES The summer of 1994 differed significantly from the other seasons with respect to the zinc and especially the copper concentrations in the various tissues/organs of C. gariepinus and L. umbratus, showing higher concentrations of these metals at both localities. The copper and zinc concentrations in the summer of 1994 were not necessarily higher in the water (see Chapter 3, Table 3-2), which may indicate that these metals may be biomagnified (accumulated through food) rather than bioconcentrated (accumulated through water) by the fish. It must be noted that only one water sample was collected at a certain point and the metal concentrations can also be a result of the regulation mechanism of the tissues/organs, as these act as depos, where the metals accumulate to a maximum concentration, after which regulation of these metals takes place (Francis, 1994). In the case of zinc, no definite seasonal trend could be established, a finding supported by data obtained by Van den Heever & Frey (1994) in a study performed in treated sewage water and natural dam water. Zinc concentrations were higher in the second year for both summer and autumn, which may be due to heavier rainfall in this period and consequent input from diffuse sources of pollution. The copper concentrations were, on the other hand, higher in the first year in the

4-38 Bioaccumulation of Zinc and Copper

summer and in the autumn, than the concentrations found in the second year, which may be due to the concentration of the metals due to the low flow of the water and effluent discharges from the mines, industries, informal settlements and combined sewage purification works. The copper concentrations appeared to be the lowest in autumn 1994 and the highest in the summer of 1994 and to some extent also in the winter of 1994 and the summer of 1995.

LOCALITY DIFFERENCES Significant differences (P5_0.05) between localities KOR1 and OR1 with regards to the bioaccumulation of zinc and copper in the different organs and tissues were recorded. A definite trend as to where the highest bioaccumulation had occurred was not always evident, but C. gariepinus accumulated more zinc in the liver at locality OR1 in February 1994. This could possibly be attributed to the absence of dilution, since the river experienced low flows during that time as a result of drier conditions. L. umbratus accumulated more zinc in the liver at locality KOR1, in February and May 1995, which coincides with the higher zinc concentrations in the water at locality KOR1 in those months (Chapter 3), due to the direct input of point source pollution from the combined sewage purification works located upstream from this locality and less rainfall and the low flow of the river, concentrating these pollutants. The lower zinc concentration at this locality for C. gariepinus can be attributed to the accumulation of zinc by the liver to the maximum concentration and the consequent regulation thereof by the liver. The higher zinc concentration in the liver was also found for C. gariepinus in treated sewage water in a study regarding the differences in bioaccumulation of zinc an copper between different tissue types in dam water and treated sewage water (Van den Heever & Frey, 1994). The copper skin and muscle concentrations were the highest for C. gariepintts at locality KOR1 in February 1994. In August 1994, the copper liver and muscle concentrations were the highest at locality OR1 and for

4-39 Bioaccumulation of Zinc and Copper

L. umbratus the highest at locality KOR1 in May 1995. There were, however, no significant differences between the two localities regarding the bioaccumulation of zinc and copper in the gills of both species.

4.5 CONCLUSION

There was a large variation in the zinc and copper concentrations in the different tissues/organs of the two species during the period February 1994 to May 1995, which suggests that the number of fish sampled at each locality should be increased, in order to minimize this variation and to be able to select fish of certain size ranges, as the results obtained from this study indicate that the accumulation patterns of metals in the muscle, skin, liver and gill tissues, vary as a function of the lengths of the fish, with higher concentrations in smaller fish.

The accumulation patterns also varied as a function of the species of the fish, mainly according to the different feeding habits of the two species and the routes of metal uptake by the fish, as well as the localities where the fish were collected. The higher concentrations at the different localities mainly coincided with the higher metal concentrations in the water at the time, with the treated sewage water being a major problem, especially at locality KOR1 during the drier periods and low flows. In order to obtain more detailed information with regards to the accumulation of zinc and copper in the different sexes, it is suggested that the gonads of both sexes should be examined, together with the different tissues/organs used in this study. However, the developmental stage of the gonads, should also be considered as this may greatly influence the data. For this reason, the use of the gonads as general monitoring tissues, is questionable. Zinc mainly accumulated in the gills, followed by the liver and the skin, with the lowest

4-40 Bioaccumulation of Zinc and Copper

concentrations found in the muscle. Very high concentrations of copper were found in the liver, which could indicate that the fish have been exposed to sublethal levels of the metal, especially in August and November 1994 and February and May 1995. It is suggested that the gills, liver, muscle and skin tissues should be used for the determination of zinc and copper bioaccumulation, as zinc and copper accumulated mainly in the gills and liver. In a general monitoring programme the muscle and skin tissue should also be included as fish are consumed by humans and in this case, by people from the informal settlements, as well as anglers at locality OR1. It is proposed that the fish be gutted and the gills removed before consumption. Additional studies need to be conducted on other species and locations in order to be able to describe the levels of these metals in fish. To understand the processes that regulates metal concentrations in different tissues and organs experimental exposures of juvenile and adult fish to specific concentrations should be performed.

The recommended concentrations for zinc in damwater is 5 mg/I (USEPA, 1976; WHO, 1984; SABS, 1984) and for copper the recommended concentration in dam water (USEPA, 1976) and treated waste water (South Africa, 1984) is 1 mg/l. The concentrations of zinc and copper in the river water at localities KOR1 and OR1 are well below this (see Chapter 3; Table 3-2), but it should be kept in mind that locality KOR1 is situated near the sewage works and an increase in the metal concentrations in the water can lead to higher accumulation of these metals by the fish, which can pose a health risk to the consumer.

4-41

Bioaccumulation of Zinc and Copper

4.6 REFERENCES

ABDULLAH, M.I.I., BANKS, J.W., MILES, D.L. & O'GRADY, K.I. (1976). Environmental dependence of manganese and zinc in the scales of Atlantic

salmon (Salmo salar L.). Freshwater Biology. 15:161 - 166.

BEZUIDENHOUT, L.M., SCHOONBEE, H.J. & DE WET L.P.D. (1990). Heavy metal content in organs of the African sharptooth catfish, Clarias gariepinus (Burchell), from a Transvaal lake affected by mine and industrial effluents. Part 1. Zinc and copper. Water SA. 16(2):125 - 129.

*BOWEN, H.J.M. (1972). The biochemistry of trace metals. In: Nuclear activation techniques in life science. Vienna. International Atomic Energy Agency. pp. 393 - 405.

BRYAN, G.W. (1976). Some aspects of heavy metal tolerance in aquatic organisms. In: Effects of pollutants on aquatic organisms. A.P.M. Lockwood [ed.]. Cambridge: Cambridge University Press. pp. 7 - 34.

CROSS, F.A., HARDY, L.H., HONES, N.Y. & BARBER, R.T. (1973). Relation between total body weight and concentrations of manganese, iron, copper, zinc and mercury in white muscle of Bluefish (Pomatomus saltatrix) and a Bathyl- bemersal fish Antimora rostrata. J. Fish. Res. Board Can. 30:1285 - 1291.

4-42 Bioaccumulation of Zinc and Copper

DE WET, L.M., SCHOONBEE, H.J., DE WET, L.P.D. & WIID, A.J.B. (1994). Bioaccumulation of metals by the southern mouthbrooder, Pseudocrenilabrus philander (Weber, 1897) from a mine polluted impoundment. Water SA. 20(2):119 - 126.

EDDY, F.B. (1981). Effects of stress on osmotic and ionic regulation in fish. In: Stress and Fish. A.D. Pickering [Ed.]. Academic Press, London. 367 pp.

FRANCIS, G.M. (1994). Toxic substances in the environment. John Wiley & Sons Inc. New York, Chichester, Brisbane, Toronto, Singapore. pp. 172 - 173.

GIESY, J.P. & WIENER, J.G. (1977). Frequency distributions of trace metal concentration in five freshwater fishes. Trans. Am. Fish. Soc. 106:393 - 403.

GRAY, H. & BERTINI, I. (1986). Progress in Inorganic Biochemistry and Biophysics, Vol. I, Zinc Enzymes. Birkhauser Boston Inc. Ed. Boston (USA).

HARRISON, F.L. & LAM, J.R. (1986). Copper-binding Proteins in Liver of Bluegills Exposed to Increased Soluble Copper under Field and Laboratory Conditions. Environ. Health Persp. 65:125 - 132.

HEATH, A.G. (1987). Water Pollution and Fish Physiology. CRC Press, Boca Raton, F.L. pp. 245.

HUGHES, G.M. (1970). A comparative approach to fish respiration. Experientia. 26:113 - 224.

4-43 Bioaccumulation of Zinc and Copper

HUGHES, G.M. & GRAY, I.E. (1972). Dimensions and ultrastructure of toadfish gills. Biol. Bull. Mar. Biol. Lab. Woods. 143:150 - 161.

JONES, J.R.E. (1938). The relative toxicity of salts of lead, zinc and copper to the stickleback (Gasterosteus aculeatus) and the effects of calcium on the toxicity of lead and zinc salts. J. Exp. Biol. 4:394 - 307.

*KODAMA, M., OGATA. T. & YAMAMORI, K. (1982). Haemolysis of erythrocytes of rainbow trout Salmo gairdneri exposed to zinc polluted water Bull. Jap. Soc. scient. fish. 43:593.

MATHIS, B.J. Sc. CUMMINGS, T.F. (1973). Selected metals in sediments , water and biota in the Illinois River. G. Wat. Pollut. Cont. Fed. 45:1573 - 1583.

MATTHIESSEN, P. & BRAFIELD, A. (19.77). Uptake and loss of dissolved zinc by the stickleback, Gasterosteus aculeatus. J. Fish. Biol. 10:399 - 410.

OLSON, K.R., SQUIBB, K.S. & COUSINS, R.J. (1978). Tissue uptake, sub-cellular distribution and metabolism of 114CH3HgC1 and CH3205HgC1 by rainbow trout. J. Fish. Res. Bd Can. 35:381 - 190.

O'NEILL, J.G. (1981). The humoral immune response of Salmo tnata L. and Cyprinus carpio L. exposed to heavy metals. J. Fish Biol. 19:297 - 306.

POURBAIX, M. (1966). Atlas of Electrochemica Equilibria. Pergamon Press Ed. Oxford (England).

4-44 Bioaccumulation of Zinc and Copper

SABS (SOUTH AFRICAN BUREAU OF STANDARDS) (1984). SABS 241 - 1984 Specification for Water for Domestic Use (2nd edition). The Council of the South African Bureau of Standards, Pretoria.

SEYMORE, T. (1994). Bioccumulation of metals in Barbus marequensis from the Olifants River, Kruger National Park and lethal levels of manganese to juvenile Oreochromis mossambicus. M.Sc. RAU.

SHEPARD, K. & SIMKISS, K. (1978). The effects of heavy metal ions on Ca' ATPase extracted from fish gills. Comp. Biochem. Physiol. 6113:69 - 72.

SKELTON, P.H. (1993). A complete guide to the freshwater fishes of Southern Africa. Southern Book Publishers, Halfway House. pp. 177 - 178 & 229 - 230.

SKIDMORE, J.F. (1970). Respiration and osmoregulation in rainbow-trout with gills damaged by zinc sulphate. J. Exp. Biol. 52:481 - 494.

SKIDMORE, J.F. & TOVELL, P.W.A. (1972). Toxic effects of zinc sulphate on the gills of rainbow trout. Water Research. 6:217 - 230.

SOLBe, J.F. (1973). The toxicity of zinc sulphate to Rainbow trout in very hard water. Water Research. 8:389 - 391.

SOUTH AFRICA (REPUBLIC) (1984). Vereistes vir die Suiwering van Afvalwater en Afloop. Regulasie 991 van 1984. Staatskoerant No. 9225. 18 Mei 1984. Staatsdrukker, Pretoria.

4-45 Bioaccumulation of Zinc and Copper

STAGG, R.M. & SHUTTLEWORTH, T.J. (1982). The accumulation of copper in Platichthys flesus L. and its effects on plasma electrolyte concentrations. J. Fish Biol. 20:491 - 500.

STEENKAMP,V.E., DU PREEZ, H.H. & SCHOONBEE, H.J. (1994). Bioaccu- mulation of copper in the tissues of Potamonautes warreni (Calman) (Crustacea, Decapoda, Branchiura), from industrial, mine and sewage-polluted freshwater

ecosystems. S. -Afr. Tydskr. Dierk. 29(2):152 - 161.

USEPA (UNITED STATES ENVIRONMENTAL PROTECTION AGENCY) (1976). Quality Criteria for Water. National Technical Information Service. PB-263 943. Office of Water and Hazardous Materials, Washington D.C. EPA. 26 July 1976).

VAN DEN HEEVER, D.J. & FREY, B.J. (1994). Human health aspects of the metals zinc and copper in tissue of the African sharptooth catfish, Clarias gariepinus, kept in treated sewage effluent and in the Krugersdrift Dam. Water SA. 20(3):205 - 212.

VAN DER WAAL, B.C.W. (1972). 'n Ondersoek na aspekte van die ekologie, teelt en produksie van Clarias gariepinus (Burchell) 1822. M.Sc. verhandeling, Randse Afrikaanse Universiteit, Johannesburg.

VAN LOON, J.C. (1980). Analytical Atomic Absorption Spectroscopy. Selected Methods. Academic Press, New York.

VAN RENSBURG, E.L. (1989). Die Biokonsentrering van atrasien, sink en yster in Tilapia sparrmanii (Cichlidae). M.Sc. Verhandeling, Randse Afrikaanse Univer- siteit, Suid Afrika.

4-46 Bioaccurnulation of Zinc and Copper

VARANASI, U. & MARKEY, D. (1978). Uptake and release of lead and cadmium in skin and mucus of coho salmon (Oncorhynchus kisutch). Comp. Biochem. Physiol. 60C:187 - 191.

VARIAN (1989). Flame Atomic Absorption Spectrometry: Analytical methods. Varian Techtron Pty Limited, Australia. pp. 146.

WHO (WORLD HEALTH ORGANIZATION) (1984). Guidelines for Drinking Water Quality. Vol I. Recommendations. World Health Organization, Geneva. 130 pp.

WIENER, J.G. & GIESY, J.P. (1979). Concentrations of Cd, Cu, Mn, Pb and Zn in fishes in a highly organic softwater pond. J. Fish. Res. Board Can. 35:270 - 279.

WINBERG, G.G. (1956). Rate of metabolism and food requirements of fish. Beloruss. State Univ., Minsk. Fish. Res. Bd Can. Trans. Ser. No 194, 1960.

YAMAMOTO, Y., ISHII, T. & IKEDA, S. (1978). Studies on copper metabolism in fishes III. Existence of metallothionein-like protein in carp hepatopancreas. Bull. Jap. Soc. Scient. Fish. 44:149 - 153.

* These articles were not viewed by the author

4-47 • Bioaccumulation of Zinc and Copper

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4-48 ▪ - ▪ Bioacctunidation of Zinc and Copper

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4-49 Bioaccurnulation of Zinc and Copper

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4-50 • ▪ • Bioaccumulation of Zinc and Copper

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4-51 Bioaccumulation of Zinc and Copper

4-52 Bioaccutnidation of Zinc and Copper

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;. CHAPTER 5

BIOACCUMULATION OF MANGANESE AND LEAD IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS AND LABEO UMBRATUS

5.1 INTRODUCTION

In addition to their natural occurrence, metals may enter and contaminate the environment as a result of anthropogenic activities. Metals are released into the environment particularly due to rapid industrialization, technology development, mining and agricultural activities. Due to the toxic properties of metals, their uncontrolled release in the environment pose serious water pollution problems. Metal toxicity to fish is often characterized by gill damage and an excessive secretion of mucus, with mortalities related to physiological disturbances to respiration, resulting in hypoxia (Neville, 1985) and ionoregulatory disturbances, resulting in body ion depletion (Neville, 1985). Because freshwaters are often sinks for metals and fish tend to accumulate them, they are potentially important sources of metals in the human diet.

Manganese (Mn) is one such toxic metal, which does not appear naturally in its pure state, but it is found in various salts and minerals, for example, manganese dioxide (pyrolusite) (Galvin, 1996), manganese carbonate (rhodocrosite) and manganese silicate (rhodonite), with the oxides being the only important manganese-containing minerals mined. Manganese is a vital micro-nutrient for both plants and animals, but is also primarily used in metal alloys, dry cell batteries, micro-nutrient fertilizer additives and the permanganates (7') are also very strong oxidizing agents of organic material. Manganese is normally present in its divalent state (Mn 2+) (Friberg, Nordberg & Vouk, 1986), with the most soluble manganese compounds being the divalent ones. Therefore

5-1 Bioaccumulation of Manganese and Lead

manganese levels in well-oxygenated waters are low and in these waters the oxidation of the divalent manganese occurs rapidly and therefore also the precipitation of non-soluble Mil' compounds (WHO, 1986). Manganese has considerable significance biologically and evidence has been presented that it is accumulated by certain species of fish (Uthe & Bligh, 1971). According to Kempster, Hattingh & Van Vliet (1982), manganese can be considered of moderate toxicity to aquatic life, although the liver glycogen and blood glucose levels can be altered by manganese in higher concentrations (Nath & Kumar, 1987).

Lead (Pb) is a toxic metal which is known to accumulate in the tissues/organs of fish. Lead accumulates mainly in the bone, scales, gill, kidney and liver of fish, with direct uptake of aqueous Pb 2+ across the gills, being the primary mode of uptake of lead in freshwater fish. Although not known to be essential to animal nutrition, lead can be . a potentially important trace element in mammals (Schwarz, 1972). In excessive concentrations, however, lead is very toxic to most plants and animals where it acts as a cumulative poison. Lead has a world-wide distribution in the environment and occurs in trace amounts in rocks, soil, water, air and in plants and animals. Most lead salts are of low solubility and lead exists in nature mainly as lead sulphide (galena). Other common natural forms are lead carbonate (cerussite), lead sulphate (anglesite) and lead chlorophosphate (pyromorphite). Stable complexes result from the interaction of lead with the sulfhydryl, carboxyl and amine coordination sites, characteristically found in living matter. Lead is widely used by man as an additive in fuels, paints, pesticide formulation and in the production of batteries and enters the aquatic environment through precipitation, lead-dust fallout, erosion and leaching from soil, municipal and industrial waste discharges and the runoff of fallout deposits from streets and other surfaces.

5-2 Bioaccumulation of Manganese and Lead

The aim of this section of the study, was to determine the extent of manganese and lead bioaccumulation in the different tissues/organs of the African sharptooth catfish, Clarias gariepinus and the moggel, Labeo umbratus as well as the difference between the accumulation of the different metals between the species and localities.

5.2 MATERIALS AND METHODS

Field and laboratory procedures for manganese and lead analyses of the fish from the Klein Olifants River (Locality KOR1) and the Olifants River (Locality OR1) (see Table 4-1), were the same as the procedures described in Chapter 4 for the analysis of zinc and copper. The statistical procedures were also the same as described in Chapter 4.

5.3 RESULTS

53.1 DIFFERENCES IN BIOACCUMULATION OF MANGANESE AND LEAD IN THE DIFFERENT

TIS SUES/ORGANS For C. gariepinus and L. umbratus, the highest concentrations of manganese in the different tissues/organs were clearly found in the gills, followed by the liver and then the skin and muscle (Fig 5-1). The differences between the different tissues/organs were mostly significant (P5_0.05), except between the liver & skin and muscle & skin for C. gariepinus at locality OR1 (Table 5-1) and the muscle & skin at locality KOR1 as well as locality OR1 for L. umbratus (Table 5-2). For lead, however, the order of

5-3 Manganese concentration (Ag/g) FIG 5-1 100 40 20 60 80 0 Feb 94 TISSUES/ORGANS OF THE MEANMANGANESECONCENTRATIONS (B) FROMFEBRUARY1994TOMAY1995 Nov 94 Feb 95 / tig Skin CLARL4S GARIEPINUS ❑ May 95 Liver 12MuscleMIGills Bioaccumulation ofManganeseandLead Manganese concentration (ttg/g) B 100 20 40 60 80 0 Feb 94 / AT LOCALITIESKOR1(A)ANDOR1 (pg/g May 94 DRY MASS)INTHEDIFFERENT / / -e7 Aug 94 Nov 94 ,n• Feb 95 L7 5-4 ,

May 95 0 ,

Bioaccumulation of Manganese and Lead

70 200

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0 0 1434 i.- 40

<1.) C.) 8 100 0 0 30 0) co) 0.) eu 00 OA C 20 ed 0

0 0 Aug 94 Nov 94 Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

Skin ❑ Liver EiMuscle U Gills

FIG 5-2 THE MEAN MANGANESE CONCENTRATIONS (liglg DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF LABEO UMBRATUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM AUGUST 1994 TO MAY 1995

5-5 Bioaccumulation of Manganese and Lead

TABLE 5-1 SUMMARY OF THE DIFFERENCES (SIGNIFICANT =P50.05; NON SIGNFICANT = P>0.05)) BETWEEN THE MANGANESE CONCENTRATIONS IN THE TISSUES AND ORGANS OF CLAMS GAREEPLVUS IN FEBRUARY 1994 (F2), MAY 1994 (M2), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

TABLE 5-2 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNFICANT = P>0.05) BETWEEN THE MANGANESE CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE0 UMBRATUS IN AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F 2) AND MAY 1995 (MO

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5-6 Bioaccumulation of Manganese and Lead

bioaccumulation in the different tissues/organs of the two species was not clear to distinguish, but is was evident that the highest concentrations occurred in the gills for C. gariepinus (Fig 5-3) and L. umbratus (Fig 5-4). The differences between the different tissues/organs of both species were significant between the gills and all the other organs as well as between the muscle & skin for C. gariepinus at locality KOR1 (Table 5-3) and between the liver & skin and the muscle & liver at locality OR1 for L. umbratus (Table 5-4).

The manganese and lead bioconcentration factors between the water and the tissues/organs (BFw) were much higher than the bioconcentration factors between the sediments and the tissues/organs (BF5). The manganese bioconcentration factors (Table 5-1 Appendix), ranged between 9.1 (in the muscle of C. gariepinus at locality OR1 in May 1994) and 11 800 (in the gills of L. umbratus in November 1994 at locality OR1), while the bioconcentration factor between the tissues/organs and the sediments fluctuated between 0.001 (as determined for the muscle of C. gariepinus in May 1994 at locality OR1 and November 1994 at locality KOR1) and 0.14 (in the gills of L. umbratus in August 1994 at locality OR1). For lead, the bioconcentration factors (Table 5-2 Appendix), between the water and the different tissues/organs (BF„,) of the two species, ranged between 9.5 (in the muscle of L. umbratus in May 1995 at locality KOR1) and 1 100 (in the gills of C. gariepinus in February 1995 at locality OR1). The BF, were in the range of 0.01 (in the skin, liver and muscle of L. umbratus in November 1994 at locality OR1) and 0.3 (in the gills of C. gariepinus in February 1994 at locality KOR1).

5.3.2 SPECIES DIFFERENCES Manganese and lead bioaccumulation in the different tissues and organs showed significant differences (P50.05) between the two species. Manganese concentrations in

5-7 Bioaccumulation of Manganese and Lead

0 30 /

■■•

25 40 LT

S = 30 0 = eS 1... 4 0 C 8 20 -cl a)cz -a

0 Feb 94 Nov 94 Feb 95 May 95 0 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

/ / El Skin ❑ Liver ❑ Muscle in Gills /

FIG 5-3 THE MEAN LEAD CONCENTRATIONS (pg/g DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF CLARL4S GARIEPINUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

5-8 Bioaccumulation of Manganese and Lead

0 35

Aug 94 Nov 94 Feb 95 May 95

r / 0 Skin ❑ Liver EiMuscle U Gills /

FIG 5-4 THE MEAN LEAD CONCENTRATIONS (ugig DRY MASS) IN THE DIFFERENT TISSUES/ORGANS OF LABE° UMBRATUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM AUGUST 1994 TO MAY 1995

5-9 Bioaccumulation of Manganese and Lead

TABLE 5-3 SUMMARY OF THE DIFFERENCES (SIGNIFICANT =P50.05; NON SIGNIFICANT = P>0.05)) BETWEEN THE LEAD CONCENTRATIONS IN THE TISSUES AND ORGANS OF CLARIAS GARIEFINUS IN FEBRUARY 1994 (F1), MAY 1994 (M1), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

TABLE 5-4 SUMMARY OF THE DIFFERENCES (SIGNFICANT = P50.05; NON SIGNFICANT = P>0.05)) BETWEEN THE LEAD CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE° IIMBRATIIS IN Aucus-r 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F 2) AND MAY 1995 (M2)

,,,,,,,,,,,,, n,,,,,,,,,,, nn,nn ,

5-10 Bioaccumulation of Manganese and Lead

the muscle, liver and gills of L. umbratus were mostly very high in comparison to that in those tissues/organs of C. gariepinus in May 1994 at locality OR1, November 1994 at locality KOR1, February 1995 at locality OR1 and May 1995 at locality OR1 (Fig 5-5). In May 1995 at locality OR1, however, the manganese muscle concentrations were higher for C. gariepinus than L. umbratus. Lead muscle and liver concentrations were higher in L. umbratus in May 1995 at locality OR1 (Fig 5-6), but C. gariepinus accumulated more lead in the liver (Fig 5-6), gills and skin (Fig 5-7) in November 1994 at localities KOR1 & OR1, February 1995 at locality OR1 and May 1995 at locality OR1 respectively.

5.3.3 RELATIONSHIP BETWEEN LENGTHS AND MANGANESE AND LEAD CONCENTRATIONS The bioaccumulation of manganese in the skin, liver, muscle and gills of the two species, showed significant negative correlations with the lengths of the fish, with correlation coefficients ranging between -0.03 and -0.27 and P-values 5 0.02. There was also a negative correlation between the lead concentration in the liver, skin and muscle and the lengths of the fish, but a significant positive correlation was found in the gills, with a correlation coefficient of 0.57.

5.3.4 DIFFERENCES BETWEEN MALES AND FEMALES No significant differences were found between the males and the females with regards to manganese and lead bioaccumulation in the different tissues/organs of C. gariepinus and L. umbratus during the period February 1994 to May 1995 at the two localities.

5.3.5 SEASONAL DIFFERENCES The only significant seasonal differences (P50.05) found for L. umbratus were between the winter

5-11 •

Bioaccumulation of Manganese and Lead

A 16 8 60

.-.14 eto ,_eg) 50 tab

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0 0 May 94 (OR1) Nov 94 (KOR1) Feb 95 (OR1 May 95 (OR1 May 95 (OR1) C 250

en -2 200

150 0 [11Clarias gariepinus ❑ Labeo umbratus

a) U 0 100

03 50

°May 94 (OR1) Nov 94 (OR1) Feb 95 OR1 May 95 (OR1) FIG 5-5 SIGNIFICANT DIFFERENCES (PS 0.05) BETWEEN CLARIAS GARIEPINUS AND LABE° UMBRATUS REGARDING THE BIOACCUMULATION OF MANGANESE IN THE MUSCLE (A), LIVER (B) AND GILLS (C) FROM MAY 1994 TO MAY 1995

5-12 • ▪ •

Bioaccumulation of Manganese and Lead

35 8

tal)

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0 Feb95 OR1) May 95 (OR1) May 95 (OR1)

Clarias gariepinus ❑ Labeo umbratus

FIG 5-6 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN CLARLAS GARIEPINUS AND LABE° UMBRATUS REGARDING THE BIOACCUMULATION OF LEAD IN THE MUSCLE (A) AND LIVER (B) FROM FEBRUARY 1995 TO MAY 1995

5-13 Bioaccumulation of Manganese and Lead B 60 10

50 bl 8 alb

Q y To 40 a) 6 c s

0 0 30

8 4 8 10:3 2 0 cis e0 V

i) 2 10

0 Nov 94 (KOR1) Nov 94 (OR1) Feb 95 (OR1) May 95 (OR1)

❑ Locality KOR1 El Locality OR1 /

FIG 5-7 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN CLARIAS GARIEPINUS AND LABE° UMBRATUS REGARDING THE BIOACCUMULATION OF LEAD IN THE GILLS (A) AND SKIN TISSUE (B) FROM NOVEMBER 1994 TO MAY 1995

5-14 Bioaccumulation of Manganese and Lead

and spring of 1994 with regards to the manganese and lead bioaccumulation in the different tissues/organs. For C. gariepinus, significant seasonal differences in the bioaccumulation of manganese were found, but not always in the same tissues/organs (Table 5-5). The summer of the first year differed significantly (P__0.05) from the other seasons with respect to the liver, skin and gills. The autumn of 1994 differed significantly from the summer and spring of the same year and this was indicated by the metal concentrations in the liver and skin respectively. Significant differences were also found between autumn 1994 and summer/autumn 1995 in the muscle, skin, liver and gills of C. gariepinus. The skin, muscle, liver and gill manganese concentrations differed significantly between the winter and spring of the first year and winter 1994 and summer/autumn 1995. These differences were mainly indicated by the concentrations in the skin, followed by that in the liver and gills and least in the muscle. The highest manganese concentrations were found in the summer months of the two years, followed by the autumn and spring seasons.

The lead concentrations in the different tissues/organs of the two species differed significantly, especially between the autumn and summer of the first year as well as between the autumn of the first year and the summer of the second year (Table 5-6). No significant differences were found however, between the summer and winter of 1994 and the autumn and spring of the same year. The highest concentrations were found in the spring, followed by the summer and autumn.

5.3.6 LOCALITIES D114 ERENCES No significant differences (P>0.05) were found for the two localities during May and November 1994 with regard to manganese and lead concentrations in the different tissues/organs of the two species. In the case of manganese, localities KOR1 and OR1 differed significantly from each other in February 1994, with respect to the muscle and liver concentrations in L. umbratus (Fig 5-8). In August 1994, the manganese

5-15 Bioaccumulation of Manganese and Lead

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5-16 Bioaccumulation of Manganese and Lead

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5-17 Bioaccumulation of Manganese and Lead

A 16 60

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0 Feb 1994 Feb 1995 Feb 1994 Feb 1995

I=1 Locality KOR1 El Locality OR1 /

FIG 5-8 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF MANGANESE IN THE MUSCLE (A) AND LIVER (B) IN FEBRUARY 1994 IN LABE° UMBRATUS AND IN FEBRUARY 1995 IN CLAR1AS GARIEPINUS

5-18 Bioaccumulation of Manganese and Lead

concentrations in the skin and gills of L. umbratus differed significantly (P..0.05) with higher concentrations at locality OR1 (Fig 5-9), while the concentrations of manganese in the muscle, liver (Fig 5-8), skin and gills (Fig 5-9) were higher at locality OR1 in February 1995. In May 1995, the two localities showed significant differences for the manganese concentration in the gills of C. gariepinus, with higher concentrations at locality OR1. For lead, there were significant locality differences with regard to the metal concentration in the skin and muscle (Fig 5-10) and the gills (Fig 5-11) of L. umbratus in February 1994 with higher values at locality KOR1 and in May 1995, where the highest concentrations were found at locality OR1 in C. gariepinus. The highest lead concentrations in the liver and gill tissue were found in February and May 1995 at locality OR1 in C. gariepinus (Fig 5-11).

5.4 DISCUSSION

Greater than natural concentrations of manganese and lead in the air, soils and surface waters have been reported in urban areas, especially those which are highly industrialized. Metals released into surface water, tend to accumulate in the sediments through adsorption and precipitation processes, but can be reintroduced into the water in a bio-available form to fish and other aquatic biota, with changing water quality conditions, for example the pH and ionic strength. Manganese and lead concentrations found in the two species examined for the purpose of this study, were variable, with sample standard deviations often as large as the sample mean itself. Giesy & Wiener (1977) found that metal concentrations in fishes usually exhibit positive skewness and are frequently non-normal, a finding which is supported by the data in this study.

5-19 Bioaccumulation of Manganese and Lead

B 30 250

s 200

r4, V 20

150 • C O es

15 4a. V 0 V 8 8 100 V 10 eiD E

V 5 _ 41 50 7

Aug 1994 Feb 1995 Aug 1994 Feb 1995 May 1995

❑ Locality KOR1 0 Locality OR1

FIG 5-9 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN LOCALITTES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF MANGANESE IN THE SKIN (A) AND GILLS (B) IN AUGUST 1994 AND FEBRUARY AND MAY 1995 IN CLAMS GARIEPINUS

5-20

Bioaccumulation of Manganese and Lead

30 25

25 t i 20 4) = = 0 7,1 15

..4-.

8 co 10 0 C

l' )

Feb 1994 May 1995 Feb 1994 May 1995

■ ❑ Locality KOR1 El Locality OR1 /

FIG 5-10 SIGNIFICANT DIFFERENCES (P:50.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF LEAD IN THE SKIN (A) AND MUSCLE (B) IN FEBRUARY 1994 IN LABE° UMBRATUS AND MAY 1995 IN CLARIAS GARIEPINUS

5-21 Bioaccumulation of Manganese and Lead

35 30

30 25 to

= i 8 -aco 10 a) es=

Feb 1995 May 1995 Feb 1994 May 1995

/ ❑ Locality KOR1 ifi Locality OR1

)

FIG 5-11 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF LEAD IN THE LIVER (A) AND GILLS (B) IN FEBRUARY 1994 IN LABEO UMBRATUS AND FEBRUARY AND MAY 1995 IN CLARL4S GARIEPINUS

5-22 Bioaccumulation of Manganese and Lead

The body tissue manganese concentrations obtained in this study, ranged from 1 to 272 Ag/g dry mass. It was found that the different tissues and organs of C. gariepinus and L. umbratus bioaccumulate different levels of manganese. It has been shown that manganese and lead can be taken up indirectly from food and ingested sediments via the gut, or directly through the gills (Bendell-Young & Harvey, 1986). The highest manganese concentrations were detected in the gills of L. umbratus, which were the main route of uptake of manganese, as little absorption of this metal occurs through the gut from the food (Katz, LeGore, Weitkamp, Cummins, Anderson & May, 1972). Lead was also noted for its specific accumulation in gills of the fish (also noted by Brooks & Rumsey, 1974), which is important, as they are in close contact with the external environment. The mean lead concentrations found in the gills of the two species, were higher than concentrations of this metal found in the gills of B. marequensis in the Lower Olifants River Catchment (Seymore, Du Preez & Van Vuren, 1995), but lower than concentrations found by Du Preez & Steyn (1992) in the Olifants River, Kruger National Park (assuming a moisture content of 80 %). Other tissues in C. gariepinus and L. umbratus, also accumulated manganese and lead, but to a much lesser degree than the gills. The concentrations of manganese recorded from the muscle tissue of C. gariepinus from treated sewage water (25 µg/g) and natural dam water (18 pg/g) (Van den Heever & Frey, 1996), were generally higher than manganese concentrations in the muscle tissue of C. gariepinus during this study (0.4 - 10 aug/g dry mass, with an exceptionally high concentration of 26 Agig dry mass at locality OR1 in May 1995), also exposed to effluent from sewage purification works. The lowest amount of manganese and lead was concentrated by the muscle and skin tissue, which is very important, as it is the edible part of the fish. It has been pointed out in this study that freshwaters and river sediments are often sinks for metals (Chapter 3) and as fish tend to accumulate metals from water and sediment, they are therefore potentially important sources of metals in the human diet. Lower concentrations in these tissues can also indicate that the skin is an important excretory organ for these metals (Khalaf, Al-Jafery, Khalid, Elis

5-23 Bioaccumulation of Manganese and Lead

& Ishaq, 1985), presumably by means of the mucus (Heath, 1987). The skin, together with the gills, is characterized by a mucus layer on the outer surface, which can indicate them as excretion routes, involving the sloughing off of mucus from these surfaces (Varanasi & Markey, 1978).

In August 1994 and February 1995 at locality KOR1 and in May 1995 at locality OR1 the mean lead concentrations in the muscle tissue for both species, ranging between 9 ,ug/g and 12 ti,g/g dry mass, exceeded the limit for human consumption (8 ;Leg dry mass, assuming a 75 % moisture content; 2/4/g wet mass) (Brown, Gardiner & Yates, 1984). It is therefore proposed that additional research should be done regarding this issue, as the fish at locality OR1 are used for consumption by humans. This is important, for manganese, as an essential trace element, is involved in haemoglobin production as well as flavo proteins and is also known as an enzymatic catalyst. Excessive ingestion of manganese is not carcinogenic, but ingestion of manganese rich waters (up to 14 mg/1) can cause brain illness, while manganese deficiency in the body, is associated with anaemia, bone malformation in children and heart disease (Galvin, 1996). Since manganese is regarded as a metal with relatively low toxicity, acute poisoning in humans is very rare. In the case of lead, a disease known as saturnism, characterized by lead accumulation in the bones and especially the nervous, renal and hepatic systems, causes anaemia, paralysis, headaches and enzymatic and lipid tissue alterations (Galvin, 1996). Lead can also be considered as a carcinogenic element (Leed, 1972) and is accumulated in the human body and excreted very slowly. For this reason, lead can also be passed on from the mother to the foetus through the placenta and causes difficulty in oxygen exchange in the cells, as it reacts with sulphur amino acids, causing mental retardation (Galvin, 1996).

The liver seems to accumulate the second highest concentration of manganese and lead. The highest manganese concentration in the liver of C. gariepinus, were found in November 1994 at locality OR1, receiving effluent from point (e.g. combined sewage

5-24 Bioaccumulation of Manganese and Lead

purification works) and diffuse sources of pollution, but this value was much higher than the values obtained in treated sewage effluent during December 1990, in the Bloem- spruit Sewage Purification Plant (Van den Heever & Frey, 1996). Appreciable levels of these metals build up in the liver to a maximum concentration, whereafter it is regulated by the liver, where metals bind to metallothionein (metal binding protein) and can be stored or regulated. These metaloproteins have many sulfhydryl groups as a result of large amounts of cysteine in the molecules and these sulthydryl groups bind the metals, making them less toxic to other cellular constituents (Heath, 1987). Despite the high individual variations, the general relationship in manganese and lead concentrations among the different tissues and organs was: G>

The bioconcentration factor can be seen as a constant of proportionality between the concentration of the chemical in the fish and the concentration in the water/sediment. Manganese bioconcentration factors were much higher than those for lead, a finding supported by a study in the lower catchment of the Olifants River (Seymore et aL, 1995), but for both manganese and lead, the factors were the highest in the gills. The bioconcentration factors calculated in this study for manganese, were much higher than those calculated in the lower parts of the Olifants River, while the bioconcentration factors for lead were lower. A high degree of bioavailability of manganese to fish is suggested by the high water bioconcentration factors, especially in November 1994, which can be due to the various physico-chemical properties of the water (Heath, 1987), as the concentrations of manganese in the water were not necessarily higher at that time (see Chapter 3, Table 3-3). It must be noted that only one water sample was collected once every three months, stressing the importance of more regular monitoring of the system. The lower bioconcentration factors for the sediment and the fish organs (BF s), indicated that less manganese and lead were available for the fish, from the sediment.

In the case of lead, the bioconcentration factors were relatively low, except in February of the second year, at locality OR1, where much higher bioconcentration factors were

5-25 Bioaccumulation of Manganese and Lead

found, which also coincided with higher lead concentrations in the water (Chapter 3; Table 3-2). The bioconcentration factors between the tissues/organs of the two species and the sediment, were much lower than those calculated for the water and tissues/organs. Although the BF„, calculated during this study, were higher than those calculated by Seymore et aL (1995), the BF, were much lower. The water of the Upper Olifants River Catchment is however fairly hard, contributing to the lower bioavailability of the metals from the sediment (see Chapter 4 for description).

Manganese and lead bioaccumulation in the different tissues/organs of the two species, varied significantly in a number of cases. Manganese accumulation were much higher in L. umbratus in May 1994 at locality OR1, November 1994 at locality KOR1 and February and May 1995 at locality OR1, than in C. gariepinus. Lead accumulation in the muscle tissue was also much higher in L. umbratus in May 1994 at locality OR1. C. gariepinus, however, accumulated more lead in the gills, liver and skin than L. umbratus. Once again, these differences can be contributed to the feeding habits of the two species, as manganese and lead can be taken up from the food and ingested sediments (Bendell-Young, et al., 1986). Results from this study indicate that the sediments contain high concentrations of manganese (See Chapter 3). Thus bottom feeding fish, such as L. umbratus are continually ingesting manganese contaminated sediments and detritus while feeding. C. gariepinus had the lowest levels of manganese, which may be related to the minimal contact they have with the stream bottom while feeding and swimming as they are predators and do not ingest large amounts of sediment or detritus while feeding. The gills of L. umbratus generally had higher manganese but lower lead concentrations when compared to the gills of C. gariepinus. This can possibly be related to the functional difference of the gills of C. gariepinus, as it is an air-breather and not so dependant on the gills for respiration. The role of the gills of C. gariepinus in the absorption and excretion of metals can be different, resulting in differences in metal uptake and excretion. This, however, still needs more detailed investigation.

5-26 Bioaccumulation of Manganese and Lead

Examination of the extent of bioaccumulation of manganese and lead in the muscle, skin, liver and gills of the two species, revealed a strong significant (P 0.05) negative correlation with the lengths of the fish, with the exception of lead bioaccumulation in the gills, which showed a significant positive correlation with the size of the fish. It is thus apparent that, predominantly, the accumulation of manganese and lead decreases with increasing fish size and the smaller the fish, the higher the bioaccumulation of these metals. Negative correlations between the metal concentrations in the skin and muscle and the lengths of the fish, can be related to the incorporation of new tissues, thus, dilution by growth. The higher metabolic rate of smaller fish can also be a dependant factor, as more oxygen is needed and thereby a higher rate of water flow over the gills is established (Winberg, 1956).

SEASONAL DIFFERENCES The highest concentrations of manganese in the different tissues and organs of C. gariepinus and L. umbratus were found in the summer of 1994, which may be ascribed to higher rainfall, causing more pronounced leaching and the manganese concentrations in the tissues and organs of the two species were the lowest in autumn and winter of 1994. The manganese concentrations in the water were not necessarily higher in this season and the pH of the water also showed no significant decrease. The significant higher manganese concentrations in the summer, compared to the autumn and winter, could possibly be attributed to the water temperature (decrease of 14.5 °C) in the winter. The much higher temperature in summer 1994, causes a higher activity and therefore ventilation rate, as increasing temperature lowers the oxygen affinity of the blood and therefore the rate of accumulation also increases. A higher metabolism may lead to more frequent feeding, which will also result in higher metal concentrations if the metals are biomagnified via the food.

5-27 Bioaccumulation of Manganese and Lead

LOCALITY DIFFERENCES No definite trend could be established as to where the highest bioaccumulation took place between the two localities, but it seems as if manganese bioaccumulation was the highest at locality OR1 in the skin, liver, gills and muscle, with the exceptions in February 1994 for C. gariepinus, where the highest manganese bioaccumulation occurred at locality KOR1 in the muscle and liver. This coincides with the higher manganese BFw at locality KOR1 at that time. The high manganese concentrations may be attributed to industrial activities in the catchment area as well as to effluents from combined sewage purification works located upstream. The manganese and lead concentrations in the skin, liver, muscle and gill tissues of C. gariepinus, collected from the Germiston lake, which had been polluted by mining and industrial effluent (Du Preez, 1995), were lower than the concentrations recorded for C. gariepinus during this study, which indicates more severe manganese and lead pollution in the area investigated. The lower concentrations in the fish from Germiston Lake can also be due to increased exposure to these metals, resulting in over regulating of the levels of manganese and lead. Nevertheless, the data presented indicate that the fish are exposed to sublethal levels of manganese and lead, resulting in the observed concentrations. The possible point and diffuse sources of manganese and lead in the upper catchment of the Olifants River, therefore needs further investigation. In the case of lead, the bioaccumulation in the skin, liver and muscle of L. umbratus was the highest at locality OR1, receiving effluent from mines and sewage purification works and surface runoff from industries, such as paint factories. Once again C. gariepinus accumulated more lead at locality KOR1 in February 1994, when the lead concentration in the water was also much higher.

5-28 Bioaccumulation of Manganese and Lead

5.5 CONCLUSION

According to the data, the gills and liver tissues accumulated the highest concentrations of manganese and lead, while the muscle and skin accumulated the lowest concentrations. The manganese concentrations recorded during this study in the muscle tissue of C. gariepinus and L. umbratus (4 Agig for both species), pose no health risk to man, according to the average daily manganese intake of approximately 10 mg (Sollman, 1957), but the lead concentrations in the muscle tissue of both species, exceeded the limit for normal daily intake of the metal. The organs suggested to sample for manganese and lead analyses in fish are the gills, liver and muscle tissue (to test its fitness for human consumption). The accumulation also seems to vary as a function of the species of fish, but the accumulation of manganese and lead is not dependant on the sex of the fish. Larger sample sizes are needed to minimize variation and to be able to select fish of a certain size, as a definite relationship exists between the size of the fish and the bioaccumulation of manganese and lead. This was not possible during this study and the feasability of this in many river systems is, however, questionable. The present data indicate that the fish have been exposed to chronic sublethal concentrations of these metals, but no serious manganese pollution problem is suggested.

5-29 Bioaccumulation of Manganese and Lead

5.6 REFERENCES

BENDELL-YOUNG, L.I. & HARVEY, H.H. (1986). Uptake and tissue distribution of manganese in the white sucker (Catostomus commersoni) under conditions of low pH. Hydrobiologia. 113:117 - 125.

BROOKS, R.P. & RUMSEY, D. (1974). Heavy metals in Baltic herring and cod. N.Z.J. Mar. Freshwater Res. 8:155 - 166.

BROWN, V.M., GARDINER, J. & YATES, J. (1984). Proposed environmental Quality Standards for List 2 Substances in Water. WRC Tech. Mem. 208.

DU PREEZ, H.H. & STEYN, G.J. (1992). A preliminary investigation of the concentration of selected metals in the tissues of the tigerfish, (Hydrocynus vittatus) from the Olifants River, Kruger National Park, South Africa. Water SA. 18(2):131 - 136.

DU PREEZ, H.H. (1995). Ekologiese bestuursaspekte van die Germistonmeer, Gauteng. Verslag deur Departement Dierkunde en Navorsingseenheid vir Akwatiese en Terrestriele Ekosisteme, Randse Afrikaanse Universiteit. pp 45, 66.

FRIBERG, L., NORDBERG, G.F. & VOUK, V.B. (1986). Handbook on the toxicology of metals. Vol. II. Specific metals. (2nd edn). Elsevier Science Publishers, B.V. Amsterdam, 704 pp.

GALVIN, R.M. (1996). Occurrence of metals in water: An overview. Water SA 22(1):7 - 18.

5-30 Bioaccumulation of Manganese and Lead

GIESY, J.P. & WIENER, J.G. (1977). Frequency distributions of trace metal concentrations in five freshwater fishes. Trans. Am. Fish. Soc. 106:393 - 403.

HEATH, A.G. (1987). Water Pollution and Fish Physiology. CRC Press, Inc. Boca Raton. Florida. 245 pp.

KATZ, M., LEGORE, R.S., WEITKAMP, D., CUMMINS, J.M., ANDERSON, D. & MAY, D.R. (1972). Effects on freshwater fish. J. WPCF. 44(6):1226 - 1250.

KHALAF, A.N., AL-JAFERY, A., KHALID, B.Y., ELIAS, S.S. & ISHAQ, M.W. (1985). The patterns of accumulation of some heavy metals in Barbus grypus from a polluted river. JBSR 16(2):51 - 74.

LEED, D.H.K. (1972). Metallic Contaminants and Human Health. Academic Press Ed. New York (USA).

NATH, K. & KUMAR, N. (1987). Toxicity of manganese and its impact on some aspects of carbohydrate metabolism of a freshwater teleost, Colisa fasciatus. Sci. Total Environ. 67:257 - 262.

NEVILLE, C.M. (1985). Physiological responses of juvenile rainbow trout, Salmo gairdneri, to acid and aluminium - prediction of field responses from laboratory data. Can. J. Fish. and Aquat. ScL 42:2004 - 2019.

SCHWARZ, K. (1972). The role of trace elements in health and disease process in man and animals, with special reference to elements newly identified as essential. In: Symposium on nuclear activation techniques in the life sciences. I.A.E.A. Ljubljana Yugoslavia, 10 - 14 April.

5-31 Bioaccumulation of Manganese and Lead

UTHE, J.F. & BLIGH, E.G. (1971). Preliminary survey of heavy metal contamination of Canadian freshwater. J. Fish. Res. Bd Can. 28:786 - 788.

VAN DEN HEEVER, D.J. & FREY, B.J. (1996). Human health aspects of certain metals in the tissue of the African sharptooth catfish, Clarias gariepinus, kept in treated sewage effluent and the Krugersdrift Dam: Iron and Manganese. Water SA. 22(1):67 - 72.

VARANASI, U. & MARKEY, D. (1978). Uptake and release of lead and cadmium in skin and mucus of coho salmon (Oncorhynchus kisutch). Comp. Biochem. Physiol. 60C:187 - 191.

* WHO (WORLD HEALTH ORGANIZATION) (1986). Directives de Qualite pour l'eau Boisson, Vol I, II y III. Geneve (Switzerland).

WINBERG, G.G. (1956). Rate of metabolism and food requirements of fish. Beloruss State Univ. Minsk. Fish. Res. Bd Can. Trans. Ser No 194, 1960.

* This article was not reviewed by the author

5-32 - p ▪ Bioaccumulation of Manganese and Lead

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(-4 esi

f74 0 0 0 0

5-34 • - ▪ ▪ ▪ ▪ Bioaccumulation of Manganese and Lead

N N 0, 00 N• 1 • " t•-• pM +1 " a CO CV _ Ce. C,‘ +1 ' t•-• ' +1 ' 0, '"C" +1 V.1 +1 ' a < +1 CO § vp aMvla en en 0 q 0 en c..4 -• VZ, co m •cr 1■4 OS CO 0 eg NJ •-• eV 0 en Cn M en 0 r, CO N N Z N

O Co co VI mcsi0 c`i M `0 el C`i < +1 o m +1 ' v1Qo +1 ' +1 ' +1 0 eNi 0 M v O e‘i ++ esc O en c-J O N Co NI 4. C-4 V' est -+

sel N er +1 co (") +1 un ,0 C.4 e.4 +I `r +1 q en ' cq 0 +1 ' b 8. +1 M4 ' esi NI VI 0 • -■ e.4 .er en 0 en CO V' N O N sr, en 'Tr VI

N 6■1 Co m -• es4 +1 0 "I' On q q 0 q +1 q ' m +1 ' O.; C-4 el 0 •-• so a C• 1 VI NI 'ce 0 en sO en VI 0 N N N ess Z est esi

N +1 g +1 g +1 g eE c E c E 4 'a . 0: a • . 'al `at e

e•I N NUED) TI

CON ble ( O Nn Nr O O O O ila a av t APPENDIX No

5-35 ▪ • Bioaccumulation of Manganese and Lead

t-- eq Tr tr) esi r„. vt >0 In ,r cv +I +1 Cq +1 ,.., tn • Cr -er • +1 CI 0 '0 0 " ■ %el Ir. 0 e0 v. v. •••• s0 .tr >el 0 r4 N 0 eV C.- 0 0

c, c0 m eV Ch rn m •-• c0 44 • 71• fq 't +1 • c? +1 ' c) 41 •',=? I4 • CO 00 VD 0 r4 e•44 0 —4 w-a (NI es so m cz, esi v-4 est •O ton eV 0 00

esi .„J• ,0 rst " +1 cs1 c0 est ND C.4 VI> +1 • C> 0 0 q 0 p +1 or CI 0 +1 co q at) q 0 vl '0 0 eV .. Tr ul -• 0 esi Un en 0 eV CO tn 0'0 00'0 m 0 CO

r4 0 ."‘ rn eV so rn " " +1 es/ "I 0 +1 ' un 0 +1 • +1 q +1 Cr. +1 00 sr> vD e4 un Tr r4 m t--4 00 Tr m 0 e. '0 00 eV 0'c VD Tr e9 000

V) +1 +1 g c E E 1.„! VQ" c co e

1.1 11 e4

X 0 0 0 0 NDI APPE 2 5- BLE TA

5-36 ▪ Bioaccumulation of Manganese and Lead

00 co rn v") N "tr r- r". N M '0 00 +1 +I ' +1 +1 o crs +1 C' le- VS. --: N O r‘l vl o N en -1 0 -• •-• ••••■ .•-■ 00 0 en tn c> — O 0■ SO 0 CV

'0 N N m •-• N N h 0 Ch + 1 ' %el O +1 O +I ' Q +1 '?. N •-• Vr C, m M•-N■ o en •-•O N 00 'Tr 1+1 0 N

m er O N v., en ev a of cs +1 ' +1 ' +1 ' esi O ' o +1 ' co + 1 ' o CO me en 0 •-• m N 0 en en N o N 0 N Os CO en 0 ee CO .ce en 0 N

• (.1 ,_, co m -• +I +1 ' C? +1 • O ' MQ o M en .-, R •-• N 00 of en 0 N

vi ili vi vi +I g +I g +I g +I g 0 ._- E . 0e --aE. • . =0 --Ea . • . e E"a. • INcoco e ±.'.' LCt LC t e i) 101 5 e i) i t LC t C

N N N

0 U oG "" 0 0 z W a. iNi

5-37 Bioaccumulation of Manganese and Lead

le b ila a v a t No

A - N/

5-38 ;c c,' " I •

rc:jc • -:•••;"

;'-;tr t. ...I, Y." .c ;1;,-, • , d'," • .= . " :•.'"e

•

• CHAPTER 6

THE BIOACCUMULATION OF CHROMIUM AND NICKEL IN THE TISSUES AND ORGANS OF CLAMS GARIEPINUS AND LABE° UMBRATUS

6.1 INTRODUCTION

Chromium is the 17th most abundant nongaseous element in the earth's crust (Schroeder, 1970), with concentrations ranging between 80 and 200 mg/kg with an average of 125 mg/kg (NAS, 1974b). In natural waters, the oxidation states of chromium range between Cr' and Cr', with Cr' considered as a minor problem, while Cr' is very toxic (Krenkel, 1974). The trivalent form of chromium is found most commonly in nature, and the biological activity of chromium, i.e. its effects as an essential metal, is restricted to this form of the metal. It forms complexes that are stable at or below pH 4, but which hydrolyse at high pH values, resulting in the formation of polynucleate bridge complexes. The trivalent form of chromium exists in large, insoluble macromolecules which precipitate and become biologically inert. For this reason, it must be supplied as a complex of suitable stability in order to be utilized (Train, 1979).

In natural waters, chromium usually occurs in low concentrations, ranging between 1 - 2 µf,/1 (Moore & Ramamoorthy, 1984; Snodgrass, 1980). The main anthropogenic sources of chromium are industrial effluents emanating from the production of corrosion inhibitors and pigments (Galvin, 1996), which can increase its levels in the water to being harmful to aquatic organisms. The toxicological action of chromium is unique compared

6-1 Bioaccumulation of Chromium and Nickel

to other metals, in that it has to be in the hexavalent form to cross biological membranes (Gray & Sterling, 1950). Doudoroff & Katz (1953), suggested that Cr(VI) behaves toxicologically different from most metals and it exists almost exclusively in the aquatic environment in the form of oxo-anions (Cr0 42-, HCrO4 Cr2072-) which have been observed to pass readily through the gill membrane and to accumulate in various tissues and organs (Knoll & Fromm, 1960). Once it is inside the biological system, it can also be reduced to the trivalent form (Mertz, 1969). The toxicological action of chromium therefore results from both the oxidizing action of Cr(VI) and the formation of complexes with various organic compounds (Mertz, 1969). Although fish appear to be relatively tolerant of chromium, hexavalent chromium is irritating and corrosive to the mucous membranes and can elicit its toxic action internally as well as on the gill surface (Buhler, Stokes & Caldwell, 1977). Chromium toxicity also varies with chemical species, chromium oxidation state and water pH (Train, 1979) as the distribution of Cr(VI)-species is altered by a change in the pH (Trama & Benoit, 1960) and therefore the pH influences the availability of Cr(VI) to fish.

Nickel is a natural ubiquitous element of the earth and its waters, but activities such as mining, melting processes of metallic alloys, as a catalyst or as part of pesticide forMulations (Galvin, 1996) can result in nickel emission in the air and water. Nickel can also be introduced in considerable quantities by fossil fuel combustion, into the atmosphere and much of this can eventually find its way into the aquatic environment. Nickel salts, except for the ferri- and ferrocyanides, sulphide and Ni-dimethylglyoxime, are reasonably soluble in water. Nickel can be found in oxidation states 1 to 4' but the 2+ state is most commonly found, with speciation and concentration dependant on the forming of complexes, adsorption, precipitation, or chelating with dissolved inorganic and organic ligands (Moore et al., 1984). Existing nickel complexes include hydrochloric, sulphuric (very stable), amine (fairly stable), oxalic, thiocyanide, metaphosphoric, pirophosphoric and cyanide complexes (Pourbaix, 1966). Reports on levels of nickel in freshwater fish, indicate a content which fluctuates between ± 10 and 120 µg/kg

6-2 Bioaccumulation of Chromium and Nickel

(Tong, 1974; Vos & Hovens, 1986), with concentrations in the water varying considerably, usually being below 5pg/1 in non-polluted environments, but concentrations can be considerably higher in urban areas (Bencko, 1983).

In this section of the study, the extent and order of bioaccumulation of chromium and nickel were determined in the different tissues and organs of Clarias gariepinus and Labeo umbratus, as well as the dependence of bioaccumulation upon the species of fish, the localities and the seasons.

6.2 MATERIALS AND METHODS

Field, laboratory and statistical procedures for chromium and nickel analyses on samples from fish collected from the Klein Olifants River (KOR1) and the Olifants River (locality OR1) (see Table 4-1), were the same as the procedures described in Chapter 4 for the analyses of zinc and copper.

6.3 RESULTS

6.3.1 BIOACCUMULATION OF CHROMIUM AND NICKEL IN THE DIFFERENT TISSUES/ORGANS The bioaccumulation pattern of chromium and nickel was not clear, but it seems as if predominantly higher concentrations of these metals were found in the gills (G), followed by the liver (L), muscle (M) and skin (S) of C. gariepinus and L. umbratus. The highest chromium (Fig 6-1; Table 6-1 Appendix) and nickel (Fig 6-3; Table 6-2 Appendix) concentrations, up to 374 Ag/g dry mass and 71 nig dry mass respectively in the gills of

6-3 Bioaccumulation of Chromium and Nickel

C. gariepinus were found in November 1994. Significant differences (P50.05) in the bioaccumulation of chromium and nickel were found between the gills and all the other tissues/organs at locality KOR1, as well as between the liver & skin and the muscle & liver at locality OR1 for C. gariepinus (Table 6-1), with the exception of nickel, where no significant difference (P >0.05) was found between the muscle and liver tissues (Table 6-3). Chromium (Fig 6-2) and nickel (Fig 6-4) concentrations were the highest in both the gills and the liver in L. umbratus, with the only significant differences (P5_0.05) between the different tissues/organs, occurring in August 1994 for the gills & skin and the gills & liver for chromium (Table 6-2) and between the gills and all the other tissues/organs for nickel (Table 6-4).

The calculated bioconcentration factors between the water and the tissues/organs (BF„) were as expected higher than the bioconcentration factors between the sediments and the tissues/organs (BF,). Bioconcentration factors (BF,,,) for chromium ranged between 19 (in the skin of C. gariepinus in February 1995 at locality OR1) and 722 (in the gills of C. gariepinus in November 1994 at locality KOR1) with very high values in November 1994 and February/May 1995 at both localities (Table 6-1 Appendix). Bioconcentration factors between the sediment and the tissues/organs ranged between 0.2 (calculated in the skin, liver and muscle of C. gariepinus in May 1994 at locality OR1) and 2.3 (calculated in the gills of C. gariepinus in November 1994 and May 1995 at locality KOR1).

The bioconcentration factors for the water and the tissues/organs of the fish (BF„,) for nickel, fluctuated between 39 (in the skin of both species in February 1995 at both localities) and 2 367 (in the gills of C. gariepinus in November 1994 at locality KOR1) (Table 6-2 Appendix). Bioconcentration factors between the sediments and the tissues/organs of the fish (BF,), ranged from 0.1 (calculated for the skin and muscle of L. umbratus at locality KOR1 in November 1994) and 0.5 (calculated for the gills and

6-4

Bioaccumulation of Chromium and Nickel

140 80

oo 60

4a. 0a) 40 8 E E 0 U 20

0 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

rig Skin ❑ Liver El Muscle $ Gills

FIG 6-1 THE MEAN CHROMIUM CONCENTRATIONS (1.1,g1g DRY MASS) IN IHE DIFFERENT TISSUES /ORGANS OF CLARL4S GARIEPINUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

6-5 Bioaccumulation of Chromium and Nickel

70 0

70 60

CO 60 4e 50 = 0 0 50 74 "',...4 40 I-. ..a. a.) a. v 8 40 = g 30 8 E O 5= 30

•. oE' ..d ,.- 20 P. U U z0

10 10

0 0 Aug 94 Nov 94 Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

/ / 3Skin ❑ Liver El Muscle EGills /

FIG 6-2 THE MEAN CHROMIUM CONCENTRATIONS (ueg DRY MASS) IN 1 HE DIFFERENT TISSUES /ORGANS OF LABE° UMBRATUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM AUGUST 1994 TO MAY 1995

6-6 Bioaccumulation of Chromium and Nickel

TABLE 6-1 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE CHROMIUM CONCENTRATIONS IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS IN FEBRUARY 1994 (F,), MAY 1994 (M,), AUGUST 1994 (A), NOVEMBER 1994

(N), FEBRUARY 1995 (F2 )

Skin liver Musele: Gills L.ocAirrY KOR1

Skin. n.s. n.s. F„ N, F2

Liver ns. F„ N, F2

:Muscle. F„ N

Lociarry OR1

F, n.s. F, , M„ F2

Li F,

Muscle F„ A, N, F2

Gills

TABLE 6-2 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE CHROMIUM CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE0 UMBRATUS IN AUGUST 1994 (A)

Skirl iiv Muscle Gills

LOCALITY KOR1

Skin n.s. n.s. A

Liv n.s. ns. Muscle n.s.

Loc.Aury OR1

Skin n.s. n.s. n.s.

n.s. A

Muscle n.s.

6-7 Bioaccumulation of Chromium and Nickel

80 50

0 30 4c-ct. 8

8 20 zC.)

0 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

I3Skin ❑ Liver 0Muscle I•Gills

FIG 6-3 THE MEAN NICKEL CONCENTRATIONS (Agig DRY MASS) IN nit, DIFFERENT TISSUES /ORGANS OF CLARL4S GARIEPINUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

6-8 Bioaccumulation of Chromium and Nickel

40 60

35 50

30 bA 40

•-0 C73 20 "a. 30 0 U

73). 15 0 z z 10

0 0 Aug 94 Nov 94 Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

Skin OLiver Li Muscle Gills

FIG 6-4 THE MEAN NICKEL CONCENTRATIONS (lgig DRY MASS) IN rriE DIFFERENT TISSUES /ORGANS OF LABEO UMBRATUS AT LOCALITIES KOR1 (A) AND OR1 (B) FROM AUGUST 1994 TO MAY 1995

6-9 Bioaccumulation of Chromium and Nickel

TABLE 6-3 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P5.0.05; NON SIGNIFICANT = P>0.05) BETWEEN THE NICKEL CONCENTRATIONS IN THE TISSUES AND ORGANS OF CL4RL4S GARTEPINUS IN FEBRUARY 1994 (F1), MAY 1994 (M1), Aucusr 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2)

Ski Liver . -Muscle •

LocAuTY KOR1

Skin n.s. n.s. F, , N, F2

'Liver n.s. F„ N

Muscle F„ N

LOCALITY OR1

n.s. F„ M„ A, N, F2

n.s. , M1 , A

MnScia F„ A

Gills

TABLE 6-4 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE NICKEL CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE0 UMBRATUS IN AUGUST 1994 (A)

Skin aver Gills

LOC.ALITY KOR1

. Skin n.s. n.s. A

n.s. A

1■4 A Gills

LocAurY OR1

n.s. n.s. A

ns. A

Mu A

Gills ..

6-10 Bioaccumulation of Chromium and Nickel

liver of both species in February 1994 and February/May 1995 at both localities).

6.3.2 SPECIES DIFFERENCES The only significant differences (P5_0.05) found between C. gariepinus and L. umbratus, regarding the bioaccumulation of chromium in the different tissues/organs, were in the gills, with the highest concentrations found in C. gariepinus in November 1994 at both localities (Fig 6-5). In the case of nickel, L. umbratus accumulated significantly (P50.05) more of the metal in the skin, liver and muscle in November 1994 at locality OR1 (Fig 6-6) than C. gariepinus, which accumulated more nickel in the gills in November 1994 at both localities (Fig 6-5).

6.3.3 RELATIONSHIPS BETWEEN THE LENGTHS AND THE CHROMIUM AND NICKEL

CONCENTRATIONS IN THE TISSUES/ORGANS In determining the relationship between the bioaccumulation of chromium and nickel in the different tissues/organs of C. gariepinus and L. umbratus and the lengths of these fish, predominantly significant (P50.05) negative correlations were found. The concentrations of chromium and nickel in the gills, however, showed significant (P50.05) positive correlations, as did the nickel concentrations in the muscle tissues.

6.3.4 DIFFERENCES BETWEEN MALES AND FEMALES The only significant difference found between males and females, was for nickel in the muscle of C. gariepinus in May 1995 at locality OR1 with the highest concentrations detected in the females.

6-11

Bioaccumulation of Chromium and Nickel

A B 250 120

--a 200 cn TAO

,4 ' 150 O O 1-.4 60 . . . . . w. . . . .

. . . . . 0 0 100 '." E 8 . . . . . . . . . . . . . . . . 1.. 3 40 * i . . . . . . . 0 O ...... ...... . . . . . . . . cd C. • . . . . . ...... . . . . . . . . . . ...... v, 50 ...... . . . . . . . . . . . . 20 •

. . . . . . . . ...... . . . . . ...... ......

. . •. •. • . . •. •. •. • . . . . . . . . 0 0 Nov 94 (KOR1) Nov 94 (OR1) Nov 94 (KOR1)

Clarias gariepinus ❑ Labeo umbratus

FIG 6-5 SIGNIFICANT DI/14 ERENCES (1'40.05) BETWEEN CLARLAS GARTEPEVUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF CHROMIUM (A) AND NICKEL (B) IN THE GILLS IN NOVEMBER 1994 AT BOTH LOCALITIES

6-12 • Bioaccumulation of Chromium and Nickel

A 3° 3

25 25 gi) dA

a) .9) 20 20 a>

0 0 73 15 15 *4.

U 8 8 10 .) 1 0 U U

ct Cd 4.)

Nov 94 (OR1) Nov 94 (OR1) C 35

7) 25 E

LI Clarias gariepinus ❑ Labe° umbratus . 5, 20

0 Cd

15 8

..s4 10

eti

0 Nov 94 (OR1) FIG 6-6 SIGNIFICANT DIFFERENCES (P 0.05) BETWEEN CLAMS GARIEPINUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF NICKEL IN THE SKIN (A), LIVER (B) AND MUSCLE (C) TISSUES IN NOVEMBER 1994 AT LOCALITY OR1

6-13 Bioaccumulation of Chromium and Nickel

63.5 SEASONAL DIFFERENCES No distinguished trend could be established with regards to the mean chromium and nickel concentrations in the various tissues/organs of C. gariepinus and L. umbratus concerning seasonal differences. The highest concentrations of chromium and nickel were found in the summer and autumn of 1995, with the chromium concentrations differing significantly (P_O.05) from all the other seasons in the summer and autumn of 1994. In the case of chromium (Table 6-5), the summer and autumn of 1994, differed significantly from all the other seasons. No significant differences (P>0.05) were, however, found in comparisons between spring/winter 1994 and summer/autumn 1994. In the case of nickel (Table 6-6), there were no significant differences between summer/autumn 1995 and summer/spring 1994 and between summer 1995 and autumn 1995, nor between spring 1994 and autumn/winter 1994. Strong significant differences were however found in all the various tissues/organs between autumn 1994 and summer/autumn 1995 as well as between winter 1994 and autumn 1995. The highest concentrations of chromium and nickel were found in the summer and autumn of the second year.

63.6 LOCALITIES DIFFERENCES The chromium (Fig 6-7; Fig 6-8) and nickel (Fig 6-9; Fig 6-10) concentrations were mostly in the same range at both localities. Significant differences (P 0.05) between the two localities, were found in February 1995 at locality OR1 for L. umbratus, while concentrations of both chromium and nickel were the highest at locality KOR1 in August 1994 and February 1994 respectively for C. gariepinus.

6-14 Bioaccumulation of Chromium and Nickel

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6-15 Bioaccumulation of Chromium and Nickel

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6-16

Bioaccumulation of Chromium and Nickel

80 100

80 0 cn 60 cnC.) = e3.0 0 g cl) = =

..... 60 0 0 Q • 0 irct •••1 c. l'I .5 40 61 4) '44) U 0 U C 40 E E . 1 E 2 ..z 0

cti 20

c'')

Feb 1995 Feb 1995

EILocality KOR1 El Locality OR1 /

FIG 6-7 SIGNIFICANT DIFFERENCES (P5.0.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF CHROMIUM IN THE GILLS (A) AND MUSCLE (B) IN FEBRUARY 1995 FOR LABE() UMBRATUS

6-17

Bioaccumulation of Chromium and Nickel

B 100 120

) CID 100 ig 8 (n k in s 80 he t in

60 ion at tr en c

con 40 ium om hr c

20 20 Mean

0 Feb 1995 Aug 1994 Feb 1995

■ ❑ Locality KOR1 El Locality OR1

FIG 6-8 SIGNIFICANT DIFFERENCES (13 0.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF CHROMIUM IN THE SKIN (A) AND LIVER (B) IN AUGUST 1994 FOR CLARIAS GARIEPINUS and FEBRUARY 1995 FOR LABE° UMBRATUS

6-18

Bioaccumulation of Chromium and Nickel

70 60

60 50 ,..i.a 4

C.> -'-.) 50 ig 7:7.0 a) O 40 a) ,.= 0 0 40 ,..,O 0 1; .—0 I... 4:3 30 'E 0 t> G 30 0 8 C a7) ,.W 75 20 -.0 -NC 0 20 .....c.) 0 0 co 0 co ''') '-) 10 10

Feb 1995 Feb 1994 Feb 1995

❑ Locality KOR1 El Locality OR1 /

FIG 6-9 SIGNIFICANT DIFFERENCES (P5.0.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF NICKEL IN THE GILLS (A) AND MUSCLE (B) IN FEBRUARY 1994 FOR CLARIAS GARIEPINUS AND AND IN FEBRUARY 1995 FOR LABE° UMBRATUS

6-19

Bioaccumulation of Chromium and Nickel

60 70

60 50 10 e3

411 50

4 0 0 30 re:w.

c..) 30 8 8 731 4) ..s4 20 U 0 20

10 10

Feb 1995 Feb 1995

❑ Locality KOR1 0 Locality OR1

FIG 6-10 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF NICKEL IN THE SKIN (A) AND LIVER (B) IN FEBRUARY 1995 FOR LABEO UMBRATUS

6-20 Bioaccumulation of Chromium and Nickel

6.4 DISCUSSION

Metal contamination of aquatic ecosystems has long been recognized as a serious pollution problem. Metal concentrations in fish have been related to morphometry and pH (Wren & MacCrimmon, 1983), alkalinity (Sheider, Jeffries & Dillon, 1979), dissolved organic matter (Wiener & Giesy, 1979), trophic relationships of fish (Rodgers & Qadri, 1982) as well as differences among species and fish mass within populations (Wren, MacCrimmon & Loescher, 1983). It is also clear from this study that the uptake of metals in fish tissues/organs and the potential use of fish in environmental assessment programmes, depend on factors such as species of the fish, fish tissues/organs, as well as modes of metal uptake and release and environmental factors.

C. gariepinus and L. umbratus showed differential accumulation of chromium and nickel in their tissues and organs. Chromium and nickel were found to have accumulated in all the tissues/organs of the two species, although the present data indicate that the gills contain the highest levels of both chromium and nickel. It is reasonable to assume that the uptake of chromium and nickel via the gills, takes place as a result of its intimate blood-water contact and the fact that the gills are the main site for absorption of toxicants from the surrounding medium but, it should also be remembered that the gills play an important role in the excretion of pollutants such as metals, probably via the mucus (Heath, 1987). Higher chromium concentrations would probably have been found in the gills, if the pH of the water at the two localities, had been more acidic, as chromium toxicity seems to be pH dependant, with increasing toxicity as pH decreases (Wepener, 1992). This finding is supported by a study performed by Van den Heever & Frey (1996), which showed that the lower pH and total hardness of the water (pH 7.4 and total hardness of 119.2 mg/1), contributed to higher concentrations of chromium in the water. Predominantly, the second highest concentrations of chromium and nickel were found in the liver of the two species. These high levels in the liver are not

6-21 Bioaccumulation of Chromium and Nickel

surprising, as the liver is associated with storage and detoxification functions. It is therefore apparent that the concentrations of chromium and nickel in the gills and to some extent the liver, can be used to detect chronic chromium and nickel poisoning. The lowest chromium and nickel concentrations were found in the skin and muscle tissues, also found by Seymore (1994), in a study performed on the bioaccumulation of chromium and nickel in the yellow fish, Barbus marequensis from the Lower Olifants River Catchment. The skin and muscle tissues are the least preferred site for chromium and nickel storage, which is important, as fish are consumed by humans. The estimated intake of chromium should average at -± 150 p.g chromium per day (Snyder, Cook, Nasset, Karhausen, Howels & Tripton, 1975). In this study, the average chromium concentration in the muscle tissue is 28p,g/g thy mass for C. gariepinus and 30 pl,/g dry mass for L. umbratus and these therefore pose no serious health risk to the consumer. These findings are in contrast to values obtained by Van den Heever & Frey (1996) for C. gariepinus in treated sewage water and natural dam water, with values of 157.3 p,g/g wet mass (768.5 p,g/g thy mass with moisture content assumed to be 80 %) in the treated sewage water and 151.9 pg/g wet mass (759.5 ,u,g/g dry mass with 80 % moisture content) in the natural dam water. Chromium is an essential trace element, acting as an insulin co-factor in glucose metabolism and also increasing the protein and amino acid uptake by cells. The trivalent form of chromium needs to be ingested in excessive amounts to cause toxicological disruptions in humans, while lesser amounts of Cr' may promote skin alterations, pulmonary problems as well as cancer (Galvin, 1996), but its mutagenic activity can be decreased by reducing agents such as human gastric juice (WHO, 1993).

Although certain nickel salts can have potentially carcinogenic and mutagenic effects, which can be explained by the reaction between the metal and the cellular DNA, nickel is generally not very toxic (USEPA, 1977). In animal experiments, ± 90 % of the ingested nickel was recovered in the faeces, meaning that ± 10 % might have been

6-22 Bioaccumulation of Chromium and Nickel

absorbed. Urinary excretion is also a major route for the elimination of nickel from the body and high ingestion can cause renal problems and also skin allergies by contact (WHO, 1990e).

The bioconcentration factors calculated for the water and the skin, muscle, liver and gill tissues of C. gariepinus and L. umbratus, provide some indication of the bio-availability of chromium and nickel to the fish, from the water. The highest bioconcentration factor for chromium, was found in the gills of C. gariepinus and the lowest in the skin of C. gariepinus. High bioconcentration factors for the water and tissues/organs of the fish, were found in May 1995 at locality KOR1, which coincided with high chromium concentrations in the water at that time. In the case of nickel, the bioconcentration factors (BFw) showed the same trend, with higher values in the gills and lower values in the skin of C, gariepinus. High bioconcentration factors (BFw) for nickel, were also found in November 1994 and May 1995 at both localities, with lower nickel concentrations in the water at both localities in November 1994, while the concentrations were higher in the water in May 1995 at both localities. The bioconcentration factors, ranging between 19 - 722 for chromium were much lower than bioconcentration factors calculated by Seymore (1994) in a study performed on the lower catchment of the Olifants River, while bioconcentration factors for nickel, ranging between 39 - 2367, were much higher than the values obtained in that study (3 - 1090), which may be ascribed to the metals forming complexes with the carbonates in the water, thus becoming less available for uptake by the fish.

The bioconcentration factors between the sediment and fish tissue/organs for both chromium and nickel, were much lower than the bioconcentration factors for the water and the skin, muscle, liver and gills of both species. This indicates that only a small fraction of these metals in the sediment is actually available for accumulation by the fish.

6-23 Bioaccumulation of Chromium and Nickel

These values are also mainly in the same range, although slightly lower than the values obtained by Seymore (1994), ranging between 0.001 - 3.45 for chromium and 0.01 - 1.69 for nickel.

The bioaccumulation of chromium and nickel is also dependant on the species of fish. In contrast to zinc (Chapter 4; Fig 4-5), C. gariepinus accumulated more chromium and nickel in the gills, whereas L. umbratus accumulated these metals in higher concentrations in the skin, liver and muscle tissues. The higher concentrations of the metals in the gills of C. gariepinus, may be as a result of weaker excretion of the metals by the gills of C. gariepinus, or the deposition of the metals in the gills, as C. gariepinus does not use its gills as actively as other fish. The differences in the order of bioaccumulation of chromium and nickel between the two species are apparent and can probably also be attributable to the feeding habits of the two species, as the uptake of these metals via the food is quite important. L. umbratus feeds by sucking up the surface layer of the sediment. Such diets, and particularly the ingestion of sediments which could quite conceivably be enriched with metals (see Chapter 3) as a result of industrial activity, would probably lead to the ingestion of greater quantities of pollutants than are ingested by fish with other feeding habits. Invertebrates accumulate higher levels of metals, than do fish (Heath, 1987) thus, predators of these invertebrates, being L. umbratus in this case, may obtain a considerable body burden from them. This phenomenon might explain the comparatively high values for chromium and nickel in the skin, muscle and liver of L. umbratus.

The accumulation of chromium and nickel in the different tissues/organs of the two species, did not differ much between the males and females of each species. The only significant difference (1'. 0.05) found regarding the accumulation between males and females, was for nickel, where the concentration of this metal was higher in the muscle tissue of the females. This supports findings obtained in a study performed on the

6-24 Bioaccumulation of Chromium and Nickel

tigerfish, Hydrocynus vittatus, in the Olifants River, Kruger National Park, with nickel concentrations in the muscle and gill tissues being slightly higher in the males than in the females (Du Preez & Steyn, 1992). Thus, the bioaccumulation of chromium and nickel for the selected tissues was predominantly independent of the sex of the fish. The use of the gonads of both sexes of the fish is suggested, in order to obtain information regarding the bioaccumulation of chromium and nickel in the testes of the males and the ovaries of the females of the two species, as explained in Chapter 4.

The correlations between the bioaccumulation of chromium and nickel in the different tissues/organs of the two species and the size of the fish, were predominantly negative. This can indicate a decrease in the concentration of chromium and nickel with an increase in fish size, with dilution by growth, changes in proximal composition of the muscle or decreased dietary intake, contributing to the negative correlation between the bioaccumulation of these metals in the muscle tissue and the size of the fish. Smaller fish are reported to take up metals via the food more rapidly than larger ones, which can be expected as they have higher rates of metabolism (per gram of body tissue) (Patrick & Loulit, 1978) and the higher ventilation rate of smaller fish, can also be the reason for the. higher metal concentration in the tissues and organs of smaller fish (Described in Chapter 4).

SEASONAL VARIATION Seasonal variation is greatest in the tissues which exhibit rapid uptake of chromium and nickel (gills and liver). No definite trend could, however, be established concerning the bioaccumulation of chromium and nickel between the different seasons. This was also the case in a study performed by Van den Heever & Frey (1996) concerning the bioaccumulation of chromium in different tissues/organs of C. gariepinus in treated sewage water and natural dam water. The lowest concentrations of these metals were

6-25 Bioaccumulation of Chromium and Nickel

found in autumn and winter of 1994, compared to the other seasons. A significant rise in bioaccumulation of chromium and nickel was observed in the summer and autumn of 1995. These significant differences (I):50.05) are correlated to the different concentrations of these metals in the water, especially at locality OR1. The summer of 1994/95 showed significantly higher concentrations of chromium and nickel in the tissues and organs of the two species, compared to the other seasons, which may be due to the higher water temperature at that time (see seasonal variation, Chapter 5).

LOCALITY DIFFERENCES The increase of the chromium and nickel concentrations in the tissues and organs of C. gariepinus and L. umbratus and in the water at locality OR1, may be due to the different industries in the Witbank area and effluent received from the Witbank Dam and the Suurstroom and Spook Spruit, both of which receive effluent from mines in their areas and flowing into the Olifants River before it reaches this locality. The metal concentrations were not necessarily higher in the water of locality OR1 in February 1995, thus the significant differences in the bioaccumulation of these metals, could be attributed to the physical/chemical properties of the water at that time. It should be noted that only one water sample was taken at each locality every third month, making precise comparisons difficult. The chromium concentrations in the skin (11 - 59 ;Leg dry mass), liver (10 - 68 pglg dry mass), muscle (11 - 56 irg/g dry mass) and gill tissues (17 - 130 p.g/g dry mass), were mostly higher than the concentrations in these tissues obtained by Seymore (1994) for B. marequensis in the lower catchment of the Olifants River, ranging between 1 - 37 p,g/g dry mass in the skin, 14 - 33 krg/g dry mass in the liver, 0.7 - 38 /Leg dry mass in the muscle and 3 - 104 ;Leg dry mass in the gills, pointing to more serious contamination by greater input from point and diffuse souces of pollution in the Upper Catchment of the Olifants River.

In the case of nickel, the concentrations in the different tissues/organs of C. gariepinus

6-26 Bioaccumulation of Chromium and Nickel

and L. umbratus, ranged between 8 - 42 p,g/g dry mass, 9 - 48 I.LgIg dry mass, 8 - 31 Ag/g dry mass and 13 - 71 p.g/g dry mass in the skin, liver, muscle and gill tissues respectively and were also much higher than the values for the lower Olifants River catchment (Seymore, 1994). These high concentrations in the tissues/organs of the two species, imply point and/or diffuse sources of these metals in the area and more intense investigations regarding this is suggested.

6.5 CONCLUSION

From the results of this section of the study, it is suggested that the gills, due to their intimate contact with the environment and the importance as an effector of osmotic and ionic regulation and the liver, as a storage/detoxification organ, may be the primary organs for chromium and nickel bioaccumulation and therefore, together with the muscle (evaluation for human consumption), should be used for the analysis of chromium and nickel to determine their concentrations in fish. It was also evident that the bioaccumulation of these metals was species specific, with higher concentrations in L. umbratus in the muscle, skin and liver tissues which may be ascribed to different trophic levels of the two species. The higher concentrations of these metals in the gills of C. gariepinus may be attributed to the possible differences in the role of metal regulation by the gills of the two species. The bioaccumulation of these metals is also dependant on the body size, with decreasing metal concentrations with increasing fish size. Larger sample sizes are therefore suggested, in order to be able to select a certain size range and thereby reducing variability. This objective may, however, be influenced by sampling difficulties and capturing success experienced during the present study. As the concentrations of metals such as chromium and nickel vary in fish tissues/organs as a funtion of species and size, the interspecific comparisons for these metals should be

6-27 Bioaccumulation of Chromium and Nickel

made in specific species and size ranges, as smaller fish tend to accumulate higher concentrations of these metals than the larger fish. The bioaccumulation of chromium and nickel in the muscle, liver, skin and gill tissues was not dependant on the sex of the fish, but it is suggested that the gonads of both sexes be included for this investigation. The bioaccumulation of chromium and nickel also differed significantly at the different localities, with locality OR1 in the Olifants River, receiving effluent from different industries and mines, showing the highest concentrations of these metals. Concentrations of chromium and nickel in the tissues/organs of C. gariepinus and L. umbratus are fit for human consumption, but more research is needed to investigate the origin of the point and diffuse sources of pollution.

6-28 Bioaccumulation of Chromium and Nickel

6.6 REFERENCES

BENCKO, V. (1983). Nickel: a review of its occupational and environmental toxicology. J. Hyg. Epidemiol. Microbiol. Immun. 27:237 - 247.

BUHLER, D.R., STOKES, R.M. & CALDWELL, R.S. (1977). Tissue accumulation and enzymatic effects of hexavalent chromium in rainbow trout (Salmo gairdneri). J. Fish. Res. Bd. Can. 34:9 - 18.

DOUDORFF, P. & KATZ, M. (1953). Critical review of literature on the toxicity of industrial wastes and their components to fish. II The metals as salts. Sewage Ind. Wastes. 25:802 - 839

GALVIN, R.M. (1996). Occurrence of metals in waters: An overview. Water SA 22(1):7 - 18.

GRAY, S.J. & STERLING, K. (1950). The tagging of red cells and plasma proteins with radioactive chromium. J. Clin. Invest. 29:1604 - 1613.

HEATH, A.G. (1987). Water Pollution and Fish Physiology. CRC Press, Inc. Boca Raton. Florida. 245 pp.

6-29 Bioaccumulation of Chromium and Nickel

KNOLL, J. & FROMM, P.O. (1960). Accumulation and elimination of hexavalent chromium in rainbow trout. Physiol. Zool. 33:1 - 8.

KRENKEL, P.A. (1974). Sources and classification of water pollutants. In: Industrial pollution. N.I. Sax [ed.]. Van Nostrand Reinhold, New York. pp. 197 - 219.

MERTZ, W. (1969). Chromium occurrence and funtion in biological systems. Physiol. Rev. 49:163 - 239.

MOORE, J.W. & RAMAMOORTHY, S. (1984). Heavy Metals in Natural Waters: Applied Monitoring and Impact Assessment. Springer-Verlag New York Inc., New York. 268 pp.

NATIONAL ACADEMY OF SCIENCES. (1974b). Chromium. U.S. government Printing Office, Washington, D.C.

PATRICK, F.M. & LOUTIT,M.W. (1978). Passage of metals to freshwater fish from their foods. Water Res., 12:395

POURBAIX, M. (1966). Atlas of Electrochemica Equilibria. Pergamon Press Ed. Oxford (England).

RODGERS, D.W. & QADRI, S.V. (1982). Growth and mercury accumulation in yearling yellow perch, Perca flavescens, in the Ottaawa River, Ontario. Environ. Biol. Fishes. 7:377 - 383.

6-30 Bioaccumulation of Chromium and Nickel

SCHEIDER, W.A., JEFFRIES, D.S. & DILLON, P.J. (1979). Effects of acidic precipitation on precambrian freshwater in Southern Ontario. J. Great Lakes Res. 5:45 - 51.

*SCHROEDER, H.A. (1970). Chromium. Air quality Monograph No 70-15. American Petroleum Institute, Washington D.C.

SEYMORE, T. (1994). Bioaccumulation of metals in Barbus marequensis from the Olifants River, Kruger National Park and lethal levels of manganese to juvenile Oreochromis mossambicus. M.Sc. RAU.

SNODGRASS, W.F. (1980). Distribution and behaviour of nickel in the aquatic environment. In: Nickel in the environment. J.O. Nriagu [ed.] Wiley, New York. pp. 203 - 274.

SNYDER, W.S., COOK, M.J., NASSET, E.S., KARHAUSEN, L.R. , HOWELS, G.P. & TIPTON, I.H. (1975). International Commission on adiological Protection. Report on the Task Group on Reference Man. ICRP Publication 23. New York.

TRAIN, R.E. (1979). Quality criteria for water. U.S. Envrionmental Protection Agency, Washington D.C. Castle House Publications Ltd.

TONG, S.S.C. (1974). Trace metals in lace Cayuga lake trout (Salvelinus namaycush). in relation to age. J. Fish. Res. Bd. Can. 31:238 - 239.

TRAMA, F.B. & BENOIT, R.J. (1960). Toxicity of hexavalent chromium to bluegills. J. Water Pollut. Control. Fed. 32:868 - 877.

6-31 Bioaccumulation of Chromium and Nickel

USEPA (US ENVIRONMENTAL PROTECTION AGENCY) (1977). Toxicology of Metals, Vol. II. Washington DC (USA).

VAN DEN HEEVER, D.J. & FREY, B.J. (1996). Human health aspects of certain metals in tissue of the African sharptooth ccatfish, Clarias gariepinus, kept in treated sewage effluent and the Krugersdrift Dam: Chromium and mercury. Water SA. 22 (1): 73 - 78.

VOS, G. & HOVENS, J.P.C. (1986). Chromium, Nickel, Copper, Zinc, Arsenic, Selenium, Cadmium, Mercury and Lead in Dutch fishery products (1977 - 1984). Sci. Total Environ. 55:25 - 40.

WEPENER, V., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1992). Die effek van heksavalente chroom by 'n varierende water-pH op die koolhidraatmetabolisme van Tilapia sparrmanii (Cichlidae). SA Tydskrif vir Natuurwetenskap en Tegnologie. 11(3):102 - 105.

WIENER, J.G. & GIESY, J.P. (1979). Concentrations of Cd, Cu, Mn, Pb and Zn in fishes in a highly organic softwater pond. J. Fish. Res. Board Can. 36:270 - 279.

WHO (WORLD HEALTH ORGANIZATION) (1990e). Environmental Criteria Health Series (no 108). Nickel. Geneve (Switzerland)

WHO (WORLD HEALTH ORGANIZATION) (1993). Guidelines for Drinking water Quality (2nd edn.), Vol. I. Geneve (Switzerland).

6-32 Bioaccumulation of Chromium and Nickel

WREN, C.D. & MacCRIMMON, H.R. (1983). Mercury levels in the sunfish, Lepomis gibbosus, relative to pH and other environmental variables of Precambrian Shield lakes. Can. J. Fish. Aquas. Sci. 40:1737 - 1744.

WREN, D.C., MacCRIMMON, H.R. & LOESCHER, B.R. (1983). Examination of bioaccumulation and biomagnification of metals in a Precambrian Shield lake. Water Air Soil Pollut. 19:277 - 291.

* This article was not viewed by the author

6-33 Bioaccumulation of Chromium and Nickel

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BIOACCUMULATION OF ALUMINIUM AND IRON IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS AND LABEO UMBRATUS

7.1 INTRODUCTION

Aluminium is abundant in the earth's crust, but its natural occurrence is limited to highly insoluble complex minerals. In aqueous solutions, aluminium can form a variety of complexes with water, hydroxide, fluoride, silicate and sulphate and the free aluminium ions are rare. Aluminium transformation processes in soil and surface water systems, occur continuously. These include processes like the mobilization of aluminium from slightly soluble sources into more soluble and reactive forms, or vice versa. Solubility, hydrolysis and molecular weight distribution of aluminium species, are affected by temperature as well as the pH of the solutions (Lydersen, Salbu, Pole() & Muniz, 1990b). The metal may be present in different physico-chemical forms, varying in molecular size and charge properties (Lydersen, Bjornstad & Englund, 1987), of which low molecular weight inorganic forms of aluminium, referred to as inorganic monomeric aluminium, are believed to be the most toxic species to freshwater fish in acidified waters (Lydersen, Poleo, Muniz, Salbu & Bjornstad, 1990a).

Losses of fish and other organisms from natural waters have been correlated with the presence of aluminium, mobilised from soils by acidic storms or snowmelt events (Harriman & Morrison, 1982). Water in regions impacted by acidic precipitation often exhibits elevated aluminium concentrations as a result of aluminium mobilization from sediments (Wright & Gjessing, 1976). The different forms of aluminium exhibit vastly

7-1 Bioaccumulation of Aluminium and Iron

different structures and activities and produce different effects on living organisms (Neville, 1985). The mechanism of aluminium toxicity to fish is thought to involve either ionoregulatory or respiratory disturbances, or both, dependant on the water pH and aluminium concentration (Neville, 1985). The precipitation of Al0H 3 and/or the binding of aluminium to organic anions on the gill surface, induce an inflammatory response. It stimulates mucus production, thickens the branchial epithelium, decreases transcellular permeability for 0 2 and CO2 and at the same time increases the permeability of paracellular channels through which the greatest amount of electrolyte loss occurs (McDonald, 1983). Significant accumulation of aluminium on the gills has also been observed with no increase in plasma aluminium levels (Youson & Neville, 1987).

Iron is the fourth most abundant by mass of the elements that make up the earth's crust. This metal is usually found in many soils, especially clay soils and may be present in varying quantities in water, with criteria of 1 mg/1 for freshwater aquatic life and 0.3 mg/I for domestic water supply (Train, 1979). These concentrations are dependant upon the geology of the area and other chemical components of the waterway. Iron can be found in different forms, including the ferrous, or bivalent (Fe') and ferric, or trivalent (Fe3+) irons, which are the primary forms of concern in the aquatic environment. The most important iron pollutant sources are mine drainage waste, industrial waste waters and iron bearing groundwaters. Iron, in the water from mine drainage, is precipitated as hydroxide (Fe(OH 3) in the presence of dissolved oxygen. This causes yellowish precipitates found in many streams draining coal mining regions. The smothering effects of settled iron precipitates may be particularly detrimental to fish eggs and bottom dwelling fish food organisms. The acute toxic action of iron seems at least partly to be related to metal precipitating on the gills, which leads to death by suffocation (Muniz & Leivestad, 1980). Recently (May 1996) the release of mine water containing high iron levels and the subsequent formation and precipitation of Fe0H 3,

7-2 Bioaccumulation of Aluminium and Iron

resulted in large fish kills in the Blesbok Spruit wetlands, Gauteng, South Africa, as a result of suffocation (Du Preez, personal communication, 1996).

In this section of the study, the extent and order of bioacccumulation of aluminium and iron were determined in the skin, liver, muscle and gill tissue of Clarias gariepinus and Labeo umbratus, as well as the dependence of the bioaccumulation upon the species of fish, the localities, sex, seasons and the lengths of the fish.

7.2 MATERIALS AND METHODS

C. gariepinus and L. umbratus were sampled and dissected as described in Chapter 4. The laboratory and statistical procedures for aluminium and iron analyses of the fish samples were the same as the procedures described for zinc and copper analyses, but in order to determine the aluminium concentration in the tissue samples of the fish by means of the Atomic Absorption Spectrophotometer, 0.5 ml of potassium chloride (KC1) solution (200 g KCl per litre distilled water ) was added to the 50 ml sample, to suppress ionisation of the aluminium (Varian, 1989).

7.3 RESULTS

73.1 DIFFERENCES IN BIOACCUMULATION OF ALUMINIUM AND IRON IN THE DIFFERENT

TISSUES /ORGANS The aluminium concentrations in the tissues/organs of the C. gariepinus and L. umbratus showed high variations. Aluminium accumulated mostly in the gills and liver of the two

7-3 Bioaccumulation of Aluminium and Iron

species (Fig 7-1, Fig 7-2), followed by the muscle and skin tissues. Comparisons between the different tissue types of C. gariepinus were predominantly significant (Ps0.05) except for the muscle and skin tissue at locality OR1 (Table 7-1). For L. umbratus, the gill tissue differed significantly (Ps0.05) from all the other tissue types, as did the muscle and liver tissue at locality OR1 (Table 7-2). The bioconcentration factors, (BFv„) for aluminium concentrations in the different tissues/organs of the two species and the water, ranged between 8 (calculated for the muscle of L. umbratus at locality KOR1 in May 1995) and 1 383 (calculated for the gills of C. gariepinus at locality OR1 in February 1994) (Table 7-1 Appendix). The bioconcentration factor (BF,) for aluminium concentrations in the tissues/organs and the sediments ranged from 0.001 to 0.007 (in the gills of C. gariepinus at locality KOR1 in February 1995).

The highest iron concentrations were found in the liver of both species (Fig 7-3, Fig 7- 4), with the exception in August 1994 at locality KOR1 for L. umbratus, where the gills accumulated higher concentrations of iron. The lowest iron concentrations were found in the muscle and skin tissues of both species. Differences in bioaccumulation of iron in the tissues/organs of C. gariepinus were mostly significant (Ps0.05), except for the muscle and skin tissue at both localities (P>0.05) (Table 7-3). For L. umbratus the bioaccumulation of the iron between the different tissues/organs in the gill and liver tissue at locality KOR1 and the muscle and skin tissue at both localities (Table 7-4) were not significantly different (P>0.05). Iron concentrations fluctuated between 42 pg/g dry mass in the skin of C. gariepinus and 3 963 p,g/g dry mass in the liver of L. umbratus. The general order of bioaccumulation for iron was: The bioconcentration factor (BF„,) for iron ranged between 43 (for the muscle tissue of C. gariepinus in February 1994 at locality KOR1) and 3 470 (for the liver of L. umbratus in November 1994 at locality OR1). The bioconcentration factors for the tissues/organs and the sediment (BF,) ranged from 0.001 (in the muscle of C. gariepinus in February 1994 at locality OR1) to 0.07 (in the liver of L. umbratus in February 1995 at locality OR1).

7-4 Bioaccumulation of Aluminium and Iron

400 250

200 oA

.? •

150

8 0 E 100

• -07

Feb 94 Nov 94 Feb 95 May 95

EgSkin ❑ Liver EMuscle •Gills

FIG 7-1 THE MEAN ALUMINIUM CONCENTRATIONS IN THE DIFFERENT TISSUES/ORGANS OF CIARIAS GARIEPINUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

7-5 Bioaccumulation of Aluminium and Iron

200

0 Aug 94 Nov 94 Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

/ / W. Skin OLiver LI Muscle iiGills

FIG 7-2 THE MEAN ALUMINIUM CONCENTRATIONS IN THE Mt 1. ERENT TISSUES/ORGANS OF LABE° UMBRATUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM AUGUST 1994 TO MAY 1995

7-6

Bioaccumulation of Aluminium and Iron

TABLE 7-1 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE ALUMINIUM CONCENTRATIONS IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS IN FEBRUARY 1994 (F1), MAY 1994 (M1), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2), MAY 1995 (M2)

.• - ...... , 11 ,,, al A; U aaLlg.a. : :La :a ":=:" - .:=1::::- :map === Ira =:::- .-- jar ; .-. "4- a "::: aLaipa,- LOCALITY KOR1 ...... ... R. 'ARM,' ,1 :.• .e. r. .1 • :.•? i.rf a 'T.,:' :•:::: ' "i?" "'• " -4' Ft F„ N, F„ m2 tmEng: --.: ,-- ..-..,-".1 ,- -,...-:.: ---:---..ha. :41 nnntr.: „ - 7+:9 r",,,i-#: ...1,=- f -Itnit: .....,-, . '1' -•' 411.:-.1Zu ,„..v. -- - -4ini.rtir , :1-....;' MI Ft , N . , . r.„., a "a .1- ... -:-M ..ri ,z. ..'- :1 :F- r:Eliirtiiinr 7:••• -1...1:1.1Rillalinn :: „ „„„ ;;;..; - 11 ,..,F L L i":::= = 1::::-24.--7.: iii' "it' ...... ,4,-•-;_._...... Ti. ,„ r, A 1---Xiiiii1:41 ,'''' Hi -,.....nniiii:11.i.„.1iFiii1 ,- . aiii, ..T 2 nn 'EM M.. ., ". := I'M _ wo..- , _7' i h --z•- -+ -• : '••- - -.= 1:::::::::= Lo c A L r r y 0 R 1

2 Slug qiP- -,r-Fillik.RA.. I M1 ns.. F„ Mt , A, F2 , M it1:::114 :am -=ar: LIaLLI.=::" " ; r : : : r: :Al 1,,, . .. - :::.==-A '' aii1.--, Lliratri"1-4- -41,. a•TiLEV. Ati- -' ...... -..d.unvire lik+,..Zi-- -..,.-...4.4.W.MLIML.-,.:-, .1... -24^ T. ....H!..i. ..9E. Mt Mt 9, A. N, F2 , M2 - = ::::-•••=7:74•it:::::: . -...te.-1.617: n11..-Tttr.1. n • - - .. r r... =:-T• 1=.. ...... „ . i.:101:11,LLr, . :::::..L--:.s7".maa41- r :::, . i: ,: = " :It :9 -mi ar,ar Ft , Mt , F2 , m2 Muscle = iffrd:--„,...,::::h,... r..,,,..., ....., ...... ::::::iii, " .,.i.:. .... . . n.... • - ... ''''''trz T1'..,.. 4 ' . Jaw: q -'411?:la FF.1. M. ".,,' F.•-,. :. '-7, -it v....? ':..4t4111 ,,--''.- .... ...... ' '''' ' 7:÷-';T:IF.T. , ... -11:L75 a". . ?'''''Pmr,1-' al LP+ - , . n.4, 77,0.1 "",..- ..Ii.IriF.F.qMi1r

TABLE 7-2 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE ALUMINIUM CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE(' UMBRATUS IN AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

11 ii^. • 1.1.4f.a.sr Aire4:-41=._ -44e, ,. -:-•••••0amiEa..r, - :11,,,, • E.• ....".1,, 14. ? ,'"1-nr,„ , ; , ' .,,„.-, a: " ..... - ..=Lar..-r-aad.41 - ^ - .^-= ,..-,.....=-.=-----re--7,-...._1_-:%-naraim. ,'....: =.: : - ."-tP--", F-,T :IE.:42:F .,..„-, -,- -%::,‘ n n'". nnn,._ .,, — LOCALITY KOR1 .46...F...., ,141: :4:-=.g..zzliflim-,,--::---,_:f..1:1-1....rzali.iig n s n.s. A, F2 ■ '''''-' -...• "".., 'tn.. . T.t"-...."97.=01,713.:11:::Er--,.-..-•-••••••-, - • • . . .÷P.... ' IttAli 11.:41 ...:-•...1.-1--- TV.Y. . „,Ilai ...! '''''''' -14.'"-:"::,-a%_ r•L:.•:Preil-' US. A, N. F2 , M2 ltvcr ''•''nanniumnTa7f.,- •,•,•• ...:1•9•,=91:2• ,.., a: .n. nu i5Eille1/4•10: . ""'''',' ' ...griiiiiiiiii:....._ -..-- ''' ''`'''-.ILIA__ 'Ti • - i-..7 ',--,.? - "..1"--1. ',--73114.7.41,,,,,f,,IMMIM,11' ,:1/4c,-... --...... ,,,,k.n--- tiatilitp,, :Ki: -I' Tiii? --f-:=FTh.- ,air,inE nr,„a ,,-_t. -4' - Ay F2 , M2 s;•;. . , FA . v ___ ..__,....1:1:1 ,r,,..... El.: a .._ala?'" :: ,,,., ==,= Z=2==2.-.1-"•A-:":" . 1:41::.-": =aaiairgifilitk-- ::::::::: .:=-- -If =,:iii, i "-is:14.,• 4:-.::,-,-, ::::. ,,,. ..., "....11.. L:•: n :MIN - ...... iiiiiir -ir LOCALITY OR1

..7. 9.-9 -9 . tr.• .: ::2...ilia, ....:Li-,Tr:....-r''' .::::7:----. - '-,==.-::,*-H- ... •,,r,"-- a:.. -'.0.-.9 — li A1:-;e1' F.:' . . . ... , ,-. ..... ,,E ,F-i-,,F :,... .- 'ii 11..-1:4. 9....."-., -".... -"-...... Prt:,'• . - .....i.T.+.! IP- :--il' -1....1.111.-!:=. F" "24'... .1: ek, ' :,... ,111,3.. r. .1.14. a 111::::. .n., ...... -. - ...... nn ...... nnn q,.: n ,..• !Elii:F.:.-....------411 ..:41-::4....:1-.F.--'r.11.01.4.^7.' 111P1: .,.1 -, .- . tn.'. , .0 . .,.., — :1-1:::: 1: . ;,-- :::-... nn,...q.niEr-Hilli+e ..,-," '1/4:=E==;•:::..,,,n1:-. "="‘": fa ''. ' := a •Lra: - =•••-•

7-7 Bioaccumulation of Aluminium and Iron

2,500 1,800

1,600

2,000 1,400

ew CiA 1,200 b C 1,500 0 O .-. —• 1,000 cu eu

U U 800 O 8 1,000 8 600 2 2

400 500

200

0 0 Feb 94 Nov 94 Feb 95 May 95 Feb 94 May 94 Aug 94 Nov 94 Feb 95 May 95

Skin ❑ Liver ❑ Muscle U Gills /

FIG 7-3 THE MEAN IRON CONCENTRATIONS IN THE DIFFERENT TISSUES/ORGANS OF CLARL4S GARIEPINUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

7-8 Bioaccumulation of Aluminium and Iron

700

600

...SOO OD -a)

8 0 2 ' 200

100

0 0 Aug 94 Nov 94 Feb 95 May 95 Aug 94 Nov 94 Feb 95 May 95

/ / 3 Skin ❑ Liver 3 Muscle /Gills /

FIG 7-4 THE MEAN IRON CONCENTRATIONS IN THE DIFFERENT TISSUES/ORGANS OF LABE° UMBRATUS AT LOCALITY KOR1 (A) AND LOCALITY OR1 (B) FROM FEBRUARY 1994 TO MAY 1995

7-9 Bioaccumulation of Aluminium and Iron

TABLE 7-3 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = P50.05; NON SIGNIFICANT = P>0.05) BETWEEN THE IRON CONCENTRATIONS IN THE TISSUES AND ORGANS OF CLARIAS GARIEPINUS IN FEBRUARY 1994 (F,), MAY 1994 (M,), AUGUST 1994 (A), NOVEMBER 1994 (N), FEBRUARY 1995 (F2), MAY 1995 (M2)

TABLE 7-4 SUMMARY OF THE DIFFERENCES (SIGNIFICANT = PS0.05; NON SIGNIFICANT = P>0.05) BETWEEN THE IRON CONCENTRATIONS IN THE TISSUES AND ORGANS OF LABE() UMBRATUS IN Auclin 1994 (A), NOVEMBER (N), FEBRUARY 1995 (F2) AND MAY 1995 (M2)

...... . . „,„

LOCALITY KOR1 ,,, • A, F2 , M2 n.s. A, F2 , M2

,, = • , ,,,,,,,,,, A, F2 IP M2 ns.

US A, DA,

Loc.Aurt OR1

A, N, F2 I M2 ns. A, N, F2 , M2

N, F2 M2 N, F2 Ern' F2 , M2

7-10 Bioaccumulation of Aluminium and Iron

4.3.2 SPECIES DIFFERENCES No definite trend could be established, concerning the bioaccumulation of aluminium in the different tissues/organs between the two species (Fig 7-5, Fig 7-6). For instance, the highest aluminium concentrations in the skin, were found for C. gariepinus in May and November 1994, while L. umbratus accumulated more aluminium in the skin in May 1995 at locality KOR1. L. umbratus also accumulated more aluminium in the muscle tissue in May and November 1994, but C. gariepinus showed higher levels of aluminium in the muscle in February 1995 at locality OR1. Higher concentrations of aluminium were found in the gills and liver of L. umbratus in May 1995 at locality OR1, whereas the highest concentrations in the gills in November 1994 at locality KOR1, were found for C. gariepinus.

In the case of iron, a trend could be established to some degree, as L. umbratus accumulated higher concentrations of the metal in the skin and muscle tissues in November 1994 at locality OR1 (Fig 7-7, Fig 7-8). Higher concentrations of iron were found in the gills of C. gariepinus in May and November 1994, but in May 1995 at locality OR1, the highest concentrations were found for L. umbratus. C. gariepinus accumulated more iron in the liver in November 1994, but higher concentrations were found in the liver in November 1994 at locality OR1 and February 1995 at locality OR1 for L. umbratus.

4.33 RELATIONSHIP BETWEEN LENGTHS AND ALUMINIUM AND IRON CONCENTRATIONS Sample sizes were too small for individual analysis concerning the relationship between the concentrations of aluminium and iron and the fish lengths, therefore these variables were grouped together in order to obtain statistically verifiable results. Significant negative correlations (P50.05) were found between the aluminium concentrations in the skin, liver, muscle and gills of the two species and the lengths of the fish. Iron showed significant negative correlations between the lengths of the fish and the concentrations

7-11 Bioaccumulation of Aluminium and Iron

500 100

ao g

..4, 1 400 (A 80

le sc mu he t

300 in 60

0 n io t a tr n e

8 200 40 E conc ium in m lu a

Ls 10o n 20 Mea

• . • • • • . • ...... May 94 (OR1) Nov 94 (KOR1) Nov 94 (OR ) May 95 (OR1 May 94 (OR1) Nov 94 (KOR1) Feb 95 (OR1)

Clarias gariepinus Ei Labeo umbratus 2

FIG 7-5 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN CLARL4S GARIEPINUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF ALUMINIUM IN THE SKIN (A) AND MUSCLE (B) FROM MAY 1994 - MAY 1995

7-12

Bioaccumulation of Aluminium and Iron

A B 400 60

__ea) 50

cn 300 a) b.0 N 40 0

0 0 c-. as

-.a- 200 •a' 30 a) c.)4) 8 8 5 E 20

...... cct 100

ets ...... a) a) 10

0 0 Nov 94 (KOR1) May 95 (OR1) May 95 (OR1)

LI Clarias gariepinus ❑ Labe° umbratus

FIG 7-6 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN CLARL4S GARIEP1NUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF ALUMINIUM IN THE Gals (A) AND LIVER (B) IN NOVEMBER 1994 AND MAY 1995

7-13

Bioaccumulation of Aluminium and Iron

A B 1,000 1,200

1,000

) 800 es0 /g g (p. kin s he 600 t

in = ion t tra ...a. 400 =0 concen 8 400 iron

Me 200 is) 200

0 0 Nov 94 (OR1) Nov 94 (OR1)

El Clarias gariepinus ill Labeo umbratus /

FIG 7-7 SIGNIFICANT DIFFERENCES (1)..50.05) BETWEEN CLARIAS GARIEPINUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF IRON IN THE SKIN (A) AND MUSCLE (B) IN NOVEMBER 1994

7-14 - •

Bioaccumulation of Aluminium and Iron

A 3,000

2,500 at 5

Sp 4

I-

0 os 1,500 3

4.) c.) 0 8 O 1,000 C 8 8 2 0 500

May 94 (ORI) Nov 94 (KOR1) May 95 (ORD Nov 94 (KOR1) Nov 94 (OR1) Feb 95 (OR1

Clarias gariepinus ❑ Labe° umbratus

FIG 7-8 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN CLARL4S GARIEPINUS AND LABEO UMBRATUS REGARDING THE BIOACCUMULATION OF IRON IN THE GILLS (A) AND LIVER (B) FROM MAY 1994 TO MAY 1995

7-15 Bioaccumulation of Aluminium and Iron

in the skin and gills, and insignificant negative correlations for the liver and muscle tissues.

7.3.4 DIFFERENCES BETWEEN MALES AND FEMALES Although comparisons between the males and females of C. gariepinus and L. umbratus at the two localities during the period February 1994 to May 1995 showed predominantly no significant (P>0.05) differences, there were significant differences (P..0.05) in November 1994 at locality OR1 for L. umbratus in the skin, and in May 1995, at locality OR1 for C. gariepinus in the muscle tissue (Fig 7-9).

73.5 SEASONAL DIFFERENCES Statistical analyses were mostly carried out for C. gariepinus, except for the comparison between winter and spring 1994, where L. umbratus was used, as a result of small sample sizes for C. gariepinus. No significant differences (P>0.05) were found between summer 1994/1995 and winter 1994 regarding the bioaccumulation of aluminium in the different tissues/organs of the two species (Table 7-5). The significant differences found between the seasons, appeared to be more likely in the skin, followed by the liver and then the gills and muscle tissues. For iron, the significant seasonal differences were found mainly for the muscle tissue, followed by the gills, the skin and the liver (Table 7-6). No significant differences were found between summer, spring 1994, summer 1995 and autumn 1995 and between winter 1994 and summer 1995.

73.6 LOCALITIES DIFFERENCES Several significant differences (P.0.05) were found between the two localities, regarding

7-16

Bioaccumulation of Aluminium and Iron

A B 1,000 600

) 800 /g g (g

kin g 400 s

he a)

t 600 0 in .... 0

ion .

t 0 300 a tr

400 concen on ir an

200 Me

0 Male Female Male Female

❑ Males M Females \

FIG 7-9 SIGNIFICANT DIFFERENCES (P50.05) BETWEEN THE MALES AND FEMALES, REGARDING THE BIOACCUMULATION OF IRON IN THE SKIN OF LABEO UMBRATUS IN NOVEMBER 1994 AT LOCALITY OR1 (A) AND IN THE MUSCLE TISSUES OF CLARIAS GARIEPINUS IN MAY 1995 AT LOCALITY OR1 (B)

7-17 Bioaccumulation of Aluminium and Iron

0C? A

7-18 • Bioaccumulation of Aluminium and Iron z 0 cd g.. 0 z 0 S p U 4tU 0 as c4

0 4 g cri in z cn i2 it r„ z z o 0 0.

.."'-' . cr, 1:4 ;a Z1 gct ›. 2 z 1 0 g 1 ,22 LaL, s x . 0 g .4 ,T., ad ...., t al t, a

.5 4az 44 w w a x z Q 1 gal 1 let Gr.

A 4 v, 4 a E *Q.. w r: w il, E 5) Li. 0 0 ; ›, 0. P> ( C..) t= t 2 cc m 0 ican co) ...... , ifn ig ts N No =

N Q

7-19 Bioaccumulation of Aluminium and Iron

the bioaccumulation of aluminium (Fig 7-10) and iron (Fig 7-11) in the skin, liver, muscle and gills of C. gariepinus and L. umbratus. In May 1994, aluminium concentrations in the muscle and gills of L. umbratus were higher at locality OR1. However, L. umbratus from locality KOR1 had accumulated high concentrations in the gills (February & May 1995) and in the muscle tissue (February 1995). The skin tissue of C. gariepinus captured during May 1995 at loality KOR1 also had significantly higher aluminium (11 0.05) concentrations compared with concentrations at locality OR1.

For iron, locality KOR1 showed higher concentrations of the metal, only in May 1994 in the muscle tissue of L. umbratus and C. gariepinus, while locality OR1 showed higher levels of iron in February and May 1995 in the muscle, liver and skin tissue of L. umbratus.

7.4 DISCUSSION

Although a number of studies have investigated the toxic effects of metals such as aluminium on fish under laboratory conditions (Neville, 1985; Schofield & Trojnar, 1980; Siddens, Seim, Curtis & Chapman, 1986; Freeman & Everhart, 1971), fewer have been concerned with the levels of contamination of this metal which occur in natural fish populations in South Africa. The present study disclosed that there is differential bioaccumulation of aluminium and iron in the different tissues/organs of C. gariepinus and L. umbratus. The distribution pattern of aluminium among the four tissue types was generally: G>L>S ,----M. Stripp, Heit, Bogen, Bidanset & Trombetta (1990), also showed that there were significant concentrations of aluminium in the gill tissue of the omnivorous

7-20 Bioaccumulation of Aluminium and Iron

A 100 B 300

0.0

Q) -0 200

C. .E 60 0 0 Ce Ce 4. 'a' 150 4) U

g 40 8 E 5 'E 100

z Ce 73. 0 20 ce cat U 50

0 May 94 Feb 95 May 94 Feb 95 May 95 C 100

Z.0 8

4). U

.—• 60 ❑ Locality KOR1 El Locality OR1 0 ;43 s-•

04.)

8 40

ief Ncl 20

May 95 FIG 7-10 SIGNIFICANT DIFFERENCES (PS 0.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF ALUMINIUM IN THE MUSCLE (A), GILLS (B) AND SKIN (C) OF LABEO UMBRA TUS FROM MAY 1994 TO MAY 1995

7-21 Bioaccumulation of Aluminium and Iron

500

04 --a 400

,..S 0.) 73v.3 z

0 c.) 200 0 8 O= --- ,-4

100

0 May 94 Feb 95 May 95 Feb 95 May 95 800

CO 600

( ❑ Locality KOR1 D Locality OR1 ..=o.) z

,-.0 0 .....0 174 400

0 0 0 0 8 o O $... at 200 N

Feb 95 May 95 FIG 7-11 SIGNIFICANT DIFFERENCES (P .5. 0.05) BETWEEN LOCALITIES KOR1 AND OR1 REGARDING THE BIOACCUMULATION OF IRON IN THE MUSCLE OF LABEO UMBRATUS AND CLARIAS GAREEPINUS (A) AND THE LIVER (B) AND SKIN (C) OF LABEO UMBRATUS FROM MAY 1994 TO MAY 1995

7-22 Bioaccumulation of Aluminium and Iron

white suckers, Catostomus commersoni and the carnivorous yellow perch, Perca flavescens and that aluminium does not accumulate in significant quantities in other tissues. For iron the order of bioaccumulation was generally: L.--G>S=M, where the higher iron concentrations in the liver are supported by a study performed by Seymore (1994) in the lower catchment of the Olifants River. The higher concentrations of iron in the liver may be due to iron-containing enzymes and the extensive vascular system of the liver, as the haemoglobin in the blood binds approximately three-quarters of the iron in the body (Voynar, 1960), explaining the high accumulation in the liver. It is also clear that there is a definite uptake of aluminium and iron via the gills, in which there is an intimate blood-water contact and the gills are important for osmotic and ionic regulation. High iron levels in the gills could be due to complexing of the metal with the mucus on the gills (Heath, 1987). Although gill and liver concentrations of aluminium and iron were much higher than concentrations in the muscle and skin tissues, mean concentrations of these metals in the muscle and skin tissues were similar. Aluminium is physiologically non-essential (Trapp, 1986) and inorganic aluminium is poorly absorbed and rapidly excreted in the urine. The major effect on humans is the possibility of neurological disorders, such as Alzheimer's disease, in renal dialysis patients treated with water obtained from drinking-water supplies rich in aluminium of more than 0.11 mg/I (WHO, 1981). Judged by the aluminium and iron concentrations in the muscle and skin tissue of C. gariepinus and L. umbratus, these fish are suitable for human consumption. According to the recommended intake of iron of 10 mg and 18 mg for males and females respectively (NAS, 1980) and therefore it does not pose any serious threat to human health.

The bioconcentration factors between the water and the different tissues/organs (BF„,) of the two species ranged between 43 and 3 470, which were lower than the bioconcentration factors calculated for fish from the lower catchment of the Olifants River, ranging between (0.2 and 49 705) (Seymore, 1994). These low bioconcentration

7-23 Bioaccumulation of Aluminium and Iron

factors for iron as well as aluminium, were found in the muscle tissues of both species and the highest bioconcentration factors (BF„,) were found for aluminium in the gills of C. gariepinus and for iron in the liver of L. umbratus. Studies have shown high concentrations of aluminium in the gills of fish from acidic lakes, but during this study, at localities KOR1 and OR1, the pH of the water was high (pH 7.6 - 9.1, see Chapter 3, Table 3-1 Appendix) thereby reducing the rate of release from the sediments. The water of the upper catchment is moderately hard, which can result in complexes forming between the metals and the carbonate in the water, resulting in low bioavailability (low bioconcentration factors) of the metals from the water and sediment. It is, however, important to note that the bioconcentration factors give no indication of relative availability of the various metals in a certain tissue/organ, if the metal concentration is regulated (Wiener & Giesy, 1979). Nevertheless, elements that are essential and therefore present in fish tissue, are regulated, while non-essential elements, such as aluminium, are not at all, or poorly regulated. The detection of these metals in the muscle, will confirm pollution but the regulation of metals in this tissue eliminates the possibility of using only muscle tissue as indicator tissue for these metals. The bioconcentration factors between the tissues/organs of the two species and the sediment (BF,), indicate that even less iron and aluminium in the sediment were bioavailable for uptake by the fish, which may be due to the alkaline pH of the water, as metals tend to be mobilised from the sediment in acidic water, thereby brought into solution.

Statistical analyses indicate various significant differences (P5.0.05) between the two species regarding the bioaccumulation of aluminium and iron in the skin, liver, muscle and gill tissues. Differences in mean aluminium concentrations were, however, much less pronounced in the liver tissues than the iron concentrations in this organ. During chronic exposure, a constant level of the metal will be seen in the blood. If an organ like the liver has a high affinity for the metal, as in this case iron, the latter will be acccumulated there and high levels of this toxicant can build up in this tissue.

7-24 Bioaccumulation of Aluminium and Iron

Differential uptake is often the result of the general structure of a chemical and instances of active uptake by one organ are observed, to a certain maximum concentration, after which this chemical is regulated. There were no consistent higher or lower aluminium concentrations in specific tissues and organs of a specific species, for example, the concentrations of aluminium in the skin tissue for C. gariepinus were the highest during May and November 1994 at both localities, while the aluminium concentrations in the skin, gills and muscle tissue during May 1995 were the highest for L. umbratus. The same trend was observed for the iron concentration in the gill and liver tissues. These differences in the accumulation of aluminium cannot be ascribed to the different feeding habits of the two species, as biomagnification of aluminium does not appear to occur in the aquatic food chain (Wren, MacCrimmon & Loescher, 1983). The bioaccumulation of aluminium and iron in this section of the study, is independent of the species of fish, as no definite trend as to where the highest concentrations were found, could be established.

Examination of the extent of bioaccumulation of aluminium and iron in the skin, liver, muscle and gill tissues of C. gariepinus and L. umbratus, revealed a strong negative correlation with the size of the fish. Results of relationship analyses between body size and aluminium and iron concentrations, generally showed decreases in metal concentrations with increasing fish size, which supports the findings of Bohn & McElroy (1976), where the same relationship was found. Decreases in metal concentrations with increasing fish size, may be due to incorporation of new tissues in the muscle and skin as the fish grows, or due to the rate of metabolism and ventilation, which is much lower in larger fish (see Chapter 4 for detailed discussion).

Bioaccumulation in the different tissues/organs of the two species showed no definite trend between the males and the females. In this case, it is also suggested that the male and female gonads of the fish should be compared to obtain significant differences in the

7-25 Bioaccumulation of Aluminium and Iron

bioaccumulation of metals between males and females, as were found by Seymore (1994) (see Chapter 4 for detailed discussion).

SEASONAL DIFFERENCES Significant differences were found between seasons in the skin, muscle, liver and gill tissues of C. gariepinus, but the differences were not always indicated by the same tissues/organs, as were the case for iron in the tissues/organs of Barbus marequensis from the lower catchment of the Olifants River (Seymore, 1994). The aluminium and iron concentrations were the lowest in the summer of 1994 and this could be due to the lower metal concentrations in the water in this season (see Chapter 3, Table 3-3) and the low bioavailability of the metals to the fish due to the moderate hardness of the water. The same seasonal trend was found for autumn 1994, compared to the other seasons, with the highest aluminium and iron concentrations found in the summer/autumn of 1995, where high concentrations of these metals were found in the water at the two localities. This may be attributed to environmental factors including the low flow of the water, due to less rainfall during this period, concentrating the metals in the water.

LOCALITIES DIFFERENCES In general, fish at locality OR1, which receives water from various industries and mines in the catchment area, accumulated more aluminium and iron than the fish at locality KOR1 in August 1994. The aluminium concentrations in the water in August 1994 were slightly higher at locality OR1, but for iron, the concentrations in the water were similar at the two localities. In May 1994, C. gariepinus accumulated more iron at locality KOR1, despite the lower iron concentrations in the water or higher exposure concentrations between sampling at this locality. In these cases, elevated accumulation of the metal may be due to different physical/chemical properties of the water at that time. The highest iron concentrations were found in fish at locality OR1 in February

7-26 Bioaccumulation of Aluminium and Iron

1995, which coincides with much higher iron concentrations in the water at locality OR1. This may be due to point and diffuse source effluent from the coal mines in the catchment area. For aluminium, the fish at locality KOR1 accumulated higher concentrations of the metal in February 1995, which can be explained by the much higher aluminium concentrations in the water at this locality at this time. Fish at locality KOR1 in May 1995, also accumulated higher levels of aluminium, as the aluminium concentrations in the water were much higher at this locality.

7.5 CONCLUSION

Different tissues and organs were analysed in order to determine the degree and order of bioaccumulation of aluminium and iron in Clarias gariepinus and Labeo umbratus. These metals may be absorbed across the entire body surface of fish, as well as by the gills and may be stored in organs and tissues like the liver and muscle. The highest concentrations of aluminium and iron were found in the gills and liver respectively and these organs, together with the muscle tissue, are suggested for use for analyses of aluminium and iron in fish. Accumulation of these metals by fish is dependant on the size of the fish and to some extent the species and localities where these fish are caught.

An important factor to take into account is the metal concentrations in the water at the different localities, as higher concentrations of especially iron were accumulated at locality OR1, receiving effluent from mines and different industries causing higher concentrations of this metal in the water, as well as the physical and chemical properties of the water. Fish absorb dissolved metals and can therefore serve as a reliable indication of the extent of pollution by these metals in an aquatic system, although factors such as the regulation of metals by the fish should be taken into account, as well as the fact that the fish are not stationary.

7-27 Bioaccumulation of Aluminium and Iron

7- 6 REFERENCES

BOHN, A. & McELROY, R.O. (1976). Trace metals (As, Cd, Cu, Fe and Zn) in Arctic cod, Boreogadus saida and selected zooplankton from Strathcona Sound, northern Baffin Island. J. Fish. Res. Board Can. 33:2836 - 2840.

FREEMAN, R.A. & EVERHART, W.H. (1971). Toxicity of aluminium hydroxide complexes in neutral and basic media to rainbow trout. Trans. Am. Fish. Soc. 4:644 - 658.

HARRIMAN, R. & MORRISON, B.R.S. (1982). Ecology of streams draining forested and non-forested catchments in an area of central Scotland subjected to acid precipitation. Hydrobiologia. 88:251 - 263.

HEATH, A.G. (1987). Water Pollution and Fish Physiology. CRC Press, Inc., Florida. 245 pp.

LYDERSEN, E., BJORNSTAD, H.E. & ENGLUND, J.O. (1987). Size distribution patterns for trace elements in waters draining a small catchment, S.E. Norway. In: Proc. Int. Symp. on Acidification of water Pathways. Bolkesjo 4 - 5 May 1987 1:107 - 116.

LYDERSEN, E., POLED, A.B.S., MUNIZ, I.P., SALBU, B. & BJORNSTAD, H.E. (1990a). The effects of naturally occurring high and low molecular weight inorganic and organic species on the yolk-sack larvae of Atlantic Salmon (Salmo salar L.) exposed to acidic aluminium-rich lake water. Aquat. Toxicol. 18:219 - 230.

7-28 Bioaccumulation of Aluminium and Iron

LYDERSEN, E., SALBU, B., POLED, A.B.S. & MUNIZ, I.P. (1990b). The influences of temperature on aqueous aluminium chemistry. Water, Air, Soil Pollut. 51:203 - 215.

McDONALD, D.G. (1983). The effects of W upon the gills of freshwater fish. Can. J. Zool. 61:691 - 703.

MUNIZ, I.P. & LEIVESTAD, H. (1980). Toxic effects of aluminium on the brown trout, Salmo trutta L. pp. 320 - 321. In: D. Drablos & A. Tollan [ed.]. Ecological impact of acid precipitation. SNSF Project, Norway.

NAS (NATIONAL ACADEMY OF SCIENCES) (1980)." National Academy of Sciences. Recommended Dietary Allowances (9 th edn.). Washington DC, Printing and Publishing Office, National Academy of Sciences.

NEVILLE, C.M. (1985). Physiological response of juvenile rainbow trout Salmo gairdneri to acid and aluminium - prediction of field responses from laboratory data. Can. J. Fish. Aquat. Sci. 42:2009 - 2019.

SCHOFIELD, C.L. & TROJNAR, J.R. (1980). Aluminium toxicity to brook trout Salvelinus fontinalis) in acidified waters, pp. 341 - 363. In: T.Y. Toribara, M.W. Miller & P.E. Morow [ed.]. Polluted rain. Plenum Press, New York.

SEYMORE, T. (1994). Bioaccumulation of metals in Barbus marequensis from the Olifants River, Kruger National Park and lethal levels of manganese to juvenile Oreochromis mossambicus. M.Sc. RAU.

7-29 Bioaccumulation of Aluminium and Iron

SIDDENS, L.K., SEIM, W.K., CURTIS, L.R. & CHAPMAN, G.A. (1986). Comparison of continuous and episodic exposure to acidic aluminium contaminated waters of brook trout (Salvelinus fontinalis). Can. J. Fish. Aquat. Sci. 43:2036 - 2040.

STRIPP, R.A., HEIT, M., BOGEN, D.C., BIDANSET, J. & TROMBETTA, L. (1990). Trace element accumulation in the tissues of fish from lakes with different pH values. Water, air, and Soil Pollution. 51:75 - 87.

TRAIN, R.E. (1979). Quality criteria for water. U.S. Environmental Protection Agency, Washington D.C. Castle House Publications LTD. pp. 82 - 94.

TRAPP, G.A. (1986). Interactions of aluminium with cofactors, enzymes, and other proteins. Kidney Int. 29(18):S12 - S16.

VARIAN (1989). Flame Atomic Absorption Spectrometry: Analytical Methods. Varian Techtron Pty Limited, Australia, 146 pp.

*VOYNAR, A.I. (1960). Biologicheskaya rol' mikroelementov v organizme zhivotnykh i cheloveka. (Biological Function of Trace Elements in Animals and Man). Vyssh, Shk. Press, Moscow.

* WHO (WORLD HEALTH ORGANIZATION) (1981). Criteres d'hygiene de l'environnement 18, Geneve (Switzerland).

WIENER, J.G. & GIESY, J.P. (1979). Concentrations of Cd, Cu, Mn, Pb, and Zn in fishes in a highly organic softwater pond. J. Fish. Res. Bd. Can. 36:270 -279.

7-30 Bioaccumulation of Aluminium and Iron

WREN, C.D., MACCRIMMON, H.R. & LOESCHER, B.R. (1983). Examination of bioaccumulation and biomagnification of metals in a precambrian shield lake. Water, Air and Soil Pollution 19:277 - 291.

WRIGHT, R.F. & GJESSING, E.T. (1976). Acid precipitation I Changes in the chemical composition of lakes. Ambio 5:219 - 223.

YOUSON, J.H. & NEVILLE, C.M. (1987). Deposition of aluminium in the gill epithelium of rainbow trout (Salmo gairdneri Richardson) subjected to sublethal concentrations of the metal. Can. J. Zool. 65:647 - 656.

* These articles were not reviewed by the author

7-31 Bioaccumulation of Aluminium and Iron

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7-32 Bioaccumulation of Aluminium and Iron

7-33 Bioaccumulation of Aluminium and Iron

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7-34 Bioaccumulation of Aluminium and Iron

7-35 Bioaccumulation of Aluminium and Iron

7-36 Bioaccumulation of Aluminium and Iron

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THE EFFECTS OF LOW PH AND ALUMINIUM ON THE HAEMATOLOGY, OSMOREGULATION AND CARBOHYDRATE METABOLISM OF OREOCHROMIS MOSSAMBICUS

8.1 INTRODUCTION

The physiological responses of freshwater fish to environmental acidity have been studied extensively over the past years (Fromm, 1980; Mc Donald, 1983; Wood, 1985). Research on the effects of surface water acidification on fish has shown that water quality exerts complex influences on fish responses to acidification (Martin, 1987). Acidity is a primary controlling and regulating variable in many chemical and biological processes and therefore changes in the pH cause many additional changes in the water quality, all of which can affect aquatic organisms. In nature, acid surges are almost exclusively a soft water problem (i.e. [Ca 2+] below 0.5 and less than 0.2 mequiv 1 4), because hard water catchments have sufficient bicarbonate alkalinity to neutralize precipitation acidity (Wood, 1985).

Of importance are the effects of pH on the solubility and speciation of metals, especially aluminium. Acidification of soil systems may cause the transfer of aluminium into aqueous solutions where it may be present in different physico-chemical forms, varying in charge and size properties. The chemistry of aluminium in natural waters is, very complex. The solubility of aluminium increases exponentially as the pH decreases (Wood, 1985), for instance, minimum solubility occurs at pH 6.0 - 7.0, while there is a greater than 7000 fold increase in solubility over the pH range from pH 5.9 - 4.0. At

8-1 Haematology, Osmoregulation and Carbohydrate Metabolism

pH 4.0 the free aequo-ion (i.e. Al') is the predominant monomeric species (90%), while at pH 6.0 the monovalent and divalent hydroxides predominate. Only inorganic monomeric aluminium seems to contribute to acute toxicity, thus aluminium bound to ligands (organic acids, fluoride, sulphate etc.) can be ignored. The toxicity of aluminium is similarly complex, as maximum toxicity occurs over the pH range of 5.0 - 5.5 and decreases at higher pH values (Muniz & Leivestad, 1980; Freeman & Everhart, 1971). The mechanism of aluminium toxicity is less well studied. According to Neville (1985), aluminium accumulates on the gills in a concentration- and pH dependant fashion, thereby producing epithelial damage and impairment of normal gill function, resembling that seen with low pH. The major trends appear to be as follows: The effect of aluminium varies with pH, changing from a great influence under moderately acidic conditions to a neutral of even protective one at a very low pH. The toxic mechanism of aluminium involves either ionoregulatory or respiratory disturbance or both, and the severity and relative contribution of each being critically dependant upon the pH and. absolute aluminium level. Increases in toxicity are associated with decreasing pH, with increasing aluminium and with decreasing Ca'. Ca2+ has a general protective effect against both aluminium and acid toxicity, but the degree of the protection is dependant upon the pH and aluminium levels.

Physiological and histological information are also important to understand the effects of aluminium and acidity at the level of the organism. Rankin, Stagg & Bolis (1982), suggest that this information may enable predictions about the health of fish populations in the field, as fish deaths from natural waters have been correlated with the presence of aluminium mobilized from soils by acidification. Field and laboratory studies show that ionoregulatory failure is the key factor leading to fish death in acidified waters. The toxic effect may also be exerted through disturbances to physiological functions such as fluid volume distribution, haematological and acid-base homeostasis and oxygen uptake

8-2 Haematology, Osmoregulation and Carbohydrate Metabolism

and transport (Packer, 1979; Milligan & Wood, 1982; Ultsch & Gross, 1979).

Different fish species react differently to combinations of aluminium and low pH (Table 8-1), but the general effects range from a reduction in feeding and growth to pathological degeneration of the gill tissues and ultimately death. Understanding the physiological effects and the mechanisms by which low pH water and combinations of low pH water and elevated aluminium concentrations are harmful to fish, could improve our ability to predict the fate of fish populations exposed to acidified surface waters. The objective of this part of the study was thus to determine the effects of low pH and combinations of low pH and sublethal concentrations of aluminium on haematological variables as well as osmoregulation and the carbohydrate metabolism of the Mozambique Tilapia, Oreochromis mossambicus.

8.2 MATERIALS AND METHODS

In this section, the haematological, osmoregulatory and carbohydrate metabolism variables were measured after a series of exposure tests (96 hours) at 23 ± 1°C at a low pH (pH 5.2) and combinations of low pH and different aluminium concentrations (0.06 mg/1, 1 mg/1, 1.5 mg/1 & 2 mg/1 aluminium). These aluminium concentrations were a range of concentrations found in the Upper Olifants River catchment between February 1994 and May 1995 (Chapter 3; Table 3-2).

8-3 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-1 EFFECTS OF ALUMINIUM AND LOW PH ON THE BEHAVIOUR, REPRODUCTION, PHYSIOLOGY AND HAEMATOLOGY OF DIFFERENT FISH SPECIES

Fish .1A0400fiT

Salvelinus fontinalis 0 ug/I + pH 4 - 43 Death due to electrolyte loss Wood, Playle, Simons, Goss & McDonald, 1988

0 141 + pH 4.9 or 5-5 Reduced growth Siddens, Seim, Curtis & Chapman, 1986

0 p.g/I + pH 52 Increase in plasma cortisol Meuller, Sanchez, Increase in blood glucose Bergman, McDonald, Proliferation and hypertrophy of Rhem & Wood, 1991 chloride cells Lifting of epithelium away from basal lamina Influx of WBC in lymphatic spaces of 2 lamellae Vacuolation and degeneration of pavement epithelial cells

0 p.g/1 + pH 6.1 Hypoxia Wood, Playle, Simons, Goss & McDonald, 1988 47 + pH 4.97 Decreased sodium and osmolality Mount, Hackett & Gem, Decrease in feeding behaviour 1988 Decrease in growth Abnormal vitellogenesis

75 - 150 + pH 52 Damaged gills (lesions extensive and Meuller, Sanchez, severe) Bergman, McDonald, Proliferation of mucus cells throughout Rhem & Wood, 1991 gill filament and secondary lamellae Increase in chloride cells Hyperplastic primary filament and secondary lamellae

228 ;AO + pH 4.4 Hypertophy of chloride cells Ingersoll, 1990 Hyperplasia of chloride cells Hypertrophy of pavement epithelial cells Vacuolation & degeneration of chloride and outer epithelial cells

239uel + pH 4.4 Reduction in feeding Ingersoll, 1990 Loss of equilibrium Increased activity Excessive mucus secretion Changes in skin pigmentation Reduced growth

8-4 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-1 (CONTINUED)

Salve!huts fontinalis 333 pg/1 + pH 4.8 Increase in ventilation Walker, Wood & Decrease in mean arterial Hb02 Bergman, 1991 Increase in PCO2

333 pg/1 + pH 52 Acute hyperventilation Walker et aL, 1991

Onchorynchus mykiss 0 ;4,11 + pH 4.0 Reduction in oxygen transport Neville, 1985 Loss of sodium & chloride ions

0 + pH 4 - 43 Increase in heart rate Milligan & Wood, 1982 Increase in mean arterial blood pressure Increase in haematocrit Plasma acidosis Reduction in plasma ions Redistribution of body water Haemoconcentration causes large increase in blood viscosity pg/1 + pH 43 Continued coughing indicates harmful Neville, 1985 irritant 0 + pH 4.7 Increase in blood glucose Brown, MacLatchy, Decrease of chloride ions Hara & Eales, 1990 Decrease of osmolality 1.6 p.M + pH 6.1 Aluminium gradually detected as an Neville, 1985 irritant 60 pg/I + pH 5.0 Increase in haematocrit Witters, Puymbroeck & Decrease in plasma sodium concentration Vanderborght, 1991 Increase in plasma cortisol Severe blood acidosis Decrease in blood PO2 Increase in epinephrine & norepinephrine Elevated levels of glucose

27 p.gli + pH 5.2 Severe decline in chloride ions Reid, McDonald & Increase in protein concentration Rhem, 1991 Increase in haematocrit

2.8 AM + pH 4 Toxicity of acid reduced rather than Neville, 1985 increased as aluminium accumulated on gill tissue

2.8 p.M + pH 43 Severe tissue damage Neville, 1985 Major disturbances in oxygen uptake efficiency and ionoregulation

2.8 p.M + pH 5 Impaired oxygen uptake Neville, 1985 Increased ventilation rate Slight loss of electrolytes due to increased permeability of gill epithelium

8-5 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-1 (CONTINUED)

Onchorynchus mylciss 2.8 pld + pH 6.1 Impaired oxygen uptake despite Neville, 1985 tremendous increase in ventilation effort 2.8 + pH 63 Gradually developing, mild, ventilatory Neville, 1985 response 333 µM + pH 5 Severe electrolyte loss Neville, 1985

Salmo salar 235 + pH 5.0 Died within period of 80 hours Polio & Muria, 1993 Mortality increased with increasing temperature Higher ventilation frequency Loss of plasma chloride and sodium Increase in haematocrit concentration (1, 6 & 10'C) At 6 'V, the mean plasma osmolality increased

8.2.1 EXPERIMENTAL PROCEDURE

8.2.1.1 EXPOSURE OF THE TEST ORGANISM Oreochromis mossambicus (Fig 8-1) was chosen as the test organism as it is easy to obtain and to keep and breed in captivity. Adult 0. mossambicus of both sexes were obtained from the University of Zululand near Empangeni, Kwazulu-Natal, and were transported to the Rand Afrikaans University where they were maintained in the aquarium in a recirculating system (Fig 8-2) in borehole water. The quality of the borehole water is presented in Table 8-2. On arrival, the fish were treated against infections by dissolving ± 200 g of coarse salt and 5g per seven kilograms of body mass Terravit (a pfizer antibiotic product) in the water. After three months, the fish were transferred to a flow through system (Fig 8-3). Each series of this system consisted of four glass tanks (volumes of each tank in Table 8-3), of which series A (illustrated) and series B were used for the exposure fish groups and series C was used for the control

8-6 Haematology, Osmoregulation and Carbohydrate Metabolism

FIG 8-1 THE TEST ORGANISM, THE MOZAMBIQUE TILAPIA, OREOCHROMIS MOSSAMBICUS

8-7 Haematology, Osmoregulation and Carbohydrate Metabolism

FIG 8-2 TIM RECIRCULATING SYSTEM CONSISTING OF TWO 1000 LURE PORTAPOOLS (A) AND A BIOLOGICAL FILTER (B)

8-8 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-2 WATER QUALITY OF THE BOREHOLE WATER USED FOR THE EXPERIMENTS

pH• 7.2

m . . ' , Conductivity ni 15.6 49 ,, . ,,,,,,,,, , • , • : : :: taI ardhess=:i 79 TEE ==alciurn. 12 Magnesium .• 8

othumw 4 Potassium 1.9 g/1) :11! 6 ulphate 17 itrate 1.96

TABLE 8-3 VOLUMES OF TANKS USED DURING EXPOSURES

ume,

8-9

Haematology, Osmoregulation and Carbohydrate Metabolism

I , ,,...... )-.. r C - t - . • ' \ .7 0, "$---1 \ \‘'-', ==1:4 ■ -7- '‘ i : ‘: ‘.; •er SYSTEM ' LC, Alma I A UGH

! \ •■•-7-eQ-"*"7 RO TH r 4 1 W-

I FLO 1"...1 ■ i L 1 I 7 ' i I

;I 1 1 • ! i ° I i i • .1 ' I EXPERIMENTA E TH OF M RA DIAG

8-10 Haematology, Osmoregulation and Carbohydrate Metabolism

fish group. The existing flow through system, as used by Nussey (1994), was upgraded by replacing the 220 litre reservoir tanks with 1000 litre glass reservoir tanks, which reduce the chance of the test solutions adsorbing to the sides of these glass tanks. The larger volumes of these tanks also saved time, as it supplied water to the systems at a constant flow rate for two days, in contrast with the smaller 220 litre tanks, where the water had to be changed every day. The experimental tanks were also placed on the same level, which improved light distribution to all the tanks.

After a one week acclimation period, the test solutions (0.06mg/I; 1mg/1; 1.5mg/1 and 2 mg/1 aluminium) were added directly to the glass tanks containing the fish (Fig 8-4 & Fig 8-5), after which a continuous supply of the specified concentrations was maintained by pumping the test solutions from each of three 1000 litre reservoir tanks (Fig 8-6) to each series of glass tanks. The pH was maintained using a Blackstone BL 7016 pH controller & pump. This pump system consists of a pH meter coupled to a pump, supplying a 1 % H2SO4 solution to the reservoir tank until the desired pH is reached. The rate of flow was regulated to each tank throughout the exposure time of 96 hours. Excess water left the system through the outlet pipe.

Aluminium was administered as aluminium chloride (A1C13.6H20, MW = 241.45), supplied by SAARCHEM. The AlC13 was dissolved in the borehole water of the experimental tanks to which the fish were acclimatised prior to addition. Water samples were also taken to determine the actual aluminium concentration in the water. These samples (50 ml) were acidified in the laboratory with perchloric acid (70%) and nitric acid (55%) in a 1:2 ratio. The mixture was concentrated on a hot plate to 25 ml, and then made up to 50 ml with distilled water, after which the total aluminium concentration was determined with a Varian atomic absorption spectrophotometer (Spectra AA10). The analytical standards for aluminium were prepared from Holpro stock solutions (Chapter 3). Table 8-4 displays the aluminium concentrations added and

8-11 Haematology, Osmoregulation and Carbohydrate Metabolism

FIG 8-4 THE EXPERIMENTAL FLOW THROUGH SYSTEM EACH SERIES CONSISTS OF 4 GLASS TANKS (1A - 1D). WATER ENTERS THE TANKS THROUGH A SUPPLY PIPE (2) AND THE WATER FLOW IS REGULATED BY A REGULATING TAP (3). THE TANKS ARE COVERED WITH NETS (4) TO PREVENT THE FISH FROM JUMPING OUT

8-12 Haematology, Osmoregulation and Carbohydrate Metabolism

FIG 8-5 THE EXPERIMENTAL FLOW THROUGH SYSTEM THE SCREENING PIPE (5), PREVENTING CLOGGING OF THE OUTLET PIPE, HAS NICKS (6), CAUSING A SUCKING ACTION THAT TRANSPORTS EXCRETION INTO THE OUTLET PIPE (7), WHICH LEADS TO THE DRAINAGE PIPE (8) THAT TRANSPORTS WATER TO THE BIOLOGICAL FILTER (9) OR A DRAIN. THE SUBMERSIBLE ELECTRIC PUMP (10) SUPPLIES THE WATER TO THE FLOW THROUGH SYSTEM THOUGH THE SUPPLY PIPE

8-13 Haematology, Osmoregulation and Carbohydrate Metabolism

m 8-6 THE EXPERIMENTAL SYSTEM THE 1000 LITRE RESERVOIR TANK (11) USED DURING EXPOSURES WITH A SUBMERSIBLE ELECTRIC PUMP (LITTLE GIANT-MODEL 2E) SUPPLYING THE LOW PH AND ALUMINIUM WATER TO THE EXPERIMENTAL TANKS THROUGH A SUPPLY PIPE. THE PH OF THE WATER IS MAINTAINED BY A PH CONTROLLER AND PUMP CONSISTING OF A PH METER (12A) AND PROBE (12B) AND A PUMP (13), FROM A 1 % SULPHURIC ACID RESERVOIR (14) THROUGH AN ACID

8-14 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-4 SUBLETHAL ALUMINIUM CONCENTRATIONS ADMINISTERED TO THE WATER DURING EXPOSURES AND CONCENTRATIONS DETERMINED THROUGH ATOMIC ABSORPTION SPECTROPHOTOMETRY

Exposure time. -:.measure e,NVater, Control 96 hours * * * pH 5.2 96 hours * * * Short-term (A) 96 hours 0.54 0.06 Mean ±Sd 0.058 -1.- 0.001 Min/max 0.056 - 0.061 Short-term (B) 96 hours 8.95 1.00 Mean ± Sd 0.97±0.007 Min/max 0.96 - 1.05 Short-term (C) 96 hours 13.42 1.50 Mean ± Sd 1.48±0.005 Min/max 1.45 - 1.55 Short-term (D) 96 hours 17.90 2.00 Mean ± Sd 1.94±0.007 Min/max 1.93 - 2.07

A1C13.6H20 was not added to the controls and pH 5.2 exposures and concentrations are not available

the actual aluminium concentrations in the water as determined by atomic absorption spectrophotometry.

The aluminium concentrations were calculated as follows:

Al concentration (mg,/1) = AAS reading (mg/1) x final volume Initial volume (50 ml)

8-15 Haematology, Osmoregulation and Carbohydrate Metabolism

8.2.1.2 CONTROLS Controls are an integral part of any toxicity test and it is important that a control test should be run for every toxicity test, because without the knowledge of what is normal, it is difficult to differentiate between the normal and pathological state. The control tests were thereby performed after the acclimation period, and were treated in the same way as the experimental organisms, except that the fish were kept in borehole water, without any aluminium or acid added.

8.2.1.3 BLOOD SAMPLING To ensure minimum stress, the fish were removed individually from the tanks with a handnet, the lengths were determined on a measuring board and the mass by means of an electric balance (Table 8-5). The eyes were covered (Fig 8-7) when the blood was drawn from the caudal aorta with a one millilitre plastic syringe and a 26G needle, rinsed with heparin as an anticoagulant, to prevent the fish from struggling.

8.2.1.4 MEASUREMENT OF VARIABLES According to Silbergeld (1972), the variables chosen to be measured, are only acceptable as indicators if: they are directly affected by the toxicant to which the fish is exposed the baseline measurement for control conditions can be identified and replicated for the species they show a significant change when the fish is exposed to the toxicant effects on the variable due to the conditions of measurement (capture, handling) are separable from effects due to toxicant exposure The haematological, osmoregulatory and metabolic variables measured in this part of the study, as well as the means of measurement are shown in Table 8-6.

8-16 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-5 MEAN MASS AND LENGTHS OF OREOCHROMIS MOSSAMBICUS, EXPOSED TO LOW PH AND SUBLETHAL CONCENTRATIONS OF ALUMINIUM

40 68±15 168±20 46 - 96 120 - 201

16 69±16 171±12 49 - 98 150 - 193

16 75±16 176± 13 47 - 112 149 - 198

16 78±35 166±10 47 - 160 150 - 180

16 78-±16 16914:11 46 - 89 150 - 165

16 74±10 164± 11 48 - 79 148 - 179

8-17 Haematology, Osmoregulation and Carbohydrate Metabolism

FIG 8-7 BLOOD SAMPLING FROM THE CAUDAL AORTA OF OREOCHROMIS MOSSAMBICUS, WITH A 1ML PLASTIC SYRINGE RINSED WITH HEPARIN

8-18 Haematology, Osmoregulation and Carbohydrate Metabolism

TABLE 8-6 THE HAEMATOLOGICAL, OSMOREGULATORY AND METABOLIC VARIABLES MEASURED IN THIS STUDY

" ..: :1:H!!!!''!:::':''"' ' .„J::::• ' , , : ' _ .: s , - _ : .•:,:e 1) White blood cells (WBC) 1) Sysmex CC-120 microcell counter 2) Red blood cells (RBC) Sysmex CC-120 microcell counter Mean corpuscular volume (MCV) 3) Sysmex CC-120 microcell counter 4) Haematocrit concentration (Htc) 4) Sysmex CC-120 microcell counter 5) Haemoglobin concentration (Hb) 5) Sysmex CC-120 microcell counter 6) Osmolality (Osmol) 6) Osmomat 030 Cryoscopic osmometer 7) Plasma sodium concentration (Na) 7) Radiometer FLM3 flamephotometer 8) Plasma potassium concentration(K) 8) Radiometer FLM3 flamephotometer 9) Plasma chloride concentration (Cl) 9) Corning Chloride Analyzer 925 10) Plasma calcium concentration (Ca) 10) Corning Analyzer 940 11) Choline Esterase activity (CHE) 11) Boehringer- Mannheim Test kits 12) Pyruvate Kinase activity (PK) 12) Boehringer- Mannheim Test kits 13) Glucose-6-Phosphate 13) Boehringer- Mannheim Test kits dehydrogenase activity (G-6-P) 14) Lactate concentration (Lac) 14) Boehringer- Mannheim Test kits 15) Blood glucose concentration 15) Boehringer- Mannheim Test kits (Gluc)

8-19 Haematology, Osmoregulation and Carbohydrate Metabolism

8.2.1.5 DATA PROCESSING A STATGRAPHICS statistical programme was used to process the data. Independent Student's t-tests were performed to prove the probability hypotheses and differences in mean values were accepted as being statistically significant if 0.005

8.3 RESULTS

Exposure to pH 5.2 and combinations of low pH and different aluminium concentrations, showed a variety of differences in the blood of 0. mossambicus (Table 8-1 Appendix). These differences are presented in Fig 8-8 - Fig 8-12, with the first group representing values compared to the control values and the second group compared to the pH 5.2 values, indicating the differences in significance levels of the differences between the two groups.

After these exposures the red blood cell count and the mean haemoglobin concentrations showed only slight changes, compared to both the control group and the pH 5.2 group (Fig 8-8, Table 8-1 Appendix). The haematocrit concentration and mean corpuscular volume were significantly higher after exposure to 1.5 mg/1 aluminium at pH 5.2, compared to the control group, while the mean corpuscular volume also showed a significant increase after exposure to 1 mel aluminium at pH 5.2 (Fig 8-8, Table 8-1 Appendix). The haematocrit concentration also showed a significant higher value after exposure to 1.5 mg/1 aluminium at pH 5.2 compared to the pH 5.2 group.

8-20

Haematology, Osmoregulation and Carbohydrate Metabolism

A Control pH 5.2 C 0.06 mg/I at pH 5.2 1 mg/I at pH 5.2 1.5 mg/I at pH 5.2 F 2 mg/I at pH 5.2

FIG 8-8 THE MEAN RED BLOOD CELL COUNT (X106)MM4 (A), HAEMOGLOBIN CONCENTRATION (g/d1) (B) AND HAEMATOCRIT CONCENTRATION (%) (C) OF OREOCHROMLS MOSSAMBICUS AFTER EXPOSURE TO PH 5.2 AND 0.06 MG/L, 1 MG/L, 1.5 MG/L AND 2 MG/L ALUMINIUM AT PH 5.2. THE FIRST GROUP REPRESENTS THE VALUES AFTER EXPOSURE COMPARED TO THE VALUES OBTAINED FOR THE CONTROL GROUP, WHILE THE SECOND GROUP REPRESENTS THE VALUES COMPARED TO THE VALUES OBTAINED AFTER EXPOSURE TO THE PH 5.2 GROUP, INDICATING DIFFERENCES IN THE SIGNIFICANCE LEVELS IN COMPARISONS BETWEEN THE TWO GROUPS SIGNIFICANCE LEVEL: * 0.005

8-21 Haematology, Osmoregulation and Carbohydrate Metabolism

Compared to the control group, a significant increase (P5.0.05) in the mean corpuscular volume occurred after exposure to 1.5 mg/1 aluminium at pH 5.2; a significant decrease occurred after exposure to 1 mg/1 aluminium at pH 5.2 and a slight decrease occurred after exposure to pH 5.2; 0.06 & 2 mg/1 aluminium at pH 5.2. Compared to the pH 5.2 exposure group, the addition of 1.5 mg/1 aluminium, caused a significant increase in the mean corpuscular volume, while an insignificant decrease occurred after addition of 0.06, 1 & 2 mg/1 aluminium (Fig 8-9, Table 8-1 Appendix).

After exposure to 0.06 mg/1, 1 mg/1 & 1.5 mg/1 aluminium at pH 5.2, there was a definite increase in the white blood cell count, with the highest value found at 1.5 mg/1 aluminium at pH 5.2 (Fig 8-9, Table 8-1 Appendix). There was however no significant change in the white blood cell count after exposure to 2 mg/I aluminium at pH 5.2.

The osmolality showed a significant decrease after exposure to 1 mg/1, 1.5 mg/I & 2 mg/I aluminium at pH 5.2 (Fig 8-9, Table 8-1 Appendix). The only significant decrease in osmolality, compared to the values obtained after exposure to pH 5.2, was after addition of 2 mg/1 aluminium. The plasma potassium concentration increased significantly after exposure to 1mg/1 aluminium at pH 5.2 and after exposure to 2 mg/I aluminium at pH 5.2, the concentration was significantly higher than the value obtained after exposure to pH 5.2 (Fig 8-10, Table 8-1 Appendix). On the other hand, the plasma sodium concentration was significantly lower than the concentrations for the control group, after exposure to 1 mg/1 and 2 mg/1 aluminium at pH 5.2 (Fig 8-10, Table 8-1 Appendix). The mean plasma chloride concentration was significantly lower than the values obtained for the control group as well as the pH 5.2 group after addition of 1 mg/1 aluminium, but increased again with higher concentrations of aluminium (Fig 8-10, Table 8-1 Appendix). The plasma calcium concentration increased significantly after exposure to pH 5.2 and combinations of 0.06 mg/1 & 1 mg/1 aluminium at pH 5.2, but a significant decrease was found after exposure to 2 mg/I aluminium at pH 5.2 (Fig 8-11, Table 8-1 Appendix).

8-22

Haematology, Osmoregulation and Carbohydrate Metabolism

A Control pH 5.2 C 0.06 mg/I at pH 5.2 1 mg/1 at pH 5.2 1.5 mg/I at pH 5.2 F 2 mg/I at pH 5.2

FIG 8-9 THE MEAN CORPUSCULAR VOLUME ( //m 3) (A), THE MEAN WHITE BLOOD CELL COUNT (X103) mm3 (B) AND THE OSMOLALITY (OSMOL/KG) (C) OF OREOCHROMIS MOSSAMBICUS AFTER EXPOSURE TO PH 5.2 AND 0.06 MG/L, 1 MG/L, 1.5 MG/L AND 2 MG/L ALUMINIUM AT PH 5.2. THE FIRST GROUP REPRESENTS THE VALUES AFTER EXPOSURE COMPARED TO THE VALUES OBTAINED FOR THE CONTROL GROUP, WHILE THE SECOND GROUP REPRESENTS THE VALUES COMPARED TO THE VALUES OBTAINED AFTER EXPOSURE TO THE PH 5.2 GROUP, INDICATING DIFFERENCES IN THE SIGNIFICANCE LEVELS IN COMPARISONS BETWEEN THE TWO GROUPS SIGNIFICANCE LEVEL: * 0.005

8-23

Haematology, Osmoregulation and Carbohydrate Metabolism

A Control pH 5.2 C 0.06 mg/I at pH 5.2 1 mg/I at pH 5.2 1.5 mg/I at pH 5.2 F 2 mg/I at pH 5.2

FIG 8-10 THE MEAN PLASMA POTASSIUM CONCENTRATION (MMOIJL) (A), MEAN PLASMA SODIUM CONCENTRATION (MMOL/L) (B) AND THE MEAN PLASMA CHLORIDE CONCENTRATION (MMOL/L) (C) OF OREOCHROMIS MOSSAMBICUS AFTER EXPOSURE TO PH 5.2 AND 0.06 MG/L, 1 MG/L, 15 MG/L AND 2 MG/L ALUMINIUM AT PH 5.2. THE FIRST GROUP REPRESENTS THE VALUES AFTER EXPOSURE COMPARED TO THE VALUES OBTAINED FOR THE CONTROL GROUP, WHILE THE SECOND GROUP REPRESENTS THE VALUES COMPARED TO THE VALUES OBTAINED AFTER EXPOSURE TO THE PH 5.2 GROUP, INDICATING DIFFERENCES IN THE SIGNIFICANCE LEVELS IN COMPARISONS BETWEEN THE TWO GROUPS SIGNIFICANCE LEVEL.: * 0.005

8-24 Haematology, Osmoregulation and Carbohydrate Metabolism

800 •• 700

800 ••

500 Q

400 0 Si •• t. 300 ••

200

100

0 ABCDEF BCDEF

A Control pH 5.2 C 0.06 mg/I at pH 5.2 1 mg/1 at pH 5.2 1.5 mg/1 at pH 5.2 F 2 mg/I at pH 5.2

FIG 8-11 THE MEAN PLASMA CALCIUM CONCENTRATION (MG/%) (A), MEAN BLOOD GLUCOSE CONCENTRATION (MG/100ML) (B) AND THE PYRUVATE KINASE ACTIVITY (MU/ML) (C) OF OREOCHROMIS MOSSAMBICUS AFTER EXPOSURE TO PH 5.2 AND 0.06 MG/L, 1 MG/L, 13 MG/L AND 2 MG/L ALUMINIUM AT PH 5.2. THE FIRST GROUP REPRESENTS THE VALUES AFTER EXPOSURE COMPARED TO THE VALUES OBTAINED FOR THE CONTROL GROUP, WHILE THE SECOND GROUP REPRESENTS THE VALUES COMPARED TO THE VALUES OBTAINED AFTER EXPOSURE TO THE PH 5.2 GROUP, INDICATING DIFFERENCES IN TIM SIGNIFICANCE LEVELS IN COMPARISONS BETWEEN THE TWO GROUPS SIGNIFICANCE LEVEL: * 0.005

8-25 Haematology, Osmoregulation and Carbohydrate Metabolism

The difference in concentrations and activities of variables involved in the metabolism of 0. mossambicus were predominantly significant. The glucose concentration increased significantly after exposure to pH 5.2 and 0.06 mg/I & 1 mg/1 aluminium at pH 5.2, but compared to the pH 5.2 group, the glucose concentration decreased significantly with the addition of 0.06 mg/1, 1.5 mg/1 & 2 mg/I aluminium (Fig 8-11, Table 8-1 Appendix). Exposure to 1 mg/1, 1.5 mg/1 & 2 mg/I aluminium at pH 5.2, caused a significant decrease in the pyruvate kinase activity (Fig 8-11, Table 8-1 Appendix).

The lactate concentration was significantly higher than the control group values, after exposure to pH 5.2, 0.06 mg/1, 1.5 mg/1 & 2 mg/1 aluminiuma at pH 5.2, but showed a significant decrease if compared to the pH 5.2 group with the addition of 1 mg/1 aluminium (Fig 8-12, Table 8-1 Appendix). The glucose-6-phosphate dehydrogenase activity increased significantly after all the exposures had been performed (Fig 8-12, Table 8-1 Appendix), with very high values after exposure to 1 mg/1 aluminium at pH 5.2. The choline esterase activity decreased significantly after exposure to pH 5.2 and 1.5 mg/1 & 2 mg/1 aluminium at pH 5.2 (Fig 8-12, Table 8-1 Appendix).

8.4 DISCUSSION

The use of haematological parameters as indicators of sublethal effects of stress can provide information on the physiological responses of fish to changing external environments. This is a result of the close association of the circulatory system of the fish with the external environment and can be used as a screening method to assess the state of health of fish in fish biology research. Different blood characteristics in the present study showed positive or negative relationships with environmental variables and from the results, it is clear that some degree of chemical stress had been imposed upon

8-26 Haematology, Osmoregulation and Carbohydrate Metabolism

• 8

0 6

• 2 •

0 if! ABCDEF B CD E F

100

•

80 A Control pH 5.2 C 0.06 mg/I at pH 5.2 60 1 mg/I at pH 5.2 1.5 mg/I at pH 5.2 Y

0 F 2 mg/I at pH 5.2 .s• •• 40 •

20 •

0 ABCDEF BCDEF

FIG 8-12 THE MEAN LACTATE CONCENTRATION (MG/100ML) (A), GLUCOSE-6-PHOSPHATE DEHYDROGENASE ACTIVITY (MU/ML) (B) AND THE CHOLINE ESTERASE ACTIVITY (U/L) (C) OF OREOCHROMIS MOSSAMBICUS AFTER EXPOSURE TO PH 5.2 AND PH 5.2 ♦ 0.06 MG/L, 1 MG/L, 15 MG/L AND 2 MG/L ALUMINIUM. THE FIRST GROUP REPRESENTS THE VALUES AFTER EXPOSURE COMPARED TO THE VALUES OBTAINED FOR THE CONTROL GROUP, WHILE THE SECOND GROUP REPRESENTS THE VALUES COMPARED TO THE VALUES OBTAINED AFTER EXPOSURE TO THE PH 5.2 GROUP, INDICATING DIFFERENCES IN THE SIGNIFICANCE LEVELS IN COMPARISONS BETWEEN THE TWO GROUPS SIGNIFICANCE LEVEL: * 0.005

8-27 Haematology, Osmoregulation and Carbohydrate Metabolism

the exposed fish. Stress can be defined as the sum of all the physiological responses by which an organism tries to maintain or compensate for a normal metabolism in the face of a physical or chemical force (Pickering, 1981). The reaction to stress can be divided into the alarm reaction, where the body is informed that foreign substances are in the immediate environment; the resistance reaction where adaption to the foreign substance takes place by reaching homeostaces; and the exhaustion reaction, where the stress cannot be coped with and the adaption to the stress is lost, and this can lead to the death of the organism.

An important feature common to the aluminium and acid toxicity to fish, also observed during the exposure of the fish to a low pH (pH 5.2) and the metal, is the appearance of mucus on the gills, as well as an increase in body mucus production. This increase is associated with mucus discharge from goblet cells on the gills and body (Lock & van Overbeeke, 1981; Eddy & Fraser, 1982). Mucus acts as a diffusion barrier to oxygen, contributing to the hypoxia observed when the fish were exposed. Precipitation of aluminium hydroxides or binding of aluminium to organic anions occurs on the gill surface, inducing an inflammatory response which stimulates mucus production, thickens and distorts the branchial epithelium, decreases transcellular permeability to 0 2 and CO2, yet simultaneously increases the permeability of paracellular channels through which the majority of electrolyte loss occurs (McDonald, 1983).

EFFECTS ON THE HAEMATOLOGY Red blood cells (erythrocytes) are produced in the haemopoietic tissue (Hoffbrand & Pettit, 1980; Smith, 1982) situated in the spleen and head kidney (Heath, 1987). The most important function of the red blood cells is that they contain haemoglobin which enables them to transport oxygen to all tissues in the body (Hoffbrand et al., 1980). Changes observed at pH 5.2 and combinations of pH 5.2 and different aluminium

8-28 Haematology, Osmoregulation and Carbohydrate Metabolism

concentrations, were statistically not significant. A low red blood cell count can be a result of inhibited production and increased destruction of erythrocytes (Narain & Srivastava, 1989). It could also be attributed to impaired osmoregulation and gill damage during exposure, which results in haemodilution which leads to a decrease in the number of red blood cells through haemolysis. This causes the fish to become anaemic (Wedemeyer & Yasutake, 1977). The development of anaemia in fish can also be attributed to the cell membrane of the red blood cells being altered through the hydrolysis of acetylcholine in the body fluids, by acetylcholine esterase. Red blood cells that aggregate in the gills can cause reduction in the number of circulating red blood cells in the stressed fish (also found by Narain & Srivastava, 1989).

Haemoglobin is a sophisticated oxygen delivery system that provides the desired amounts of oxygen to the tissue under a wide variety of circumstances and. Low haemoglobin levels signal anaemia. Very small organisms do not require a protein such as haemoglobin, because their respiratory needs are satisfied by the simple passive diffusion of oxygen through their bodies. However, since the transport rate of a diffusing substance varies inversely with the square of the distance it must diffuse, the oxygen diffusion rate through tissue thicker than 1mm is too slow to support life (Voet & Voet, 1990). Therefore, animals such as fish require a circulatory system that actively transports oxygen and nutrients to the tissues. The blood of such animals must therefore contain an oxygen transporter such as haemoglobin, because the solubility of oxygen in blood plasma is too low ( 10 4M under physiological conditions) to carry sufficient oxygen for metabolic needs. Many pollutants enter the red blood cell and either oxidize or denature the haemoglobin. The data obtained in this study showed a slight decrease to 15.3 01 in the haemoglobin concentration after exposure to pH 5.2, but in contrast to significant decreases in the haemoglobin concentration in the blood of Tilapia sparrmand after exposure to sublethal concentrations of hexavalent chromium (Wepener, Van Vuren & Du Preez, 1992a). With the addition of 1 mg/l aluminium, the haemoglobin

8-29 Haematology, Osmoregulation and Carbohydrate Metabolism

concentration increased (high of 18.3 g/dl), which may be due to an increased production by the erythropoietic tissues to elevate the oxygen capacity of the blood. This is a process whereby the body produces an increased amount of haemoglobin to replace the oxidized or denatured haemoglobin, formed as a result of metal exposure.

Haematocrit is an instrument for determining the amount of plasma and corpuscles in the blood (measurement of packed erythrocytes) and is used to determine the oxygen carrying capacity of the blood (Larsson, Haux & Sjobeck 1985). Furthermore, it is defined as the volume occupied by erythrocytes in a given volume of blood, usually measured as the number of erythrocytes per 100 ml of blood. The significant increase in the haematocrit concentration found after exposure to 1.5 mg/I aluminium, at pH 5.2 may be attributed to an increase in the number of erythrocytes due to catecholamine- induced mobilization from the spleen, swelling of the erythrocytes due to osmotic shifts (Witters, Vangenechten & Van Puymbroeck, 1987), or a decreased plasma volume caused by fluid shifts (Milligan et al., 1982).

The mean corpuscular volume (MCV) indicates the size or status of the red blood cells and reflects a normal or abnormal cell division during erythropoiesis (Larsson et al., 1985). The significant decrease (P.0.05) in the mean corpuscular volume after exposure to 1 mg/1 aluminium at pH 5.2, is in contrast to increases in the mean corpuscular volume in 0. mossambicus exposed to sublethal copper concentrations (Nussey, Van Vuren & Du Preez, 1995) and suggests that the erythrocytes may have shrunk, either due to hypoxia, stress or impaired water balance, or a large concentration of smaller, immature erythrocytes that have been released from the erythropoietic tissue to counteract the pathological action of aluminium (Larrson et al., 1985). The significant (P:50.05) increase in the mean corpuscular volume after exposure to 1.5 mg/I aluminium at pH 5.2, on the other hand, may be due to swelling of the erythrocytes due to impaired water balance (osmotic stress) or due to hypoxia setting in.

8-30 Haematology, Osmoregulation and Carbohydrate Metabolism

The main function of the circulating white blood cells is the immune reaction of the body against pathogens. The white blood cells of fish include four types of cells, namely, the lymphocytes, which form antibodies during the immunological process, the granulocytes and monocytes, which take part in the elimination of foreign substances, and trombocytes which play an important role during blood coagulation. An increase in the number of white blood cells, as seen after exposure to pH 5.2 and 0.06 mg/I, 1 mg/I & 1.5 mg/1 aluminium at pH 5.2, is therefore a normal reaction of the body against foreign substances that alter the normal physiological processes of the fish. It is also known that metals can block the active area of the antibody molecules, which may cause a disturbance in the metabolism, ionic balance and cell division of these cells.

EFFECTS ON THE OSMOREGULATION The results of this study confirm previous findings (Muniz & Leivestad, 1980; Neville, 1985) in showing that low ambient pH alone and in combination with aluminium exposure produce disturbances to ion balance. The data indicate that ionic fluxes and plasma ion concentrations become more severely disrupted with increased concentrations of either H+ or aluminium. It is important for the fish to be able to maintain water and ion homeostasis and thereby survive in a changing environment. A disturbed osmotic and ion regulation may affect the ability of the fish to function normally. Critical loss of body electrolytes, which reduces the osmotic concentration, leads to the loss of water, which leads to haemoconcentration and circulatory collapse. Therefore survival through acute challenges relies upon the fish's ability to reduce ion efflux while restoring sufficient uptake to maintain body ion levels.

In some cases, the effect of low pH is exacerbated by the presence of aluminium. As a result of a change in pH in the branchial micro-environment, aluminium accumulates and precipitates (Booth, McDonald, Simons & Wood, 1988) on the gills as a result of the

8-31 Haematology, Osmoregulation and Carbohydrate Metabolism

forming of hydroxides with lower solubility. In this way, aluminium could act as an irritant on the gill surface, which stimulates mucus production and decrease the branchial permeability for CO2 and 02 exchange. The distortion of the gill epithelium may contribute to greater permeability of the paracellular channels and therefore the elevated loss of electrolytes (Witters et aL, 1987).

Changes in osmoregulation in stressed fish are elucidated by measuring the blood plasma sodium, potassium, calcium, chloride and osmolality (Heath, 1987). The sodium, potassium and chloride ions are responsible for the preservation of the osmolarity and crystalloid osmotic pressure of the plasma (Meyer, 1988b). The gills, kidneys and to some extent the integument (in its role as a barrier), regulate the plasma ion concentrations (Bond, 1979). The plasma potassium concentration increased significantly (P50.05) after exposure to 1 mg/1 aluminium at pH 5.2, despite of the branchial losses (also found by Booth et al., 1988), as commonly observed in acid stressed fish. This reflects both haemoconcentration and an efflux of potassium ions from the intracellular compartment of white muscle (Wood & McDonald, 1982). The increase in plasma potassium can also be as a result of the exchanging of potassium ions with the hydrogen ions at cellular level to get rid of the excess hydrogen ions. In contrast to this increase in plasma potassium concentrations and the increase found by Nussey et al. (1995) after exposure of 0. mossambicus to sublethal levels of copper, the decrease is due to impaired active ion uptake via the gills or the intestine, or impaired renal ion retention.

The plasma sodium concentration remained stable up to a concentration of 1 mg/I aluminium at pH 5.2. After exposure to 1.5 mg/1 & 2 mg/1 aluminium at pH 5.2, the sodium concentration decreased significantly, also found after longterm exposure of 0. mossambicus to sublethal concentrations of copper. This may be due to the loss of sodium by the gills and kidneys, as structural changes to the gills can cause the excretion of sodium. The uptake of sodium from the environment does not take place

8-32 Haematology, Osmoregulation and Carbohydrate Metabolism

as a result of the inhibition of the Na-, K-ATPase enzyme in the gills (Heath, 1987). The damaging of the gills similarly prevents reabsorption of sodium. The Na-, K-ATPase enzymes play an important role in the kidney, where glomerular filtration and reabsorption play an important role in osmoregulation. Reabsorption of sodium in the renal tubes is also decreased as a result of the inhibition of Na-, K-ATPase in the kidney and intestinum (Kuhnert, Kuhnert & Stokes, 1976). The plasma chloride concentration decreased significantly after exposure to 1 mg/I aluminium at pH 5.2, which may be attributed to the nett loss of chloride by the damaged gills.

Calcium functions in ion regulation, membrane permeability, muscle and nerve cell function, skeletal bone metabolism and blood coagulation. It has been reported that the toxicity of metals is counteracted by calcium and other antagonistic metallic cations (Lewis & Lewis, 1971), as Ca' protects largely by reducing ion loss. Significant increases in plasma calcium concentrations were noted after exposure to pH 5.2 and 0.06 mg/1 & 1 mg/1 aluminium at pH 5.2, which can be due to tissue damage, as well as the secretion of parathormone which could also withdraw calcium from bone, which serves as a calcium reservoir and thus increase the plasma concentration. Decreases in plasma calcium concentration, as found after exposure to 2 mg/I aluminium at pH 5.2, could be due to impaired tubular reabsorption in the kidney, or an impaired intestinal uptake of calcium (Larsson et aL, 1981).

EFFECTS ON THE CARBOHYDRATE METABOLISM AND ENZYME ACTIVITY Fish respond to exposure to pollutants by a series of biochemical and physiological stress reactions. These responses are mediated by enhanced releases of hormones from the pituitary, the interrenals and the chromaffin tissues and by an increased activity of the sympathetic nervous system (Mazeaud, Mazeaud & Donaldson, 1981). Besides the effects on haematology and osmoregulation, the carbohydrate metabolism is also

8-33 Haematology, Osmoregulation and Carbohydrate Metabolism

affected. The generation of metabolic energy from carbohydrates begins with glycolysis (Fig 8-13), a ten-step, enzymatically catalyzed pathway, which converts one molecule of glucose to two molecules of pyruvate, with the concomitant generation of two molecules of ATP. The overall reaction being:

Glucose + 2NAD+ + 2ADP + 2P; 2NADH +2 pyruvate + 2ATP +2H20 + 4H+

Glucose occupies a central role in metabolism, both as a fuel and as a precursor of essential structural carbohydrates and other biomolecules. The significant increase in blood glucose levels after exposure to pH 5.2, therefore reveals a disruption in carbohydrate metabolism, which is probably a hypophysis-adrenal response (Christensen, McKim, Brungs & Hunt, 1972). This is a stress response, involving an increase in blood sugar with the eventual secretion of glucocorticoids and catecholamines (Nath & Kumar, 1987). Catecholamines may deplete glycogen reserves in stressed fish by stimulating glucogenolysis and gluconeogenesis (Heath, 1987). Exposure to pH 5.2 caused a significant increase in the blood glucose levels. Exposure to 0.06 mg/1 aluminium at pH 5.2, however, showed only a slight increase in the blood glucose levels. It seems as if at this low aluminium concentration, the toxicity of the acid was reduced rather than increased. This was also found by Neville (1985), with exposure of juvenile rainbow trout, Oncorhynchus mykiss exposed to pH 4 and 2.8 AM inorganic aluminium, where it appeared that the low concentration of aluminium accumulated on the gill tissue at that pH, afforded some protection to the gill epithelium. Exposure to 1 mg/1 & 1.5 mg/I at pH 5.2, once again caused a significant increase in the blood glucose levels, while exposure to 2 mg/I aluminium at pH 5.2 showed no significant change. This could be due to the possibility that the aluminium formed polymers too large to adhere closely to the gill epithelium.

In reaction ten of glycolysis, the enzyme pyruvate kinase (PK) couples the free energy

8-34

Haematology, Osmoregulation and Carbohydrate Metabolism

+ 2ADP (2NAD +2P

Fructose-1.6—biphosphate

2ATP 2NADH

Pyruvate kinase

2 Pyruvate 1 1

Anaerobic homolactic Aerobic oxidation fermentation 2NADH + 2NAD

Citric acid cycle 2 Lactate

Oxidative phosphorylation

2NAD 4011'

6C02 + 6H 0 2

FIG 8-13 THE GENERATION OF METABOLIC ENERGY THROUGH GLYCOLYSIS

8-35 Haematology, Osmoregulation and Carbohydrate Metabolism

of phospho-enolpyruvate hydrolysis to the synthesis of ATP to form pyruvate, which, through gluconeogenesis, is converted back to glucose. The optimum pH for pyruvate kinase is in the range of pH 7.2 - 7.8 (Randall & Anderson, 1975) and it is therefore clear that the addition of 1 mg/1, 1.5 mg/I & 2 mg/1 aluminium exacerbated the effect of the low pH on the pyruvate kinase activity, by causing significant decreases in the activity of this enzyme. These decrease in enzyme activity in the plasma shows a possibility that there may be an increase in the pyruvate kinase activity elsewhere in the fish. That may be an indication that there is an increase in glycolysis in the muscle tissue of the fish (Knox, Walton & Conwey, 1980) The noncarbohydrate precursors that can be converted to glucose, include the glycolysis products lactate and pyruvate, citric acid cycle intermediates and most amino acids.

The significant increase in lactate concentration after exposure to pH 5.2 and 0.06 mg/1, 1.5 mg/I and 2 mg/1 aluminium at pH 5.2, implies an increase in anaerobic metabolism (Heath, 1987), which can occur as a result of gills damaged by aluminium and low pH. This hypoxic state, was also found with exposure to hexavalent chromium, when anaerobic cell respiration takes place with muscle glycogen as fuel (Van Waarde, Van den Thillart & Kesbeke, 1983). Under aerobic conditions, the pyruvate formed by glycolysis is further oxidized by the citric acid cycle and oxidative phosphorylation to CO 2 and water. Under anaerobic conditions, however, the pyruvate is now converted to a reduced end product, which is lactate (homolactic fermentation; Fig 8-13). Much of the lactate is exported from the muscle cells and carried by the blood to the liver, where it is reconverted to glucose. In vertebrates, some of the tissues, such as red blood cells, derive most of their energy from anaerobic metabolism. Skeletal muscle, which derives most of its energy from respiration when at rest, relies heavily on glycolysis during exertion/stress, when glycogen stores are rapidly broken down, or mobilized to provide substrates for glycolysis.

8-36 Haematology, Osmoregulation and Carbohydrate Metabolism

Normally the lactate produced diffuses from the tissue and is transported through the bloodstream to highly aerobic tissues, such as the heart and liver. The aerobic tissue can catabolize lactate further, through respiration, or can convert it back to glucose, through gluconeogenesis. However, if lactate is produced in large quantities, it cannot be readily consumed. Then the blood pH falls and the Bohr effect functions to increase oxygen supplies to the tissues (Voet & Voet, 1990). The decrease of pH in the capillaries lowers the oxygen affinity of haemoglobin, allowing even more efficient release of the last traces of oxygen. The response of haemoglobin to the decrease in pH is called the Bohr effect. The general reaction can be described as: Hb.402 + .1-1+ — Hb. .H+ + 402 where n has a value somewhat greater than 2 (Matthews & Van Holde, 1991).

From a physiological viewpoint, this reaction has two consequences: in the capillaries, hydrogen ions promote the release of oxygen by driving the reaction to the right. When the venous blood then recirculates to the gills, the oxygenation has the effect of releasing the H+ by shifting the equilibrium to the left. This causes the release of CO 2 from bicarbonate dissolved in the blood by the reversal of the bicarbonate reaction (above). The free CO2 can then be expired. Lactate accumulation in animals thus occurs when the need for tissues to generate energy exceeds their capacity to oxidize the pyruvate produced in glycolysis.

In this study, the activity of glucose-6-phosphate dehydrogenase, which is the key regulatory enzyme in the pentose-phosphate pathway, was determined, in order to determine the possible effect of low pH and aluminium on this pathway. In the pentose phosphate pathway, additional energy rich compounds are produced. Glucose-6- phosphate dehydrogenase oxidizes glucose-6-phosphate to fructose-6-phosphate, which is converted to pyruvate during glycolysis, which in turn is converted to glucose. The toxicological action of aluminium is mainly confined to the gills and there may be a

8-37 Haematology, Osmoregulation and Carbohydrate Metabolism

higher need for energy in the gills for the active regulation of the osmotic balance of the fish. There may therefore be an increase in glucose-6-phosphate dehydrogenase activity in the gills to produce this additional energy for the osmoregulatory process (Bhaskar & Govindappa, 1985). This is probably the reason for the increase in glucose-6-phosphate dehydrogenase activity in 0. mossambicus after exposure to pH 5.2 and 0.06 mg/1, 1 mg/I, 1.5 mg/1 and 2 mg/1 aluminium at pH 5.2.

Exposure to low pH and combinations of low pH and different aluminium concentrations, caused little activity in the fish. Nerve impulses are transmitted electrically along a neuron and, in most cases, chemically between neurons and from neurons to muscle or gland cells. Nerve impulses are chemically transmitted across most synapses by the release of neurotransmitters. Such a neurotransmitter is acetylcholine (ACh), which is packaged in synaptic vesicles that are exocytotically released into the synaptic cleft and degrades an acetylcholine molecule in the few milliseconds before the potential arrival of the next nerve impulse. • As can be seen in this study, exposure to pH 5.2 and 1.5 mg/I & 2 mg/1 aluminium at pH 5.2, causes a significant decrease in the acetylcholine esterase activity. When acetylcholine esterase is inhibited, the accumulation of acetylcholine desensitises the membrane receptors, which after an initial period of repetitive firing can no longer respond, so that neuromuscular transmission fails. Nerve impulse transmission is therefore blocked at cholinergic synapses. Locomotor activity would therefore be reduced to such an extent that the fish remained motionless at the bottom of the tank, as observed during the exposures.

8.5 CONCLUSION

It is clear that low pH and sublethal concentrations of aluminium have a physiological

8-38 Haematology, Osmoregulation and Carbohydrate Metabolism

effect on 0. mossambicus. In some cases, however, for example an aluminium concentration of 0.06 mg/1 aluminium and a pH of 5.2, the toxicity of the acid was reduced rather than increased by the aluminium, which was clearly seen in the blood glucose concentration. It seems as if the most toxic concentration of aluminium at a pH of 5.2, was in the range of 1 mg/1 & 1.5 mg/I aluminium. Concentrations greater than this, in this case 2 mg/1 aluminium at pH 5.2, had statistically little effect on the haematology of the fish (white and red blood cell counts, haemoglobin and haematocrit concentrations and the mean corpuscular volume, as well as on the plasma potassium and blood glucose concentrations). It seems that at certain concentrations, the aluminium is toxic to the fish, up to a point where the fish starts to regulate the metal. The different routes for possible excretion of harmful chemicals are the gills, bile (via faeces), kidney and skin. The liver is the main organ for metal homeostasis, where the metal-binding protein metallothionein is of key importance. The metabolism of the fish was however severely affected after exposure to low pH and combinations of low pH and different aluminium concentrations. There was a definite increase in blood glucose and lactate concentrations, which indicates a stress response and anaerobic metabolism respectively. It was also clear that the aluminium exacerbated the effect of the low pH (5.2) on the pyruvate kinase activity, by causing a significant decrease in the activity of the enzyme.

From these results and the aluminium concentrations recorded in the Upper Olifants River Catchment at localities KOR1 and OR1, it is evident that some degree of physiological stress would have been imposed on the fish if the pH is reduced to values smaller than 5.5, making the aluminium more bioavailable to the fish. It is clear from this section of the study that aluminium in combination with low pH is toxic to the fish, but the higher pH values at localities KOR1 and OR1, cause less toxicity of the metal, since the availability of aluminium is hampered by higher pH values.

8-39 Haematology, Osmoregulation and Carbohydrate Metabolism

8.6 REFERENCES

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BOND, C.E. (1979). Biology of fishes. Saunders College publishing, Philadelphia. pp 514.

BOOTH, C.E., McDONALD, D.G., SIMONS, B.P. & WOOD, C.M. (1988). Effects of aluminium and low pH on net ion fluxes and ion balance in the brook trout Salvelinus fontinalis. Can. J. Fish. Aquat. Sci 45:1563 - 1574.

CHRISTENSEN, G.M., McKIM, J.M., BRUNGS, W.A. & HUNT, E.P. (1972). Changes in the blood of the brown bullhead (Ichtaiurus nebuloses Lesueur) following short and log term exposure to copper (II). Toxicol. Appl. PharmacoL 23:417 - 427.

EDDY, F.B. & FRASER, J.E. (1982). Sialic acid and mucus production in rainbow trout (Salmo gairdneri Richardson). Comp. Biochem. PhysioL 73C:357 - 359.

FREEMAN, R.A. & EVERHART, W.H. (1971). Toxicity of aluminium hydroxide complexes in neutral and basic media to rainbow trout. Trans. Am. Fish. Soc. 4:644 - 658.

8-40 Haematology, Osmoregulation and Carbohydrate Metabolism

FROMM, P.O. (1980). A review of some physiological and toxicological responses of freshwater fish to acid stress. Environ. Biol. Fishes 5:79 - 93.

HEATH, A.G. (1987). Water pollution and fish physiology. C.R.C. Press, Inc., Boca Randton, Florida. pp. 245.

HOFH3RAND, A.V. & PEMT, J.E. (1980). Essential Haematology. Blackwell Scientific Publications, Oxford. pp. 227.

INGERSOLL, C.G., SANCHEZ, D.A., MEYER, J.S., GULLEY, D.D. & TIETGE, J.E. (1990). Epidermal response to pH, aluminium and calcium exposure in brook trout (Salvelinus fontinalis) fry. Can. J. Fish. Aquat. Sci 47:1616 - 1622.

KNOX, G., WALTON, M.J. & CONWEY, C.B. (1980). Distribution of enzymes of glycolysis and glyconeogenesis in fish tissues. Mar. Biol. 56:7 - 10.

KUHNERT, P.M., KUHNERT, B.R. & STOKES, R.M. (1976). The effects of in vivo chromium exposure on Na/K+ - and me+ ATPase activity in several tissues of rainbow trout (Salmo gairneri). Bull. Environ. Contam. ToxicoL 15:383 - 390.

LARSSON, A., HAUX, C. & SJOBECK, M. (1985). Fish physiology and metal pollution: Results and experiences from laboratory and field studies. Ecotoxicol. Environ. Sal 9:250 - 281.

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LEWIS, S.D. & LEWIS, W.M. (1971). The effect of zinc and copper on the osmolality of blood serum of the channel catfish, ktalurus punctatus Rafinesque, and golden shiner, Notemigonus crysoleucas Mitchell. Trans. Amer. Fish. Soc. 4:639 - 643.

LOCK, R.A.C. & VAN OVERBEEKE, A.P. (1981). Effects of mercuric chloride and methylmercuric chloride on mucus secretion in rainbow trout, Salmo gairdneri Richardson. Comp. Biochem. Physiol. 69C:67 - 73.

MARTIN, H.C. (1987). Acidic precipitation. D. Reidel Publishing Co., Dordrecht, Holland, pp. 1118.

MATHEWS, C.K. & VAN HOLDE, K.E. (1991). Biochemistry. The Benjamin/ Cummings Publishing Company, Inc. pp. 231 - 233.

MAZEAUD, M.M., MAZEAUD, F. & DONALDSON, E.M. (1981). Adrenergic responses to stress in fish. In Stress and fish (A.D. Pickering, ed.), pp. 49 - 75. Academic Press, London, New York.

McDONALD, D.G. (1983). The effects of Hi- upon the gills of freshwater fish. Can. J. Zool. 61:691 - 703.

MEYER, B.J. (1988b). Bylae. In: Die Fisiologiese Basis van Geneeskunde - Vierde uitgawe. B.J. Meyer [ed.]. JAUM Uitgewery, Pretoria. pp. i - xv.

MILLIGAN, C.L. & WOOD, C. (1982). Disturbances in haematology, fluid volume distribution and circulatory function associated with low environmental pH in the rainbow trout, Salmo gairdneri. J. Exp. Biol. 99:397 - 415.

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MOUNT, D.R., HOCKETT, J.R. & GERN, W.A. (1988). Effect of long-term exposure to acid, aluminium and low calcium on adult brook trout Salvelinus fontinalis) 2. Vitellogenesis and osmoregulation. Can. J. Fish. Aquas. Sci. 45:1633 - 1642.

MUELLER, M.E., SANCHEZ, D.A. & BERGMAN, H.L. (1991). Nature and time course of acclimation to aluminium in juvenile brook trout (Salvelinus fontinalis). II. Gill Histology. Can. J. Fish. Aquat. Sci. 48:2016 - 2027.

NARAIN, A.S. & SRIVASTAVA, P.N. (1989). Anemia in the freshwater teleost Heteropneustes fossilis under the stress of environmental pollution. Bull. Environ. Contam. Toxicol. 43:627 - 634.

NATH, K. & KUMAR, K. (1987). Toxicity of manganese and its impact on some aspects of carbohydrate metabolism of a freshwater teleost, Colisda Fasciatus. Sci. Total Environ. 67:257 - 262.

NEVILLE, C.M. (1985). Physiological response of juvenile rainbow trout, Salmo gairdneri, to acid and aluminium - prediction of field responses from laboratory data. Can. J. Fish. Aquat. Sci. 42:2009 - 2019.

NUSSEY, G., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1995). Effect of copper on the haematology and osmoregulation of the Mozambique tilapia, Oreochromis mossambicus (Cichlidae). Comp. Biochem. Physiol. 111C(3):369 - 380.

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PACKER, R.K. (1979). Acid-base balance and gas exchange in brook trout, Salvelinus fontinalis, exposed to acidic environments. J. exp. Bio. 79:127 - 134.

PICKERING, A.D. (1981). Introduction: The concept of biological stress. In: Stress and fish. A.D. Pickering [ed.] Academic Press, London. pp. 1 - 11.

POLO:), A.B.S. & MUNIZ, I.P. (1993). The effect of aluminium in soft water at low pH and different temperatures on mortality, ventilation frequency and water balance in smoltifying Atlantic salmon, Salmo salar. Environmental Biology of Fishes. 36:193 - 203.

RANDALL, R.F. & ANDERSON, P.J. (1975). Purificaion and properties of pyruvate kinase of sturgeon muscle. Biochem. J. 145:569 - 573.

RANKIN, J.C., STAGG, R.M. & BOLIS, L. (1982). Effects of pollutants on gills, pp. 207 - 219. In: D.F. Houlihan, J.C. Rankin & T. J. Shuttleworth [ed.] Gills. Cambridge University Press, Cambridge.

REID, S.D., McDONALD, D.G. & RHEM, R.R. (1991). Acclimation to sublethal aluminium: Modifications of metal-gill surface interactions of juvenile rainbow

trout (Oncorhynchus mykiss). Can. J. Fish. Aquat. Sci. 84:1996 - 2005.

SIDDENS, L.K., SEIM, W.K., CURTIS, L.R. & CHAPMAN, G.A. (1986). Comparison of continuous and episodic exposure to acidic aluminium- contaminated waters of brook trout (Salvelinus fontinalis). Can. J. Fish. Aquat. Sci. 43:2036 - 2040.

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SILBERGELD, E.K. (1972). Blood glucose: A sensitive indicator of environmental stress in fish. Dissertation.

SMITH, L.S. (1982). Introduction to fish physiology. T.F.H. Publications, Inc., Neptune, USA. pp. 346.

ULTSCH, G.R. & GROSS, G. (1979). Mucus as a diffusion barrier to oxygen: possible role in oxygen uptake at low pH in carp (Cyprinus carpio) gills. Comp. Biochem. Physiol. 62A:685 - 689.

VAN WAARDE, A., VAN DEN THILLART, G. & KESBEKE, F. (1983). Anaerobic energy metabolism of the european eel, Anguilla anguilla. L J. Comp. Physiol., 149:469.

VOET, F. & VOET, J.G. (1990). Biocl'thmisby. John Wiley & Sons, New York, Chichester, Brisbane, Toronto, Singapore. pp. 425 - 457.

WALKER, R.L., WOOD, C.M. & BERGMAN, H.L. (1991). Effects of long-term preexposure to sublethal concentrations of acid and aluminium on the ventilatory response to aluminium challenge in brook trout (Salvelinus fontinalis). Can. .I. Fish. Aquat. Sci. 48:1989 - 1995.

WEDEMEYER, G.A. & YASUTAKE, W.T. (1977). Clinical methods for the assessment of the effects of environmental stress on fish health. U.S. Tech. Pap. U.S. Fish Wildl. Serv. 89:1 - 18. Washington, D.C., USA.

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WEPENER, V., VAN VUREN, J.H.J. & DU PREEZ, H.H. (1992a). The effect of hexavalent chromium at different pH values on the haematology of Tilapia sparrmanii (Chichlidae). Comp. Biochem. Physiol. 101C(2):375 - 381.

WITTERS, H.E., VANGENECHTEN, J., VAN PUYMBROECK, S. & VANDERBORGHT, 0. (1987). Ionoregulatory and haematological responses of rainbow trout, Salmo gairdneri, to chronic acid and aluminium stress. Ann!. Soc. R. Zool. Belg. 117(1):411 - 421.

WITTERS, H.E., VAN PUYMBROECK, S. & VANDERBORGHT, J. (1991). Adrenergic response to physiological disturbances in rainbow trout, Oncorhynchus mykiss, exposed to aluminium and acid pH. Can. J. Fish. Aquat. ScL 48:414- 420.

WOOD, J.M. (1985). Effects of acidification on the mobility of metals and metalloids: an overview. Environ. Health Perspect. 63:115 - 119.

WOOD, C.M., & McDONALD, D.G. (1982). Physiological mechanisms of acid toxicity to fish, pp. 197 - 226. In: R.E. Johnson [ed.] Acid rain fisheries. American fisheries Society, Bethesda, M.D.

WOOD, C.M., PLAYLE, R.C., SIMONS, B.P., GOSS, G.G. & McDONALD, D.G. (1988). Blood gases, Acid-base status, ions and haematology in adult brook trout (Salvelinus fontinalis) under acid/aluminium exposure.

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■ : : •• ••.- CHAPTER 9

SUMMARY AND RECOMMENDATIONS

9.1 GENERAL SUMMARY

9.1.1 WATER AND SEDIMENT To ensure sustainable use of the water resources in the Upper Olifants River Catchment, information on the aquatic ecosystem has to be collected. The present study forms part of a much larger project based on the potential pollution of the Olifants River and its tributaries. Two specific localities were chosen, one situated in the Olifants River, at the Olifants River Lodge holiday resort, while the other is in the Klein Olifants River, in the nature reserve, Botshabelo. These two localities were chosen to determine the effects of the point and diffuse sources of pollution on the Olifants River and Klein Olifants River, before their confluence. At these two localities, the physical and chemical characteristics of the water were determined, as well as the manganese, copper, lead, chromium, zinc, iron, nickel and aluminium concentrations in the water and different particle sizes of the sediment. Two species of fish were also collected at these localities, namely the African sharptooth catfish, Clarias gariepinus and the moggel, Labeo umbratus. The bioaccumulation of the selected metals in the skin, liver, muscle and gill tissues was determined.

The chemical and physical characteristics of the water from these localities, showed values well within the guideline limits permitted for the protection of aquatic life, with the exception of the phosphate concentrations. The values were much higher (up to 3 mg/I) than the permitted levels for effluent water (1 mg/I). This can be attributed to effluent received from combined sewage purification

9-1 Summary and Recommendations

works located upstream from locality KOR1 and alongside the Olifants River near locality OR1, as well as effluent from informal settlements in the area and agricultural runoff. These high phosphate also caused excessive growth of algae and aquatic weeds in the river at the two localities. The 1 mg/I phosphate standard for effluent water, is therefore not acceptable and should be revised to also consider drought and low flow periods, when there are practically no dilution.

The water at the two localities were alkaline, well oxygenated and moderately hard. The hardness of the water can cause the reduction of the bioavailability of the metals in the water, as these metals form complexes with the carbonates in the water. This was evident from the bioconcentration factors calculated between the sediments and the tissues/organs of C. gariepinus and L. umbratus, which were much lower than the bioconcentration factors calculated for the water and the tissues/organs of the two species. The high metal concentrations in the water and sediment, did not necessarily indicate toxic conditions to aquatic life. The metal concentrations in the water were much lower than the concentrations in the sediments, with exceptionally high concentrations of aluminium and iron, due to the adsorption of the metals to the sediment particles. The sediment from both localities consists mainly of fine sand, followed by coarse silt and the least fine silt and fine clay particles, with the latter containing the highest concentrations of the metals. The order of metal occurrence (mean concentrations) in the water and sediment, was as follows: Fe > Cr> AI>Ni >Pb> Zn >Mn >Cu for the water and: Al >Fe > Mn > Cu > Cr >Ni > Zn > Pb for the sediment (mean concentrations).

A definite relationship was also found for especially zinc, at locality KOR1, between the water and the sediment, as zinc is a common pollutant of surface waters in many industrial areas. Considering the water quality of the Olifants

9-2 Summary and Recommendations

River and Klein Olifants River, it is important that a detailed biomonitoring programme be implimented. This should include more regular monitoring of specifically the water, for example every two weeks, as well as the sediment, once every month and this should form a part of chemical monitoring studies performed by the Department of Water Affairs and Forestry.

9.1.2 BIOACCUMULATION OF THE SELECTED METALS IN THE TISSUES AND ORGANS OF C. GARIEPINUS AND L. UMBRATUS

A large variation was found in the metal concentrations in the different tissues and organs of C. gariepinus and L. umbratus during the period February 1994 - May 1995. A summary of the statistical differences regarding the bioaccumulation of the selected metals in the different tissues and organs, as well as its dependence on the species and sex of the fish, the seasons and the localities where the fish were collected and the relationship between the metal concentrations in the water and the different tissues/organs, is given in Table 9-1. Metals were bioaccumulated mainly by the gills, but copper and iron concentrations were the highest in the liver tissue. These organs are therefore suggested to be used in a general biomonitoring programme for the assessment of the extent of bioaccumulation of metals in fish tissues/organs. The lowest concentrations of the selected metals were found in the muscle and skin tissue, which should always be included in the biomonitoring programme, as it is the edible part of the fish.

Manganese, copper and iron were accumulated in the highest concentrations by L. umbratus. Lead was also bioaccumulated in the liver and muscle tissue,

9-3 Summary and Recommendations

TABLE 9-1 SUMMARY OF STATISTICAL DIFFERENCES REGARDING THE BIOACCUMULATION IN THE SKIN, LIVER, MUSCLE AND GILL TISSUES, BETWEEN THE SPECIES, CLAMS GARIEPINUS (C.G.) AND LABE° UliEBRATUS (L.U.), THE SEASONS, LOCALITIES KOR1 AND OR1 AND THE CORRELATIONS BETWEEN THE METAL CONCENTRATIONS IN THE TISSUES/ORGANS AND THE LENGTHS OF THE TWO SPECIES

Tissu ...... ,,, Bioaccumul a a a a a O a

Sp. C.g. Sp. C.g. Sp. Lu. Sp. Lu.

Seasonal 4

Localities, Loc OR1 Loc OR1 Loc OR1

Relations corn - COTT - COTT - COTT - COTT - corr - COTT -con Bioaccumulation • • • • • • • !Ipeci Sp. Lu. Sp. Lu. Sp. Lu. Sp. Lu.

a a a

Loc OR1 Loc OR1 Loc OR1

...... „ ...... ,, ... . . - sort ations - COTT - COTT - sort - corr + con - corr - COTT a a a a a a a a

Sp. Lu. Sp. Lu. Sp. Cg. Sp. Lu. Sp. Lu.

Loc OR1 Loc OR1

Relations - corr -con -con - COTT - COTT -con - corr -corn Bioaccurnulation • • • • • • • •

Sp. Lu. Sp. Cg. Sp. Cg. Sp. Lu. Sp. Cg.

ason a a

Localities Loc OR1 Loc OR1 Loc OR1

- corr + COTT + COTT corr - COTT + COTT - COTT

Low bioaccumulation High bioaccumulation Significant seasonal difference

9-4 Summary and Recommendations

zinc in the gills and nickel in the skin, liver and muscle tissues. C. gariepinus accumulated the highest concentrations of chromium, as well as lead in the skin and gill tissues, zinc in the skin and muscle tissues and nickel in the gills. Metal concentrations in the tissues/organs of the two species can be attributed to the different feeding habits of these species, as well as the physiological role of the gills. The high metal concentrations in the muscle tissue of L. umbratus can be expected, as it feeds on the sediments of the river and high concentrations of metals occurred in the sediments at both localities. L. umbratus is also dependant on its gills for breathing and this may contribute to the high concentrations of zinc and manganese in its gills. The higher metal concentrations in the gills of C. gariepinus, are due to the metals being deposited in the gills and not regulated efficiently. Furthermore, the difference in concentrations of some metals found in the gills of the two secies may be attributed to a difference in the physiological roles of the gills in the two species. In addition to the gills, C. gariepinus has a respiratory tree enabling it to extract oxygen from the air present in the suprabranchial chamber. After gulping air, as it does periodically, the gills can obtain oxygen from the air in the branchial chamber where its concentrations in higher than the surrounding water (Prof. J.H. Swanepoel, personal communication, 1996).

No definite trend could be established regarding the differences in bioaccumulation of the metals between the different seasons. Significant differences were, however, found in the skin tissue for manganese, copper and lead, in the liver tissue for manganese and in the gill tissue for zinc. Strong significant differences were found for copper and nickel in all the tissue types between autumn 1994 and summer/autumn 1995 and between winter 1994 and autumn 1995. These were at times related to the available metal concentrations in the water. Predominant negative correlations were found between the metal

9-5 Summary and Recommendations

concentrations in the different tissues/organs of the two species and the lengths of the fish, indicating higher metal concentrations in the smaller fish. This is due to the higher metabolism and ventilation rate of smaller fish.

9.1.3 TOXICITY TESTS OF LOW PH AND ELEVATED ALUMINIUM CONCENTRATIONS ON THE

HAEMATOLOGY, OSMOREGULATION AND CARBOHYDRATE METABOLISM OF THE

MOZAMBIQUE TILAPIA, OREOCHROMIS MOSSAMBICUS

From the study, it is evident that adult 0. mossambicus exposed to pH 5.2 and combinations of pH 5.2 and 0.06 mg/1, 1 mg/1, 1.5 mg/1 and 2 mg/I aluminium, were subjected to physiological stress. In some cases, the toxicity of the acid to the fish were reduced, rather than increased by the aluminium, a finding clearly seen in the blood glucose concentrations. Concentrations of 2 mg/I at pH 5.2, had predominantly no significant effect on the haematology of the fish. The low pH and elevated aluminium concentrations were toxic to the fish at concentrations of 1 mg/I and 1.5 mg/I aluminium and pH 5.2, whereafter the fish regulated the metal. The metabolism of the fish was severely affected, as there was a definite increase in the blood glucose and lactate concentrations, indicating stress responses and anaerobic metabolism, respectively, and inhibition of certain enzyme activities.

Should the pH and hardness of the river water in the Upper Catchment of the Olifants River decrease, as were the case for other streams in the upper catchment, aluminium toxicity could have a major impact on fish populations in the area, as extremely high concentrations were also found in the sediments at localities KOR1 and OR1. Acid mine drainage, especially during periods of drought and low flow of the river or the accidental release of acid water can

9-6 Summary and Recommendations

therefore cause dramatic effects on fish populations in the area. A good example of this is the fish kills found in the area as a result of low pH and elevated aluminium concentrations.

9.2 RECOMMENDATIONS

More intensive studies into the point and especially diffuse sources of pollution in the study area are recommended, in order to determine their impact on the Upper Olifants River Catchment. Information about the distribution, form, toxicity and bioavailability of metals in the water and sediment is also needed and the development of sediment toxicity tests is also proposed, as high metal concentrations were found in the sediments. Investigations are mostly done by water analysis while little attention is paid to the analysis of fish tissue. This approach has a serious drawback, since in many cases, the causative agent may have been diluted to a level which does not allow an unambiguous interpretation at the time of sampling. It is therefore recommended that the water and sediment samples be increased in order to decrease variation and should be monitored more often. The number of fish should also be increased in order to be able to select fish of the same species and size, as the concentrations of metals vary between the different sizes of the fish. Therefore, interspecific comparisons for metals should be made only among fish of the same size range. The fish should also be large enough to obtain one gram of dried tissue, as working on a dry mass basis, decreases variation in metal concentrations. This was not possible at both localities during this study and the localities should be chosen accordingly. Biomonitoring of other aquatic organisms, such as the freshwater crabs, is also suggested, as they are known to bioaccumulate metals.

9-7 Summary and Recommendations

By monitoring biological indicators such as fish and crabs, sublethal changes in the water quality can be detected, which may go undetected with the chemical monitoring programmes. These sublethal changes in water quality can, however, severely affect aquatic populations over a period of time. Metals were detected in the fish and therefore water quality as well as aquatic organisms should be used in general biomonitoring programmes. Other tests, such as the breeding of, for example C. gariepinus under controlled conditions, after which it is exposed to different pollutants in order to obtain information regarding the bioaccumulation of these substances, are also proposed. These can meet the information requirements of water resource managers, as well as effluent dischargers, required to comply with limits set for protecting the aquatic environment.

Further degradation of the water quality of the Olifants River cannot be afforded and therefore field studies, combined to such experimental work are suggested. Measures should be taken to reduce the impact of mines, industries and especially sewage purification works on the water quality of the Upper Olifants River Catchment, especially during the drier and low flow periods. Ecologically sound management objectives related to the health of the system should therefore be set and monitoring of wether those objectives are being achieved, should take place. If not, the appropriate actions should be taken.

- o o o 0 oo o -

9-8