Page 1
Chapter 5 – Field Studies
74
CHAPTER 5
FIELD STUDIES
5.1 Introduction
To protect biological recourses depends on the ability to identify and predict the
effects of human actions on these biological systems and to distinguish between
natural and human-induced variation in biological conditions (Karr and Chu,
1999a). Substances that can be potentially toxic, enter the aquatic environment
from a number of different sources. At direct point sources, discharges enter a
water source at a single point, for example discharges of domestic sewages and
industrial effluents. Non-point or diffuse sources where toxic substances enter
surface and underground water through runoff from urban and industrial areas,
leachates from domestic and solid waste disposal sites and mining operations.
Spillage and the release of agricultural chemicals are classified as unquantified
point source discharges; there is little or no data available due to their irregular
discharges (Roux, 1994; Sutton and Oliveira, 1987; Heath and Claassen, 1999).
For the evaluation of the biomarkers under field conditions, a reference site (Rust
de Winter Dam) and two polluted sites, Loskop Dam and Hartebeespoort Dam,
were chosen. During biological assessment, a standard is needed to evaluate
the conditions at one or more sites of interest, against. This standard/reference
condition provides the baseline for site evaluation (Karr and Chu, 1999b).
Reference conditions describe the characteristics of water resources that are
least impaired by human activities. When there are no undisturbed sites, “least
impacted sites” or “best attainable conditions” may be used as reference (Roux,
1994).
Page 2
Chapter 5 – Field Studies
75
5.2 Reference site
5.2.1 Rust de Winter Dam
The Elands River is the most westerly tributary of the Olifants River and rises a
few kilometers south of Rayton neat Kaztan, north of the N4-highway. It drains
northwards through hilly country (± 60 km) to the Rust de Winter Dam. From this
point it flows in a westerly direction to the south of the Springbok Flats to its
confluence with the Olifants River, downstream of Marble Hall (Figure 5.3)
(Theron, Prinsloo, Grimsehl & Pullen Inc., 1991a).
The Rust de Winter Dam is the uppermost major dam in the Elands River and is
surrounded by a nature reserve (Figure 5.2). The dam has a catchment area of 1
147 km2 and a total storage capacity of 28.1 x 106 m3. The dam was completed
in 1933 to provide irrigation water to vegetable farms below the dam, but has in
recent years become a popular recreational area for fishing and boating (Figure
5.1) (Butty et al., 1980; Theron, Prinsloo, Grimsehl & Pullen Inc., 1991a).
Rainfall occurs predominantly in the summer months usually between October
and March, with January usually experiencing the heaviest rainfall. There is no
development that can have a major impact on the dam. Irrigation is the dominant
water user from the dam and agriculture is the major land use in the area.
Agricultural activity is restricted to an area close to Zonderwater and to farms
near the impoundment itself. The greatest portion of the catchment is
undeveloped bushveld, which is utilised for cattle ranching (Butty et al., 1980;
Theron, Prinsloo, Grimsehl & Pullen Inc., 1991a). Taking the last statement into
consideration, the RDW dam was thus the best choice as a reference site since it
contains relative little pollution.
Page 3
Chapter 5 – Field Studies
76
Figure 5.1 : The Rust de Winter Dam.
Page 4
Chapter 5 – Field Studies
77
Figure 5.2 : A map showing the catchment areas, rivers and
urban/industrial developments around the Rust de Winter Dam (A) and
Loskop Dam (B).
(A)
(B)
Page 5
Chapter 5 – Field Studies
78
Figure 5.3 : The Rust de Winter Catchment (Buttey et al., 1980).
Page 6
Chapter 5 – Field Studies
79
5.3 Polluted sites
5.3.1 Loskop Dam
The Upper Olifants River catchment comprises the drainage areas of the Olifants
River, Klein Olifants River and Wilge River, with tributaries down to the Loskop
Dam (Figure 5.2). The headwaters of these rivers are located along the Highveld
Ridge in the Secunda-Bethal areas and the rivers then flow in a northerly
direction towards the Loskop Dam. The natural rivers and streams have been
extensively dammed with the result that the stream flow is now highly regulated.
The major impoundments upstream of the Loskop Dam include the Witbank
Dam, Middleburg Dam, Bronkhorstspruit Dam and Premiere Mine Dam (BKS
(Pty) Ltd., 1998). In 1990, the population of the Olifants River catchment was
±2.5 x 106. Two thirds of this population lives in rural or semi-urban (settlements)
conditions. Middelburg and Witbank are the largest urban concentrations (Heath
and Claassen, 1999).
Over the past few years, the Olifants River has been systematically impaired
because of an increase in agricultural and mining activities, industrial
development and urbanisation. This river system is often described as one of the
most polluted systems in South Africa and is known as “The Battered River” (Van
Vuren et al., 1999). Along the Olifants River there are intensive and subsistence
agriculture as well as numerous point and diffuse sources of industrial pollution
(Heath and Claassen, 1999).
Loskop Dam was built in 1939, 48 km north of Middelburg and raised in 1977 by
9.1 m. The storage capacity rose from 180 x 106 m3 to 348.1 x 106 m3. The total
catchment area for the dam is 12 261 km2 (SANCOLD, 1978). The total
catchment incorporates the most industrialised region of the Olifants River basin
and the Loskop Dam is the biggest storage unit in the Olifants River catchment
(Figure 5.4) (Theron, Prinsloo, Grimsehl & Pullen Inc., 1991b). Rainfall occurs
mainly in the summer months; with January experiencing the heaviest rain and
Page 7
Chapter 5 – Field Studies
80
90% of the water available from the Loskop Dam is used for irrigation (James
and Van Wyk, 1993).
The Loskop Dam has been described as a sink for heavy metals deriving from
the upper catchment and the whole of the Olifants River has been described as
degraded and contaminated with metals and other chemicals. These concerns
have been expressed as a consequence of the large number of agricultural,
industrial and mining activities in the catchment (Grobler et al., 1994). A large
number of mines, predominately coal mines, are located in the Loskop Dam
catchment and are concentrated mainly in the Olifants and Klein Olifants River
catchments upstream of the Witbank and Middelburg Dams respectively (Du
Plessis and Maré, 1999). The most extensive coal mining takes place at the
Witbank Coalfields and Highveld Coalfields. Coalmines provide essential fuel to
local power stations e.g. Arnot, Hendrina, Kriel, Komati, Duhva, Matla and
Kendal, as well as to the domestic and international markets. Coal mining and
industries in the Witbank-Middelburg and Phalaborwa areas also impact the
Olifants River. These mine effluents contain a complex of chemicals, many of
which may have deleterious effects for aquatic systems (Van Vuren et al., 1999).
Water discharges from the mines can originate from various sources, including
sewage treatment plants, seepage from opencast and underground mining
operations. The return flows from sewage treatment plants are released into
natural streams or re-used in mining operations. Return flows are also used for
irrigation purposes. Seepage and decanting from mines can result in serious
water quality related problems (BKS (Pty) Ltd., 1998; Du Plessis and Maré,
1998).
Presently, Loskop Dam supplies domestic, industrial and irrigation water users.
The impoundment mainly supplies the large irrigation schemes downstream of
the dam (BKS (Pty) Ltd., 1998).
Page 8
Chapter 5 – Field Studies
81
Figure 5.4 : The Loskop Dam
Page 9
Chapter 5 – Field Studies
82
5.3.2 Hartebeespoort Dam
The Hartebeespoort Dam is situated on the Crocodile River, about 16 km
southwest of the town of Brits and 37 km due east of Pretoria (SANCOLD, 1978)
and in the Highveld region of northern South Africa, 250 km south of the tropic of
Capricorn (Figure 5.6) (Hely-Hutchinson and Schumann, 1997). The 5
catchment basins of the dam are, from west to east, the Magalies/Skeerpoort, the
Crocodile, the Jukskei, the Hennops and the Swartspruit basin (Van Riet, 1987).
The Crocodile River is the most intensive irrigation system in South Africa with
numerous point and diffuse sources of domestic and industrial pollution (Figure
5.7) (Heath and Claassen, 1999).
The Hartebeespoort Dam was built in 1923 downstream of the confluence of the
Crocodile River and the Magalies River, and was raised in 1971 with 2.12 m.
The dam has a total storage capacity of 185.49 x 106 m3 and a catchment area of
4 112 km2 (Rossouw, 1992). Rainfall is highly seasonal and occurs mainly
between October and March. Land usage in the Hartebeespoort Dam catchment
can be devided into two categories, namely rural and urban. The commercial,
residential and industrial areas that are associated with the northern suburbs of
Johannesburg and other smaller towns on the Witwatersrand make up the urban
land use, while the rest of the area is used for natural reserves and agriculture
(National Institute for Water Research, 1985) (Figure 5.5).
The rivers that flow into the Hartebeespoort Dam are carrying an ever-increasing
volume of wastewater form a rapidly growing industrial and urban complex
(Aucamp et al., 1987) and Van Riet (1987) stated that the water of the
Hartebeespoort Dam is becoming unsuitable for agriculture, development and
recreation. The upper reaches of the Crocodile River drains the Johannesburg
Northern suburbs and its Hennops tributary drains Kempton Park, Tembisa,
Midrand and Centurion. The Magalies River drains the town of Magaliesburg and
Swartspruit drains the town of Hartebeespoort (Sutton and Oliveira, 1987). Other
Page 10
Chapter 5 – Field Studies
83
catchment areas include towns like Clayville, Olifantsfontein, Alexandra and a
part of Atteridgeville and Saulsville (Rossouw, 1992).
Due to the intense urbanisation of this catchment it has the potential to decrease
the water quality of the natural resources due to the dumping of effluents and
solid-waste, mines, industrial activities, etc. There are also the sewerage
treatment plants of Johannesburg, Midrand, Kempton Park, Centurion,
Olifantsfontein, Randfontein, Kurgerdorp and Roordepoort in the catchment area.
Industrial dumping sites include AEK (Pelindaba and Valindaba), AECI-
Modderfontein and the Kelvin power station. There is also the potential
contamination of storm water runoff from industrial areas like Clayville, Isando
and Eastleigh as well as residential areas like Tembisa, Alexandra, Atteridgeville,
etc. The biggest influence on the water quality of the Hartebeespoort Dam is
form the Modderfontein stream that is upstream form the confluence with the
Jukskei River and thus the Crocodile River (Rossouw, 1992).
Figure 5.5 : The Hartebeespoort Dam
Page 11
Chapter 5 – Field Studies
84
Figure 5.6 : A map showing the catchment areas, rivers and
urban/industrial areas of the Hartebeespoort Dam.
Page 12
Chapter 5 – Field Studies
85
Figu
re 5
.7 :
The
catc
hmen
t are
as o
f the
Har
tebe
espo
ort D
am (B
KS
(Pty
) Ltd
., 19
92).
Page 13
Chapter 5 – Field Studies
86
5.4 Materials and Methods
5.4.1 Water Quality
Physical water quality variables such as temperature, pH, conductivity, total
dissolved salts (TDS) and dissolved oxygen were determined at each of the sites
for the summer and winter surveys, using field instruments (Cyberscan DO100
Handheld Dissolved Oxygen Meter; Waterproof pHScan WP2 Tester; Cyberscan
EC-con-300 TDS/Conductivity Meter) (Table 5.1).
5.4.2 Fish sample collection and preservation
Twenty (20) fish were collected in the Rust de Winter Dam (S 24º49.800’; E 027º
29.102’), the Loskop Dam (S 25º14.220’; E 028º30.402’) and the Hartebeespoort
Dam (S 25±º45.678’; E 027º52.656’) during the summer and winter of the year,
2000. Samples were taken seasonally to establish increases and decreases in
biomarker activity. Twenty fish were collected to ensure more reliable results by
reducing variation in the data. The fish were collected using gill nets (70 – 100
mm stretched mesh sizes). The fish were removed from the nets and placed into
a portable holding tank with constant water circulation, till dissection to reduce
handling stress. The standard length (cm), mass (kg) and gender (male/female)
of each fish were recorded as well as other comments such a lesions or cysts on
the skin or liver of the fish, etc. (See Table 5.2 and 5.3)
After the above biological parameters were recorded, blood was drawn from the
caudial vein of the fish with a 2,5 ml pre-heparinised syringe (0,1 ml of 5 000
iu/mΡ sodium heparin) and a 0,6 x 30 mm needle. The blood was transferred to
a 5 ml vacutainer and kept on ice. The fish were then decapitated on a
polythene dissection board using clean, stainless steel tools. The liver was
removed and placed in a cryotube and frozen in liquid nitrogen at -196ºC. After
all 20 fish were dissected, 210 µl of the blood was removed for ALAD analysis
and placed in a clean vacutainer and frozen at -20ºC. The remaining blood was
centrifuged at 3 000 r.p.m. (1 000 g) for 10 min in an automatic refrigerated
centrifuge (Sorvall Superspeed RC2-B) and prepared according to the method of
Page 14
Chapter 5 – Field Studies
87
Wittaker (1984) for further analysis. The plasma was placed in a cryotube and
frozen in liquid nitrogen. The red blood cells were kept in the vacutainers and
frozen at -20ºC. At the laboratory the samples were removed from the liquid
nitrogen and kept in a -70ºC freezer.
5.4.3 Method of biomarker analysis
All the protein analyses were done according to the colourimetric method of
Bradford (1976). Table 5.4 shows the different sample preparation and
biomarker analysis methods used during this study.
Table 5.4 : Methods and apparatus used in biomarker analysis of field
samples.
Biomarker Fish tissue
Preparation Method
Preparation Apparatus
Analysis Apparatus
Absorbancy Wavelength
(nm)
Analysis Method
AChE Red blood cells
Wittaker (1984)
Automated Refrigerated Centrifuge (Sorvall Superspeed RC2-B)
See Table 4.8 405 Ellman et al. (1961)
ALAD Whole blood
- Automated Refrigerated Centrifuge (Sorvall Superspeed RC2-B)
SP-8-100 Ultraviolet Spectro- Photometer (PYE Unicam)
555 Schaller and Berlin (1984)
EROD Liver Besselink et al. (1997)
Automated Refrigerated Centrifuge (Sorvall Superspeed RC2-B) Beckman L8-70M Ultracentrifuge
See Table 4.8 Exitation: 510mm Emmision: 586mm
Burke and Mayer (1974)
Glucose Plasma - See Table 4.8 See Table 4.8 546 See Table 4.8
Glycogen Liver - - SP-8-100 Ultraviolet Spectro- Photometer
620 Seifter et al. (1950)
Page 15
88
Table 5.1 : Physical water qualities of the reference and polluted sites during the summer and winter surveys.
Reference site Polluted sites Variables
Rust de Winter Dam Loskop Dam Hartebeespoort Dam
Date 14/01/00 05/06/00 15/01/00 06/06/00 08/02/00 08/06/00
Time 14:45 15:20 12:10 14:10 16:00 16:25
Temperature (ºC) 25.7 15.5 26.3 18.1 23.1 16.8
pH 8.3 6.3 9.2 8.6 9.3 8.5
Dissolved Oxygen – mg/l 11.1 3.02 12.3 5.2 7.6 3.23
– % 163 30.6 112 56.5 117 34.4
Conductivity(µS/m) 178 136 409 309 511 477
Total dissolved salts (TDS) 88.9 - 203 308 233 -
- = Variables not measured
Page 16
89
Table 5.2 : Biological variables (standard length, mass and gender) measured of O. mossambicus collected at the
reference and polluted sites during the summer survey.
Reference site Polluted sites Rust de Winter Dam Loskop Dam Hartebeespoort Dam
Fish no.
Standard Length(cm)
Mass (kg)
Gender Comments Standard Length(cm)
Mass (kg)
Gender Comments Standard Length(cm)
Mass (kg)
Gender Comments
1 38 1.15 ? - 39 1.25 ? - 23.5 0.30 ? - 2 34 0.65 ? - 38.5 1.20 ? - 24.5 0.35 ? - 3 35 0.75 ? - 38 1.10 ? - 24 0.35 ? - 4 33 0.55 ? One eye
blind 36 0.85 ? - 24.5 0.35 ? -
5 36 0.90 ? - 42 1.55 ? - 24 0.35 ? - 6 41 1.40 ? - 41 1.45 ? - 25.5 0.40 ? - 7 42 1.50 ? - 36 0.90 ? - 25 0.40 ? - 8 35 0.75 ? - 35.5 0.80 ? - 34 0.55 ? - 9 35.5 0.80 ? - 37 1.00 ? - 24 0.30 ? -
10 36 0.90 ? - 37.5 1.00 ? - 26 0.40 ? - 11 42 1.50 ? - 37 1.00 ? - 23 0.30 ? - 12 36 0.90 ? Many
Argulus parasites
35 0.75 ? - 23 0.30 ? -
13 36 0.90 ? - 44 1.75 ? - 24 0.35 ? - 14 39 1.15 ? - 37.5 1.00 ? - 24 0.35 ? - 15 29 0.45 ? - 39 1.30 ? - 24 0.35 ? - 16 28.5 0.50 ? - 35 0.75 ? - - - - - 17 33 0.55 ? - 47 2.10 ? - - - - - 18 36 0.90 ? - 40 1.35 ? - - - - - 19 - - - - 37 1.00 ? - - - - - 20 - - - - 47 2.10 ? - - - - -
Page 17
90
Table 5.3 : Biological variables (standard length, mass and gender) measured of O. mossambicus collected at the
reference and polluted sites during the winter survey.
Reference site Polluted sites Rust de Winter Dam Loskop Dam Hartebeespoort Dam
Fish no.
Standard Length(cm)
Mass (kg)
Gender Comments Standard Length(cm)
Mass (kg)
Gender Comments Standard Length(cm)
Mass (kg)
Gender Comments
1 42.5 1.26 ? - 41 1.75 ? - 28.5 0.45 ? Cysts on operculum
2 36 0.77 ? - 36 1.15 ? - 40 1.25 ? One eye blind 3 25 0.23 ? - 41 1.40 ? - 47 1.90 ? - 4 39.5 1.25 ? - 38.5 1.50 ? - 41.5 1.50 ? - 5 38 1.25 ? - 42 1.80 ? Liver dark,
hard with cysts
36.5 1.00 ? -
6 42 1.25 ? - 45.5 2.20 ? - 42 1.50 ? Liver contains cysts
7 36 0.78 ? - 39.5 1.80 ? - 38 1.20 ? - 8 39 1.27 ? - 39.5 1.20 ? - 45.5 1.70 ? - 9 37 0.90 ? - 49 2.40 ? - 42 1.30 ? Bottom lip
deformed 10 35 0.90 ? - 45 2.20 ? - 37 1.10 ? One eye
cataract 11 36.5 1.10 ? - 36.5 1.20 ? - 38 1.25 ? One eye
cataract 12 40 1.25 ? - 43 1.75 ? Liver hard,
with cysts 44.5 1.80 ? -
13 34 0.85 ? One eye blind 45 2.25 ? - 44 1.75 ? - 14 39 1.20 ? - 38.5 1.25 ? - 40 1.40 ?
Liver hard with cysts
15 40 1.05 ? - 39.5 1.50 ? - - - - - 16 42 1.25 ? - 40 1.50 ? - - - - - 17 43 1.60 ? Nematode in
brain cavity 41.5 1.70 ? - - - - -
18 38 1.10 ? - 43 1.90 ? Liver contains cysts
- - - -
19 40.5 1.30 ? - 42 2.00 ? - - - - - 20 35 0.79 ? - 42 1.75 ? - - - - -
Page 18
Chapter 5 – Field Studies
91
5.4.4 Statistical Analysis
Statistical analysis was done on the values for the different biomarker
variables measured during the field studies. The statistical analysis was done
by a consultant from STATKON at the Rand Afrikaans University (RAU) by
using the SPSS Computer Systems. The variables were evaluated
statistically with ANOVA. When the ANOVA indicated statistical differences,
Dunnett’s test was employed to test for significance differences between the
reference site (Rust de Winter Dam) and the polluted sites, Loskop and
Hartebeespoort Dam. Significant differences for a dam between the two
seasonal surveys were also determined.
5.5 Results
5.5.1 Physical water quality
The selected physical water quality results for the dams can be seen in Table
5.1.
The variation in temperature for the different dams sampled, is a function of
seasonality. The water temperatures are higher during the summer, and a
drop in water temperature during the winter. All three dams showed a slight
decrease in pH. The Rust de Winter Dam showed the highest dissolved
oxygen content (163% saturation) during the summer, with Loskop and
Hartebeespoort Dam with lower levels of 112 and 117% saturation,
respectively. There was a drop in dissolved oxygen levels in all three dams
during the winter. Conductivity is the ability to conduct an electrical current
due to the presence of ions in the water. Loskop Dam (summer = 409 µS/m;
winter = 309 µS/m) and Hartebeespoort Dam (summer = 511 µS/m; winter =
477 µS/m) showed the highest values for both seasons. Total dissolved salts
(TDS) are a measure of all the salts dissolved in the water and Loskop Dam
and Hartebeespoort Dam showed the highest values during the summer, with
203 and 233 respectively (Table 5.1).
Page 19
Chapter 5 – Field Studies
92
5.5.2 Differences between genders
There were no significant differences for the values measured for the different
variables, between the male and female fish collected at the different sites for
both surveys.
5.5.3 Different protein values
Table 5.5 shows the protein values for the different tissues used for the
biomarker analysis. All the biomarker results were expressed in terms of
protein.
Table 5.5 : Protein content of different tissues used for biomarker
analysis.
Tissue protein content (mg/ml )
Dam Season WB RBC Plasma Liver
Summer n
Mean±Sd
Min/max
P
18
63.60±22.81
27.83-108.79
*
16
278.06±200.73
29.90-706.00
-
18
53.39±14.10
27.08-74.71
-
18
29.09±11.24
12.97-61.03
*
RDW
Winter n
Mean±Sd
Min/max
P
19
130.27±35.52
57.93-207.31
*
20
225.12∀27.19
189.60-276.91
-
17
47.15±9.28
30.26-66.96
-
20
43.07±11.85
11.46-66.55
*
Summer n
Mean±Sd
Min/max
P
19
58.36±17.56
20.58-90.90
*
18
474.03±240.45
182.70-897.65
* / <
19
57.10±15.26
21.65-77.00
*
20
34.57±13.01
23.24-84.38
-
LD
Winter n
Mean±Sd
Min/max
P
18
112.03±38.63
25.47-210.69
*
20
259.37±47.97
218.09-412.85
* / <
20
46.27±6.73
37.06-64.47
*
20
34.83±8.87
12.97-52.72
<
Summer n
Mean±Sd
Min/max
P
14
123.35±22.49
84.42-165.60
<
15
184.11±9.27
169.45-204.9
*
14
53.27±16.11
32.02-85.87
*
15
18.78±9.69
10.07-42.79
* / <
HBP
Winter n
Mean±Sd
Min/max
P
13
125.16±45.18
62.81-203.23
-
14
302.12±155.76
195.95-728.69
*
14
42.88±8.05
33.11-64.01
*
14
39.90±9.77
22.30-54.11
*
RDW=Rust de Winter Dam* = Significant difference (p<0.05) between the same dam LD=Loskop Dam different seasonal surveys. HBP=Hartebeespoort Dam < = Significant difference (p<0.05) between the polluted sites WB=Whole Blood (LD/HBP) and the reference site (RDW) during the same RBC=Red Blood Cells survey. - = No significant difference
Page 20
Chapter 5 – Field Studies
93
There is a significant difference (p<0.05) between the whole blood protein
levels in the Rust de Winter Dam between the surveys, 63.60 and 130.27
mg/ml for summer and winter respectively. Loskop Dam also showed a
significant difference with 58.36±17.56 and 112.03±38.63 mg/ml for the two
surveys. The Hartebeespoort Dam showed a significant difference with
123.35±22.49 mg/ml and Rust de Winter Dam (63.06±22.81 mg/ml) during
the summer survey.
Rust de Winter Dam showed no significant difference in the red blood cell
protein levels for the two surveys but Loskop Dam did with values of
474.03±240.45 mg/ml for the summer and 259.37±47.49 mg/ml for the winter
survey. Hartebeespoort Dam also showed significant differences (p<0.05) for
the two surveys, with 184.11±9.27 and 302.12±155.76 mg/ml for both the
summer and the winter surveys. The red blood cell protein levels for Loskop
Dam for both summer and winter showed significant differences form the
reference site, Rust de Winter Dam.
The plasma protein content of the Rust de winter Dam showed no significant
differences during the two surveys. Both the Loskop Dam
(summer=57.10±15.26 mg/ml; winter=46.27±6.73 mg/ml) and the
Hartebeespoort Dam (summer=53.27±16.11 mg/ml; winter=42.88±8.05
mg/ml) showed significant differences (p<0.05) for the two surveys. There
were no significant differences between the po lluted sites and the reference
site for the summer and winter surveys.
Both Rust de Winter Dam (summer=29.09±11.24 mg/ml; winter=43.07±11.85
mg/ml) and Hartebeespoort Dam (summer=18.78±9.69 mg/ml;
winter=39.90±9.77 mg/ml) showed significant differences (p<0.05) for the liver
protein content between the two seasonal surveys. Hartebeespoort Dam also
showed a significant difference from the reference site for the summer survey
and Loskop Dam for the winter survey.
Page 21
Chapter 5 – Field Studies
94
5.5.4 Different biomarker analysis
Figures 5.8 to 5.12 show the results obtained for the different biomarkers
during the two seasonal surveys.
The Acetylcholinesterase (AChE) enzyme activity (Figure 5.8) for the summer
survey showed inhibition for Loskop Dam (1.59 x 10-4±1.81 x 10-4 OD/min.mg
protein) and induction for Hartebeespoort Dam (5.48 x 10-4±2.57 x 10-4
OD/min.mg protein), but there was no significant difference (p<0.05) between
the reference and polluted sites. There were also no significant differences
between the dams for the winter survey, although there was some inhibition at
the polluted sites. A significant increase was seen in AChE activity in the
Loskop Dam between the two surveys with 1.59 x 10-4±1.81 x 10-4 OD/min.mg
protein for the summer and 4.17 x 10-4±2.66 x 10-4 OD/min.mg protein for the
winter survey.
0.00E+00
2.00E-04
4.00E-04
6.00E-04
8.00E-04
1.00E-03
1
AC
hE
(OD
/min
.mg
pro
tein
)
RDW LD HBP
0.00E+002.00E-044.00E-046.00E-048.00E-041.00E-03
1
AC
hE (O
D/m
in.m
g pr
otei
n)
RDW LD HBP
Figure 5.8 : Acetylcholinesterase (AChE) enzyme activity during the summer
(A) and winter (B) surveys (RDW=Rust de Winter Dam, LD=Loskop Dam, HBP=Hartebeespoort Dam).
A
B
Page 22
Chapter 5 – Field Studies
95
05
10152025
1
AL
AD
(U/h
.mg
pro
tein
)
RDW LD HBP
00.5
11.5
22.5
1ALA
D (U
/h.m
g pr
otei
n)
RDW LD HBP
Figure 5.9 : ∗-Aminolevulinic acid dehydratase (ALAD) enzyme activity during
the summer (A) and winter (B) surveys (RDW=Rust de Winter Dam, LD=Loskop Dam, HBP=Hartebeespoort Dam; * = p<0.05).
0
0.5
1
1.5
1
ER
OD
(nM
/min
.mg
prot
ein)
RDW LD HBP
0
0.5
1
1.5
2
1
ER
OD
(nM
/min
.mg
prot
ein)
RDW LD HBP
Figure 5.10 : Ethoxyresorufin-O-deethylase (EROD) enzyme activity during the
summer (A) and winter (B) surveys (RDW=Rust de Winter Dam, LD=Loskop Dam, HBP=Hartebeespoort Dam; * = p<0.05).
A
B
* *
A
B
*
*
Page 23
Chapter 5 – Field Studies
96
0.00E+00
1.00E-02
2.00E-02
3.00E-02
4.00E-02
1
Pla
sma
Glu
cose
(mg
g
luco
se/m
g p
rote
in)
RDW LD HBP
0.00E+00
1.00E-02
2.00E-02
3.00E-02
4.00E-02
1
Pla
sma
Glu
cose
(mg
gluc
ose/
mg
prot
ein)
RDW LD HBP
Figure 5.11 : Plasma Glucose content during the summer (A) and winter (B)
surveys (RDW=Rust de Winter Dam, LD=Loskop Dam, HBP=Hartebeespoort Dam; * = p<0.05; **= p<0.1).
0200400600800
100012001400
1
Liv
er G
lyco
gen
(m
g
gly
cog
en/1
00 g
live
r)
RDW LD HBP
0500
1000
15002000
2500
1
Live
r G
lyco
gen
(mg
glyc
ogen
/100
g li
ver)
RDW LD HBP
Figure 5.12 : Liver Glycogen content during the summer (A) and winter (B)
surveys (RDW=Rust de Winter Dam, LD=Loskop Dam, HBP=Hartebeespoort Dam; * = p<0.05).
*
A
B
A
B **
Page 24
Chapter 5 – Field Studies
97
Figure 5.9 shows the ALAD enzyme activity for both surveys. Loskop Dam
(5.602 U/h.mg protein) and Hartebeespoort Dam (7.725±3.40 U/h.mg protein)
showed significant differences (p<0.05) in the inhibition of ALAD activity form
Rust de Winter Dam (16.093±7.48 U/h.mg protein) for the summer survey.
There was no significant differences between the dams for the winter survey,
although it looks like inhibition did occur, and all three sites showed significant
decreases in enzyme activity between the two surveys.
EROD synthesis for the three sites is shown in Figure 5.10. Induction did
occur in Loskop Dam (1.011±0.33 nM/min.mg protein) and Hartebeespoort
Dam (1.059±0.28 nM/min.mg protein) during the summer, but not statistical
significant. A significant difference did occur between Loskop Dam
(0.493±0.38 nM/min.mg protein), Hartebeespoort Dam (1.295±0.38
nM/min.mg protein) and Rust de Winter Dam (0.874±0.47 nM/min.mg protein)
for the winter survey with Loskop Dam showing inhibition and Hartebeespoort
Dam, induction. Loskop Dam also showed a significant difference between
the two surveys.
Plasma Glucose levels (Figure 5.11) showed no significant differences for the
summer survey even though Loskop Dam showed a slight increase (2.241 x
10-2±8.00 x 10-3 mg glucose/mg protein) and Hartebeespoort Dam (1.454 x
10-2±9.84 x 10-3 mg glucose/mg protein) a decrease against the reference
site, Rust de Winter Dam. The fish caught at Loskop Dam did show a
significant increase in plasma glucose levels (2.14 x 10-2±7.47 x 10-3 mg
glucose/mg protein) during the winter survey, but there was no difference in
plasma glucose levels between the two surveys. Hartebeespoort Dam (p<0.1;
2.162 x 10-2±6.56 x 10-3 mg glucose/mg protein) showed a significant increase
from Rust de Winter Dam during the winter survey. Rust de Winter Dam did
show a significant decrease in glucose levels for the summer and winter
surveys with 2.147 x 10-2±8.19 x 10-3 mg glucose/mg protein and 1.577 x 10-2
±5.40 x 10-3 mg glucose/mg protein, respectively. Hartebeespoort Dam also
showed a significant increase with glucose levels of 1.454 x 10-2±9.84 x 10-3
mg glucose/mg protein for the summer and 2.162 x 10-2±6.56 x 10-3 mg
glucose/mg protein for the winter surveys.
Page 25
Chapter 5 – Field Studies
98
Figure 5.12 shows the liver glycogen content determined during the summer
and winter surveys. Hartebeespoort Dam showed a significant decrease in
values form the reference site. The standard deviation for this survey was
extremely high. There was a significant increase in liver glycogen content for
all three sites form the summer to the winter surveys, with Rust de Winter
Dam increasing from 567.0±634.51 to 1349.864±432.26 mg glycogen/100 g
liver, and Hartebeespoort Dam with 86.0±147.59 to 1569.391±616.88 mg
glycogen/100 g liver. There was no significant difference (p<0.05) between
the sites during the winter survey.
Plasma glucose levels and EROD enzyme activity both showed the smallest
standard deviation and ALAD and AChE enzyme activity and liver glycogen
content the highest. This should be taken into account when choosing a
suitable battery of biomarkers.
5.6 Discussion
Factors such as overpopulation, industrialisation and the improvement in
agricultural practices for better crops production, have all contributed to the
general deterioration of the environment, the same environment that humanity
is completely dependent on for life. Industrial effluents lead to heavy metal
enrichment of the aquatic environment during melting and smelting operations
and these metals can produce various physiological changes in the aquatic
life (Sastry and Shukla, 1994).
The variability of the physical environmental conditions directly affects the
biotic patterns such as abundance of species and micro- and macro-
geographic distribution of all organisms (Roux, 1994). Depending on the
variable, the ambient water quality as well as the organism involved can be
affected. The thermal characteristics of an aquatic ecosystem are reliant on
hydrological, climatical and structural features of its catchment under normal
conditions. Anthropogenic activities can decrease or increase the natural
variation in water temperature negatively. Water temperature as a variable
should be seen as a factor influencing the toxicity of pollutants (Van Vuren et
Page 26
Chapter 5 – Field Studies
99
al., 1999) for example, metals like cadmium and zinc increase in toxicity with
an increase in water temperature (DWAF, 1996). The seasonal water
temperature changes, seen in Table 5.1, is still within the tolerance ranges of
Oreochromis mossambicus (Chapter 3).
Geological and atmospheric influences determine the natural pH of a
waterbody (Van Vuren et al., 1999) and the pH range that is not directly lethal
to fish is between 5 and 9. It is however important to remember that the
toxicity of several common toxicants like metals, is affected by pH changes
within this range (Alabaster and Lloyd, 1980).
Dissolved oxygen (DO) is essential for maintaining aquatic life and low oxygen
levels create an increase in the metabolic rate of fish, thereby causing an
increased rate of water pumping over the gills. Thus increasing the amount of
toxin in contact with the gill surface, where it is absorbed (Alabaster and
Lloyd, 1980; Van Vuren et al., 1999). The target range of DO levels are
between 80-120 % saturation (Van Vuren et al., 1999). The drop in oxygen
levels during the winter could be a function of seasonality. Higher inflow of
the rivers around the impoundments takes place because most of the rainfall
occurs during October to March, with January experiencing the most rainfall
for all the sites. Thus during the summer, large amounts of organic matter
enter the water from industrial/domestic wastes and could utilise a large
amount of DO due to microbial respiration (Van Vuren et al., 1999). During
the winter survey the DO levels were extremely low. The high rainfall
experienced during the summer could have caused a large influx of mining
runoff and industrial/domestic waste into the impoundments studied, where
microbial respiration utilised the a large amount of DO. The above-mentioned
as well as a reduction of inflow of the rivers surrounding the impoundments
during the winter could attribute to a decreased DO level.
Total dissolved salts (TDS) concentrations are a measure of the salts
dissolved in the water while the conductivity refers to the water’s ability to
conduct an electrical current due to the presence of ions in the water. The
ions have the ability to carry an electrical charge (CO32-, HCO-, Cl-, SO4
2-,
Page 27
Chapter 5 – Field Studies
100
NO3-, Na+, Ca2+, Mg2+). Geological weathering and atmospheric conditions
contribute to the TDS of natural aquatic systems, however domestic and
industrial discharges and surface runoff from urban, industrial and agricultural
areas, together with evaporation can also increase the TDS levels. Natural
fluctuations in TDS could be the dissolution of rocks, soils and decomposing
plant material (Van Vuren et al., 1999). A heavy summer rain season could
enhance the above-mentioned fluctuations. The high conductivity and TDS
values for Loskop Dam could be related to high land usage disturbances such
as mining, and in the case of Hartebeespoort Dam, agricultural runoff and
urban and rural settlement activities, as well as water works (water care
facilities) in the catchment could cause these high levels.
The inhibition of AChE enzyme activity is specific to organophosphorus and
carbamate compounds. These compounds are widely used in pesticides and
the inhibition of AChE has been used for the assessment of
organophosphorus pesticide pollution. Fish can detoxify these compounds
more easily than invertebrates and significant AChE inhibition in fish can only
be detected at relatively high concentrations. Organisms that survived acute
effects of pesticide pollution show a recovery of AChE activity that is slow but
dependant on the spontaneous dephosphorylation of the inhibiton site and on
the synthesis of new AChE (Peakall, 1992; Ibrahim et al., 1998).
Organophosphorus and carbamate compounds may not be the only pollutants
that cause AChE inhibition in fish, chemicals like zinc, mercury, cadmium and
copper can cause some inhibition of AChE, meaning that AChE activity may
not be especially diagnostic for pesticide poisoning (Heath, 1995). The
inhibition of brain AChE activity may remain for several weeks following
exposure to pesticides. The inhibition of blood cholinesterase activity is also
indicative of exposure to a toxicant but the inhibition of enzyme activity is
short-lived and more definitive results are usually obtained with brain AChE
than with blood AChE (Melancon, 1995). The higher the AChE level in a
tissue, the more susceptible it is to inhibition with inhibition the greatest in the
brain, followed by muscle, gill and liver (Heath, 1995). The effectiveness of
using AChE inhibition as an indicator of pollution in field-collected animals
depends on the quality of the reference values so that possible inhibited
Page 28
Chapter 5 – Field Studies
101
samples show a significant difference from the control values (Melancon,
1995). For continuous exposure studies, blood sampling has the advantage
that the organism does not have to be killed for sampling and serial samples
may be collected form the same animal. Keeping the above -mentioned in
mind and looking at Figure 5.8, the red blood cell AChE activity was extremely
low for all three sampling sites. The standard deviation was also very high
and in future field surveys; brain AChE activity should be measured for more
accurate results. Blood AChE should not be used as part of a battery of
biomarkers, as it gives unreliable results. However, the possible use of brain
AChE as an indicator of toxicant exposure should be investigated in the
future.
∗-Aminolevulinic acid dehydratase (ALAD) is an important enzyme in the
haem synthesis, converting ∗-Aminolevulinic acid (ALA) to porphobilinogen
and ferrochelatase, and inserting iron into protoporphyrinogen (Nussey,
1994). The ALAD enzyme activity is found in almost every tissue since it also
participate in the synthesis of all other haem proteins. Red blood cells are
formed in the spleen and kidney and an increase in ALAD in the red blood
cells would indicate stimulation of haem synthesis in these two organs
(Nussey, 1994; Westman et al., 1975). The inhibition of ALAD activity is
specific for lead (Wepener, 1990; Johansson-Sjöbeck and Larsson, 1979) and
reduction in the enzyme activity occurs rapidly and can be detected at
exposure concentrations near “no effect” level (Schmitt et al., 1984). Lead is
a naturally occurring heavy metal and is widely distributed by industrial
activities (Ho and Ho, 1997) and the contamination of natural waters by lead is
caused by activities related to increasing mining operations and industrial use
of lead (Tewari et al., 1987). Fish exposed to cadmium, copper, zinc and
mercury showed no erythrocyte ALAD inhibition, indicating that this enzyme is
quite specific to lead (Johansson-Sjöbeck and Larsson, 1979). Mining
activities release a large amount of metals, including lead, into the
environment. The significant decrease (p<0.05) in ALAD enzyme activity
obtained during the field evaluation shows the possibility of high lead
concentrations present in the Loskop Dam catchment. When investigating an
Page 29
Chapter 5 – Field Studies
102
area with possible lead contamination, or when conducting lead exposure
experiments in the laboratory, the inhibition of ALAD enzyme activity should
be included in the battery of biomarkers because of the ALAD enzyme
sensitivity to lead.
Ethoxyresorufin-O-deethylase (EROD) is a sensitive indicator of Cytochrome
P1501A and high levels of environmental contaminants by chemicals may
result in an increase in this mixed-function oxidase (MFO) activity in fish
(Chen et al., 1998; Parke, 1981). Organic pollutants like PAHs, PCBs and
dioxins as well as complex mixtures including municipal and industrial
effluents cause the induction of EROD activity (Jimenez and Stegeman,
1990), explaining the induction seen in the Hartebeespoort Dam during the
two surveys. EROD levels from fish caught in the Loskop Dam showed
inhibition of EROD activity during the second survey. It has been reported
that heavy metals have an inhibitory effect on cytochrome P4501A induction
in fish hepatoma cells (Chen et al., 1998; Gagné and Blaise, 1993). Copper
inhibits enzyme activity by binding to SH-residues of the enzymatic proteins of
the MFO system, or as a consequence of the enhancing of lipid peroxidation
of the membranes. This leads to a more general alteration of the structure
and function of the endoplasmic reticulum (Stien et al., 1997). High levels of
PCB congeners and metals such as cadmium also cause inhibition of
cytochrome P4501A mediated catalytic activity (Jakšic et al., 1998).
Environmental variables such as seasonal changes, which are associated
with temperature and sexual factors, age and nutritional status of the fish, are
the most important influences on the MFO activity of fish (Jimenez and
Stegeman, 1990). The EROD enzyme activity proved to be a relative
sensitive biomarker for use in field surveys. By concentrating the enzyme
concentration even more during the preparation stage of the assay, it could be
possible to increase the sensitivity of the assay. Before using EROD enzyme
synthesis as an indicator of pollution, chemical analysis of the river or
impoundment studied, should be carried out to determine the specific
constituents and toxicant levels present in the water. The analysis will explain
the results since liver EROD enzyme activity can be both inhibited and
inducted by pollutants, as mentioned above.
Page 30
Chapter 5 – Field Studies
103
In fish, an increase in blood glucose levels and a decrease in liver glycogen
levels, are one of the first signs of stress on the carbohydrate metabolism
(Wepener, 1990). The rise in blood glucose, is the most characteristic general
response to stress, and occurs when the physical activity of the fish exceeds
what is normal (Love, 1980). Blood glucose levels are elevated in fish during
exposure to various pollutants, including pesticides and these stressful stimuli
elicit the rapid secretion of hormones, glucocorticoids and catecholamines,
from the adrenal tissue of fish, producing rapid hyperglycemia (Cerón et al.,
1996). Prolonged hyperglycemia could result in depletion of energy reserves
and an insufficient energy production. Continuous elevated blood glucose
levels cause a shift form aerobic to anaerobic metabolism and increases in
anaerobic metabolism is a response against the depletion of energy caused
by lack of oxygen (Cerón et al., 1996; Solomonson, 1981). Short-term
changes in glucose levels can be induced by handling stress, changes in
temperature, pH, water velocity, hypoxia, or other seasonal variations
(Folmar, 1993). To reduce the possibility of handling stress, all the fish from
the reference as well as the polluted sites were treated equally. The different
blood glucose levels obtained could be attributed to differences in rate and
degree of digestion, absorption and utilisation of glucose, due to an impaired
carbohydrate metabolism (Hilmy et al., 1980). When including plasma glucose
in a battery of biomarkers, it should be remembered that as already
mentioned, elevated blood glucose levels are a general response to stress,
including exposure to toxicants and other environmental changes like water
temperature variations.
Stress response in fish is generally characterised by an increase in adrenalin
causing mobilisation of liver glycogen into blood glucose (Swallow and
Flemming, 1970). The blood glucose levels do not necessarily reflect the
level of glycogen (Love, 1980). The high liver glycogen levels at the polluted
sites for the winter survey showed that the fish at the polluted sites had
greater liver glycogen stores than the reference site. Cortisol lowers the liver
glycogen and an increase in blood glucose with the depletion of liver glycogen
to stress. Metabolic consequence of cortisol impairment may be a reduced
Page 31
Chapter 5 – Field Studies
104
capacity to mobilise liver glycogen stores (Hontela et al., 1995). In the field
study, liver glycogen values showed no significant differences between the
reference and polluted sites for both the surveys. High standard deviations
were also obtained (Figure 5.5). When looking at all of the above-mentioned
information, there seems to be a large variation in liver glycogen levels under
normal conditions. It is thus difficult to identify changes in liver glycogen
levels caused by exposure to pollutants. Liver glycogen levels should not be
used as an indicator of pollutants and should not be included in a battery of
biomarkers.
Of the five biomarkers evaluated during this field study, only three biomarkers
showed significant results (p<0.05). ALAD, EROD and plasma glucose levels
can thus be included in a battery of biomarkers to be used as indicators of
exposure to pollutants. Even though these three biomarkers showed
significant results, the biomarkers will be even more accurate and sensitive at
higher levels of pollution. Although erythrocyte AChE and liver glycogen
showed no significant results during this study, it is still possible that these two
biomarkers will show more accurate and significant results at higher pollution
levels. Together with the use of biomarkers as indicators of deteriorating
water quality due to the influx of pollutants, chemical water analysis should
also be carried out. Chemical analysis will show what toxicants/pollutants are
present in the water while the biomarkers will show the level of effect of the
toxicant on the organisms. More reliable results will thus be obtained.
Page 32
Chapter 5 – Field Studies
105
5.7 References
ALABASTER, J.S. and LLOYD, R. (1980) Water Quality Criteria for
Freshwater Fish. Butterworths & Co. (Publishers) Ltd., London, 283
pp.
AUCAMP, P.J.; PIETERSE, S.A. and VIVIER, F.S. (1987) Health Problems
of the Hartebeespoort Dam. In: Hartebeespoort Dam – Quo Vadis?
(Eds. J.A. Thornton and R.D. Wamsley), FRD Ecosys. Prog. Occ.
Rep., 25: 83 – 93.
BESSELINK, H.T.; VAN SANTEN, E.; VORSTMAN, W.; VETHAAK, A.D.;
KOETMAN, J.H. and BROUWER, A. (1997) High Induction of
Cytochrome P4501A Activity without Changes in Retinoid and Thyroid
Hormone Levels in Flounder (Platichthys Flesus) Exposed to 2,3,7,8-
Tetrachlorodibenzo-p -dioxin. Environ. Toxicol. Chem., 16 (14): 816 –
823.
BKS (Pty) Ltd. (1992) Krokodilrivier (Wes-Transvaal) Opvangebiedstudie.
Watergehalte Situasie Ontleding van die Bo-Krokodilrivier en
Hartebeespoortdam. Verslag no. P A200/00/2792.
BKS (Pty) Ltd. (1998) Development of an Integrated Water Resource Model
of the Upper Olifants River (Loskop Dam) Catchment. Water
Requirements and Return Flows , Report No. PB B100/00/0698.
BRADFORD, M.M. (1976) A Rapid and Sensitive Method for the Quantitation
of Microgram Quantities of Protein Utilizing the Principle of Protein-dye
Binding. Analytical Biochemistry, 72: 248 – 254.
BURKE, M.D. and MAYER, R.T. (1974) Ethoxyresorufin: Direct Fluorimetric
Assay of a Microsomal O-dealkylation which is Preferentially Inducible
by 3-Methylcholanthrene. Drug Metabolism and Disposition, 2(6): 583
– 588.
Page 33
Chapter 5 – Field Studies
106
BUTTY, M.; WALMSLEY, R.D. and ALEXANDER, C.J. (1980) Rust Der
Winter Dam. In: The Limnology of Some Selected South African
Impoundments, (Eds. R.D. Walmsley and M. Butty), Published by The
Water Research Commission, pp. 71 – 80.
CERÓN, J.J.; SANCHO, E.; FERRANDO, M.D.; GUTIERREZ, C. and
ANDREU, E. (1996) Metabolic Effects of Diazinon on the European
Eel Anguilla anguilla. J. Environ. Sci. Health, B31 (5): 1029 – 1040.
CHEN, C-M.; UENG, T-H; WANG, H-W; LEE, S.Z. and WANG, J.S. (1998)
Microsomal Monooxygenase Activity in Milkfish (Chanos chanos) from
Aquaculture Ponds near Metal Reclamation Facilities. Bull. Environ.
Contam. Toxicol., 61: 378 – 383.
DEPARTMENT OF WATER AFFAIRS AND FORESTRY (DWAF). (1996)
South African Water Quality Guidelines (second edition). Volume 6:
Agricultural Use: Aquaculture, pp. 46, 75, 183.
DU PLESSIS, J.A. and MARÉ, H.G. (1999) Development of an Integrated
Water Resource Model of the Upper Olifants River (Loskop Dam)
Catchment. Base Conditions, DWAF Report No. PB B100/00/0598.
ELLMAN, G.L.; COURTNEY, D.; ANDRES, V. (JR) and FEATHERSTONE,
R.M. (1961) A New and Rapid Colorimetric Determination of
Acetylcholinesterase activity. Biochemical Pharmacology, 7: 88 – 95.
FOLMAR, L.C. (1993) Effects of Chemical Contaminants on Blood Chemistry
of Teleost Fish : A Bibliography and Synopsis of Selected Effects.
Environmental Toxicology and Chemistry, 12: 337 – 375.
GAGNÉ, F. and BLAISE, C. (1993) Hepatic Metallothionein Level and Mixed
Function Oxidase Activity in Fingerling Rainbow Trout (Oncorhynchus
Page 34
Chapter 5 – Field Studies
107
mykiss) after Acute Exposure to Pulp and Paper Mill Effluents. Wat.
Res., 27 (11): 1669 – 1682.
GROBLER, D.F.; KEMPSTER, P.L. and VAN DER MERWE, L. (1994) A
Note on the Occurrence of Metals in the Olifants River, Eastern
Transvaal, South Africa. Water SA, 20 (3): 195 – 204.
HEATH, A.G. (1995) Water Pollution and Fish Physiology. Second Edition,
CRC Press Inc., Florida, 359 pp.
HEATH, R.G.M. and CLAASSEN, M. (1999) An Overview of the Pesticide
and Metal Levels Present in Populations of the Larger Indigenous Fish
Species of Selected South African Rivers. WRC Report No. 428/1/99,
318 pp.
HELY-HUTCHINSON, J.R. and SCHUMANN, E.H. (1997) The Anatomy of a
Flash Flood in the Hartebeespoort Dam Catchment. Water SA, 23 (4):
345 – 356.
HILMY, A.M.; SHABANA, M.B. and SAIED, M.M. (1980) Blood Chemistry
Levels after Acute and Chronic Exposure to HgCl2 in the Killifish
Aphanius dispar (Rupp). Water, Air and Soil Pollution, 14: 409 – 417.
HO, J.W. and HO, A.W. (1997) Environmental Lead Exposure Induces
Changes in the Heme Biosynthetic Pathway. Environ. Toxicol. Water
Qual., 12: 245 – 248.
HONTELA, A.; DUMANT, P.; DUCLOS, D. and FORTIN, R. (1995)
Endocrine and Metabolic Dysfunction in Yellow Perch, Perca
flavescens, Exposed to Organic Contaminants and Heavy Metals in the
St. Lawrence River. Environmental Toxicology and Chemistry, 14 (4):
725 – 731.
IBRAHIM, H.; KHEIR, R.; HELMI, S.; LEWIS, J. and CRANE, M. (1998)
Effects of Organophosphorus, Carbamate, Pyrethroid and
Page 35
Chapter 5 – Field Studies
108
Organochorine Pesticides and a Heavy Metal on Survival and
Cholinesterase Activity of Chironomus reparius Meigen. Bull. Environ.
Contam. Toxicol., 60: 448 – 455.
JAKŠIK, Ž.; BIHARI, N.; MÜLLER, W.E.G.; ZAHN, R.K. and BATEL, R.
(1998) Modulation of Cytochrome P4501A in Sea Bass Liver by Model
Substances and Seawater Extracts. Aquatic Toxicology, 40: 265 –
273.
JAMES, K.M and VAN WYK, N.J. (1993) An Overview of the Surface Water
Resources in the Olifants River Basin. Imiesa, September, pp. 3 – 11.
JIMENEZ, B.D. and STEGEMAN, J.J. (1990) Detoxication Enzymes as
Indicators of Environmental Stress on Fish. In: Biological Indicators of
Stress in Fish. (Edited by S.M. Adams), American Fisheries
Symposium 8, Bethesda, Maryland, pp. 67 – 79.
JOHANSSON-SJÖBECK, M-L. and LARSSON, Å. (1979) Effects of
Inorganic Lead on Delta-Aminolevulinic Acid Dehydratase Activity and
Hematological Variables in Rainbow Trout, Salmo gairdnerii. Arch.
Environm. Contam. Toxicol., 8: 419 – 431.
KARR, J.R. AND CHU, E.W. (1999a) Key Concepts for Using Watershed
Biological Indicators. From: Restoring Life in Running Water.
http://www.epa.gov/ceisweb1/ceishome/atlas/bioindicators/keyconcepts
.html
KARR, J.R. AND CHU, E.W. (1999b) Premise 8: Understanding Biological
Responses Requires Measuring Across Degrees of Human Influence.
From: Restoring Life in Running Water.
http://www.epa.gov/ceisweb1/ceishome/atlas/bioindicators/premise_8.h
tml
Page 36
Chapter 5 – Field Studies
109
LOVE, R.M. (1980) The Chemical Biology of Fishes. Volume 2: Advances
1968 – 1977. Academic Press Inc., London, 943 pp.
MELANCON, M.J. (1995) Bio-indicators Used in Aquatic and Terrestrial
Monitoring. In: Handbook of Ecotoxicology. (Eds. D.J. Hoffman, B.A.
Rattner, G.A. Burton (Jr) and J. Cairns (Jr)), Lewis Publishers, Boca
Raton.
NATIONAL INSTITUTE FOR WATER RESEARCH. (1985) The Limnology of
Hartebeespoort Dam . South African National Scientific Programmes
Report No. 110, 269 pp.
NUSSEY, G. (1994) The Effect of Copper on the Blood Coagulation and
General Haematology of Oreochromis mossambicus (Cichlidae).
M.Sc.-Thesis, RAU.
PARKE, D.V. (1981) Cytochrome P-450 and the Detoxication of
Environmental Chemicals. Aquatic Toxicology, 1: 367 – 376.
PEAKALL, D. (1992) Animal Biomarkers as Pollution Indicators. Chapman &
Hall, London, 290 pp.
ROSSOUW, J.D. (1992) Krokodilrivier (Wes-Transvaal)
Opvangsgebiedstudie. Beskrywing van die Fisiese Waterstelsel,
Departement van Waterwese, Verslag No. PA200/00/1292.
ROUX, D.J. (1994) Role of Biological Monitoring in Water Quality
Assessment and a Case Study on the Crocodile River, Eastern
Tranvaal. M.Sc-Thesis, RAU, 130 pp.
SASTRY, K.V. and SHUKLA, V. (1994) Acute and Chronic Toxic effects of
Cadmium on Some Haematological, Biochemical and Enzymological
Parameters in the Fresh Water Teleost Fish, Channa punctatus. Acta
hydrochim. Hydrobiol., 22(4): 171 – 176.
Page 37
Chapter 5 – Field Studies
110
SCHALLER, K-H and BERLIN, A. (1984) ∗ -Aminolaevulinate Dehydratase.
In: Methods of Enzymatic Analysis. Volume IV, Enzymes 2: Esterases,
Glyucosidases, Lyases, Ligases. (Eds. H.U. Bergmeyer, J. Bermeyer,
and M. Gra∃l), Verlag Chemie, Weinheim, pp. 363 - 368.
SCHMITT, C.J.; DWYER, F.J. and FINGER, S.E. (1984) Bioavailability of Pb
and Zn form Mine Tailings as Indicated by Erytrhocyte ∗-Aminolevulinic
Acid Dehydratase (ALA-D) Activity in Suckers (Pisces: Catostomidae).
Can. J. Fish. Aquat. Sci., 41: 1303 – 1040.
SEIFTER, S.; DAYTON, S.; NOVIC, B. and MUNTWYLER, E. (1950) The
Estimation of Glycogen with the Anthrone Reagent. Arch. Biochem.,
25: 191 – 200.
SOLOMONSON, L.P. (1981) Cyanide as a Metabolic Inhibitor. In: Cyanide
in Biology. (Eds. B. Vennesland; E.E. Conn; C.J. Knowles; J. Westley
and F. Wissing), Academic Press Inc., London, 548 pp.
SOUTH AFRICAN NATIONAL COMMITTEE ON LARGE DAMS (SANCOLD).
(1978) Loskop Dam. In: Typical Large Dams in South Africa,
Published by CIGB ICOLD.
STIEN, X.; RISSO, C.; GNASSIA-BARELLI, M.; ROMÉO, M. and LAFAURIE,
M. (1997) Effect of Copper Chloride in vitro and in vivo on the Hepatic
EROD Activity in the Fish, Dicentrarchus labrax. Environmental
Toxicology and Chemistry, 16 (2): 214 – 219.
SUTTON, D.F. and OLIVEIRA, M.P. (1987) Hartebeespoort Dam as a
Receiver of Return Flows. In: Hartebeespoort Dam – Quo Vadis?
(Eds. J.A. Thornton and R.D. Walmsley), FRD Ecosys. Prog. Occ.
Rep., 25: 49 – 61.
Page 38
Chapter 5 – Field Studies
111
SWALLOW, A. and FLEMMING, P. (1970) The Effect of Oxaloacetate,
ACTH and Cortisol on the Liver Glycogen Levels of Tilapia
mossambica. Comp. Biochem. Physiol., 36: 93.
TEWARI, H.; GILL, T.S. and PANT, J. (1987). Impact of Chronic Lead
Poisoning on the Hematological and Biochemical Profiles of a Fish,
Barbus conchonius (Ham). Bull. Environ. Contam. Toxicol., 38: 748 –
752.
THERON, PRINSLOO, GRIMSEHL and PULLEN (Pty) Ltd. (1991a) Water
Resource Planning of the Olifants River Basin. Basin Study Report,
Volume 3, Part 3, Sub-catchment B310, Report No. PB B100/00/0591.
THERON, PRINSLOO, GRIMSEHL and PULLEN (Pty) Ltd. (1991b) Water
Resources Planning of the Olifants River Basin. Basin Study Report,
Annexure 16, Part 1, Water Availability from Major Dams, Basin
Upstream of Loskop Dam, Report No. PB 000/00/3491.
VAN RIET, W.F. (1987) The Hartebeespoort Dam – A magnet to millions?
In: Hartebeespoort Dam – Quo Vadis? (Eds. J.A. Thornton and R.D.
Walmsley), FRD Ecosys. Prog. Occ. Rep., 25: 83 – 93.
VAN VUREN J.H.J.; DU PREEZ, H.H.; WEPENER, V.; ADENDORFF, A.;
BARNHOORN, I.E.J.; COETZEE, L.; KOTZÉ, P. and NUSSEY, G.
(1999) Lethal and Sublethal Effects of Metals on the Physiology of
Fish: An Experimental Approach with Monitoring Support. WRC
Report No. 608/1/99.
WEPENER, V. (1990) Die Effek van Swaarmetale by Variërende pH op die
Bloedfisiologie en Metaboliese Ensieme van Tilapia sparrmanii
(Cichlidae). M.Sc-Verhandeling, RAU.
WESTMAN, I. JOHANSSON-SJÖBECK, M-L. and FÄNGE, R. (1975) The
Effect of PCB on the Activity of Delta -Amino-Levulinic Acid
Page 39
Chapter 5 – Field Studies
112
Dehydratase (ALA-D) on Some Hematological Parameters in the
Rainbow Trout, Salmo gairdneri. In: Sublethal Effects of Toxic
Chemicals on Aquatic Animals. (Eds. J.H. Koeman and J.J.T.W.A.
Strik), Elsevier Scientific Publishing Company, Amsterdam, pp. 111 –
118.
WITTAKER, M. (1984) Cholinesterase: Acetylcholinesterase. In: Methods of
Enzymatic Analysis. Volume IV, Enzymes 2: Esterases, Glyucosidases,
Lyases, Ligases. (Eds. H.U. Bergmeyer, J. Bermeyer, and M. Gra∃l),
Verlag Chemie, Weinheim, pp. 52 – 63.