Makalah Self Purification
Makalah Self Purification
Makalah Self Purification
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CONTENT
1.
Introduction.......................................................................................
1
2.
3
3.
8
4.
Conclusion.........................................................................................
9
5.
Reference..........................................................................................
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SELF PURIFICATION
1. Introduction
Running water is capable of purifying itself with distances through a process known as self
purification. This is the ability of rivers to purify itself of sewage or other wastes naturally.
It is produced by certain processes which work as rivers move downstream. These
mechanisms can be inform of dilution of polluted water with influx of surface and
groundwater or through certain complex hydrologic, biologic and chemical processes such
as sedimentation (behind obstruction), coagulation, volatilization, precipitation of colloids
and its subsequent settlement at the base of the channel, or lastly due to biological
uptake of pollutants. On the other hand, certain streams are capable of adding-up more
materials as they flow downstream from riparian inputs (Ongley, 1987; 1991).
Quality of water is of paramount importance because of its role to human health, aquatic
life, ecological integrity and sustainable economic growth. Indeed, without good quality
water sustainable development and environmentally sound management of water
resources will be meaningless. For example, on a global scale, water borne disease is
estimated to be responsible for about 3 million deaths and also to render sick a billion
people (World Bank, 1993).
The extent of self purification in any stream depend on certain factors some of which are:
temperature; level of river; river velocity; amount of inorganic compound in the stream
and the arrow; distribution and types of aquatic weeds along the channel. If the
concentration of oxidisable material be excessive, the river-water will suffer considerable
or complete deoxygenating, and a nuisance will result owing to the septic condition
caused by the anaerobic decomposition of the organic matter. On the other hand, if there
be sufficient dilution, the organic matter can be oxidized and thus destroyed without
depriving the river-water of oxygen to any appreciable degree. The suspended matter will
also be sediment in the form of a thin film distributed over a considerable area of riverbed, and no nuisance will thus result through the formation of foul mud-banks.
Recovery from pollution, or self-purification, as it is termed, thus depends on the
conditions obtaining with the particular river. Ordinarily towns situated on the same river
are sufficiently separated to give time for the river to recover from the effects of the
upper pollution before it is subjected to the next. On the other hand, if towns be close
together, a nuisance may result, and the river may become unfit to receive a further
volume of sewage lower down, until a considerable length of time and dilution from
tributaries enable purification to be effected.
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the characters of river-waters in this country, we find that they necessarily depend on the
nature of the soils and rocks over which they flow.
pH
Particulate suspension
temperature profiles
atmospheric loadings
aquatic eco-community
In the following text, some insight is given into the most important of the above factors:
oxidation, biological activity and sedimentation.
Oxidation
As in rivers, the major inputs of dissolved oxygen are from atmospheric re-aeration,
exchange mechanisms with water richer in oxygen e.g. rainfall, photosynthesis and, in
some circumstances, chemical reduction of nitrate and sulphate. The major demands on
the oxygen are from biological and chemical processes in the hypolimnion and sediments.
The assimilative capacity of a lake and the resulting dissolved oxygen levels are normally
determined as part of the overall oxygen budget. The process is similar to that used for
streams but there are some important differences. Thermal stratification separates the
major input (surface aeration) from the major demand (sediments). Further, in lakes both
the sediment and water column demands are functions of dissolved oxygen levels in the
water. Sediment demands for eutrophic waters are 0.5 to 3.0 g O2/m2/day and a change of
4 mg/l of dissolved oxygen doubles the demand (see Polak and Haffner, 1978).
Some measurements of sediment oxygen demand have been related to the percentage of
organic matter (dry weight) for lakes. In a typical case, 1.3, 12 and 30 - 40 percent
organic matter have oxygen uptake rates of 0.01, 0.1 and 0.15 - 9.18 g O2/m2/h at 15C
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(Edberg and Hofsten, 1973). It should be noted that these rates are temperature specific
and quite different uptake rates will be found at 10C and 20C. Further, the general
applicability of these results must be tempered by sediment depth and its physical,
chemical and biological quality. The final uptake will depend on whether only the surface
of the sediment requires oxidation or the sediment is being disturbed so that demands in
the sediment are exerted at depth. Bacterial and macro invertebrate respiration and
nitrate concentrations at the sediment water interface may also be factor affecting
oxygen uptake rates. Water oxygen demands in an urban area are 0.2 to 1.0 g 0 m d and
double
for
an
mg/l
rise
in
dissolved
oxygen
levels
in
the
water
(see Polak and-Haffner, 1978). Both these demands are also a function of temperature,
although the temperature variation in many lakes is small compared to the variation of
the dissolved oxygen levels in the water.
The oxygen demand is also frequently spatially variable. Measuring changes in biological
oxygen demand and chemical oxygen demand in the effluent plume of a shoreline
discharge in a large lake would normally require tracing the plume for periods in excess of
12 to 14 hours (Polak and Palmer, 1977).
Atmospheric re-aeration is highly variable and difficult to measure. Normally all the other
inputs and demands are measured and the re-aeration determined by difference. In most
instances re-aeration ranges from 1 to 9 g O2/m2/day.
The maintenance of reasonable dissolved oxygen levels allows the conversion of
potentially biologically toxic chemicals like H2S and ammonia into less harmful
components. These oxidations are normally very rapid and usually are a function of the
dissolved oxygen stock available. Depending on the length of thermal stratification, reaeration of the hypolimnion may be impossible for months. Full oxygen depletion and
significant concentrations of H2S will persist for long. In water free of oxygen above the
sediments and within the sediments, remobilization especially of iron and manganese
occurs. The presence of either of these metals causes great difficulties for drinking water
supply. They form deposits in the pipes and removal requires expensive treatment.
Discharges of free chlorine are normally reduced rapidly in a natural environment
provided ammonia is not present, in which case toxic chloramines are formed. Very low
oxygen levels allow reducing conditions to establish which can release nutrients from the
sediments particularly phosphorus thereby enriching the water column above at the time
of overturn. A water column with alternating oxic and anoxic conditions at the sediment
interface can act as a nutrient sink for nitrogen through nitrification/denitrification
(Keeney, 1973).
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Biological activity
If slightly polluted water courses are impounded, the biological activity becomes much
more intensive both in time and space than it was in flowing water, i.e. the degradation
effect on the constituents of the waste water is increased. This effect is utilized in
drinking water reservoirs, which, if properly dimensioned have such biological activity
that a considerable improvement in water quality is achieved (Lack and Collingwood,
1975).
The material budget of storage reservoirs and lakes is governed by the phytoplankton. The
production of algae is a function of nutrient availability, light and efficiency of its
utilization, grazing intensity of the water column and in some instances, the presence of
toxins or parasites. In many instances, algal production, can be related to total
phosphorus loading (Vollenweider, 1968, Vollenweider and Dillon, 1974) but predictions of
the effects of phosphorus on algal production have been questioned (Thomann, 1977)
because of the interactions of other variables. The effects of these other variables have
been expressed in production models (Bannister, 1974 and Lehman et al., 1975) which may
be better ways to produce predictive statements on algal production. As more variables
are considered the prediction becomes more realistic. A model especially developed for
shallow lakes was described by Oskam, (1973).
As a result of eutrophication, production rates are often accelerated, resulting in high
standing crops of algae and locally high concentrations of dissolved oxygen. However, the
eventual decomposition of these populations results in high oxygen demands being
exerted on the water column. Should these demands exceed rates of oxygen supply,
anoxic conditions will arise.
Extensive zones of macrophytes have some effect on the concentration of nutrients and
degradation of pollutants carried by the incoming water. However, their presence also
involves problems of silting and the accumulation of persistent pollutants, particularly
heavy metals, in the sediments.
Bacteria are an important component of the bioactivity in lakes and reservoirs
(Golterman, 1975). Consequently, productivity estimates for bacteria should be made.
Unfortunately little information is available on the nutrient requirements for bacteria in
general other than some efforts to correlate bacterial populations with some nutrients.
Heterotrophs, nitrosomonas and nitrifying bacterial populations are important in the
assimilation of nutrients. The presence of pathogenic bacteria limits water use for both
drinking and body contact recreation. Estimates of bacterial populations are prone to at
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least three major influences. The time of sampling is critical as Bellair et al. (1977) have
shown that sunlight can produce a two-fold variation in population numbers within a day.
Water temperature is a further important influence. Geldreich (1968) has shown that
Salmonella typhimurium, Escherichia coli and Aerobacter aerogenes were reduced by 90%
after 1.3, 1.9 and 3.8 days respectively when stored in storm water at 200C, while Faecal
streptococci were only reduced by 83% after 14 days. At 100C the 90% reduction in
numbers took 7.6, 9.3 and 5.8 days respectively. Faecal streptococci were reduced by 48%
after 14 days at this temperature.
Distance and time of travel from the discharge is another influence to be taken into
account (Zanoni et al.,1978). The decay of the bacteria with distance is normally
computed by considering the time of travel and water temperature. Receiving bodies of
water vary in their characteristics which affect the decay rates and the extent must be
assessed with actual field measurements. To permit an assessment of this variation,
bacterial surveys
are normally required over a period of time (approximately 5 days) with at least two
samples per day. Data from these surveys are useful for developing decay rates for the
bacteria. Generally, bacterial decay is a power function of distance from the sourceTypically, reductions of bacterial levels by a factor of two orders of magnitude occur in a
kilometer along the shoreline of large lakes (Cherry et al., 1974).
Perhaps the most important roles of bacteria are their part in the assimilative capacity of
the body of water. They break down large organic molecules, help stabilize organic
content in sediments, and even break down harmful toxins. This assimilative capacity
generates high oxygen demands. Stabilization of sediments can result in an oxygen uptake
rate in the range of 0.05 to 9.18 g O2 m .d at 20oC and the water column demands are
frequently greater.
There is evidence that bottom fauna assimilates insecticides, polychlorinated biphenyls,
heavy metals and radioactivity. This fauna is the major source of these substances in fish,
either directly or through predation. While the bottom fauna is probably not a significant
mechanism for the removal of these substances from water compared to other
mechanisms, it is a good indicator of the degree of contamination.
Sedimentation
Pollutants discharged to near shore waters by industries and sewage treatment plants are
either transported in dissolved, complexes or suspended form to offshore waters or
sediment by various competing mechanisms. Numerous studies have been performed on
the mechanisms of trace metal transport and sedimentation in rivers, lakes and estuaries.
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Important mechanisms of sedimentation (Gibbs, 1973; Stumm and Morgan, 1970) include
incorporation in inert crystalline structures such as various silicate minerals; precipitation
and co-precipitation as oxides, hydroxides, carbonates, sulphides etc; absorption (physical
and chemical) on minerals such as clays; or biological incorporation in sedimentation.
The chemical nature of the sediment-water interface plays a profound role in the
sedimentation of pollutants and their possible re-solution. It has been known for many
years (Mortimer, 1941) that anaerobic conditions in the overlying water allow reduction
and mobilization of absorbed or co-precipitated phosphate and silica. Other factors
governing the mobilization of trace metals include:
Microbial activity can affect the physical and redox properties of sediments
bringing about reducing conditions. Bacteria are also involved in the formation of
soluble organometallic compounds (e.g. with mercury).
An example may be drawn from the Hamilton Harbour study (Ontario, Ministry of the
Environment, 1974) to illustrate the importance of these sedimentation mechanisms on
the control of heavy metal concentration in water. If cultural and natural inputs to
Hamilton Harbour, but not lake exchange, are considered the average residence time for
water in the harbour is 1.25 years. Using this figure expected metal concentrations were
computed for iron, chromium and zinc from industrial data and compared with the
observed concentrations. The fraction remaining in solution is generally 5 per cent or less.
Although some of the metal may have been removed by lake exchange, the majority is
undoubtedly in the sediments. The greatest concentration of heavy metals is found in
deep water sediments or adjacent to the discharges. These produce enrichment of surface
layers of the sediment compared to deeply buried layers. Average enrichment factors
(top/bottom concentration ratios in cores) of up to six have been found in Hamilton
Harbour. In Lake Erie, Walters et al. (1974) found concentration factors of 2 to 50 for
different metals.
Contamination of deep water sediments is indicative of polluting discharges. However,
determining the pattern of contamination and relating it to the discharge is difficult. It is
suggested that if contamination is indicated in the sediments or biological species,
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d. Clean Zone
In this zone, the river has reached a condition as before. Aquatic life in this zone is
also back as before. A few pathogenic organisms may still live in this zone, because the
river water that has been contaminated once cannot be used as drinking water
although it has been processed. Dissolved oxygen conditions in the clean zone located
at 8 ppm, which is the normal concentrations DO in the waters and on the conditions
of low BOD. In this zone the animals require oxygen water in concentrations normally
grow well.
e. Capacity
Capacity is the ability of the water pollution in a water source to receive input load of
pollution without resulting in the water became polluted. Water pollution can occur
any other substances/elements that go into the water, causing the water quality to be
down. These elements can be derived from elements of the non-conservative
(relegated) and conservatives (the element that is not degraded).
4. Conclusion
5. Reference
Bellair, J. T., Parr-Smith, G. A.and Wallis, I. G., 1977. Significance of diurnal variations in fecal
coliform die-off rates in the design of ocean outfalls. J. Water Pollution Control Fed. 49,
2022-2030.
Cherry, D. S., Guthrie, R. C. and Harvey, R.S., 1974. Temperature influence on bacterial populations
in three aquatic systems. Wat. Res. 8, 149-155.
Edberg, N. and Hofsten, B.V., 1973. Oxygen uptake of bottom sediments studied in situ and in the
laboratory. Water Res. 7, 1285-1294.
Forstner, U., 1976. Forms and sediment association of trace metals. Presented at "Fluvial transport
of sediment-associated nutrients and contaminants". PLUARG Workshop. Kitchener,
Ontario, October, 1976.
Geldreich, E. E., 1968. Bacteriological aspects of storm water pollution. J. Wat. Pollut. Control.
Fed. 40, (11), 1861-1872.
Gibbs, R. J., 1973. Mechanisms of trace metal transport in rivers. Science, 180, 71-73.
Golterman, H. L., 1975. Physiological limnology. Developments in Water Science 2, Elsevier,
Amsterdam.
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Keeney, D. R., 1973. The nitrogen cycle in sediment-water systems. J. Environ. Qual. 2, 15-29.
KnOpp, H., 1968. Stoffwechseldynamische Untersuchungsverfahren far die biologische
Wasseranayse. Int. Revue. ges. Hydrobiol. 53, (3), 409-441.
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