2.1 Leachate Characterization
Prior to 1965 very few people were aware of the fact that water passing
through solid waste in a sanitary landfill would become highly contaminated. This
water, termed leachate, was generally not a matter of concern until few cases of
water pollution were noted where leachate had caused harm (Boyle and Ham 1974).
Boyle (1974) also reported that many contaminants released from sanitary landfill, if
allowed to migrate, may pose a severe threat to surface and ground water.
2.1.2 Factors Affecting Leachate Quality
Leachate quality was highly variable. It contains larger pollutant loads than
raw sewage or many industrial wastes. The variation in leachate quality are
generally attributed to a myriad of interacting factors such as type and depth of solid
waste, age of fill, the rate of water application, landfill design and operations, and the
interaction of leachate with its environment. The effects of some of these variables
upon leachate quality are presented below. The quality variations can also be
attributed to sampling procedures, sample preservation, handling and storage, and
analytical methods used to characterize the leachate.
220.127.116.11 Processed Refuse
Leacahte characteristics from shredded or baled refuse fills also differ greatly.
Lu et al. (1984) indicated that leachates from shredded fills has significantly higher
concentrations of pollutants than those from unshredded refuse. Attainment of field
capacity is also delayed, but the rate of pollutant removal, solid waste decomposition
rate, and the cumulative mass of pollutants released per unit volume is significantly
increased when compared with unshredded fills.
18.104.22.168 Depth of Refuse
One of the earlier studies performed by Qasim and Chiang (1994) reported
that substantially greater concentrations of constituents are obtained in leachates
from deeper fills under similar conditions of precipitation and percolation. Deeper
fills, however, require more water to reach saturation, require longer time for
decomposition, and distribute the bulk of extracted material over a longer period of
time. Water entering from surface of the landfill and travelling down through the
refuse will successively transfer to the percolating water. Deep fills offer greater
contact time and longer travel distance, thus higher concentrations will result.
22.214.171.124 Age of Fills
Variation of leachate quality with age of fill is expected, because organic
matter will continue to undergo stabilization. It should be noted that release of
constituents from solid waste is obviously governed by decomposition processes, and
the rate and volume of water infiltrating through the fill. Age is merely a convenient
means of measuring and monitoring changes in leachate composition, and extraction
of pollutants from the refuse bed. As a result, many studies describe leachate quality
as a function of time based on water input rate and leachate generation.
Lu et al. (1984) indicated that pollutant concentration in leachate peak in
early life (that is within 2-3 years), followed by a gradual decline in ensuing years.
This trend applies to most of the constituents, but in particular to organic indicators
(e.g. BOD, COD, TOC), and microbiological population. Most other constituents
exhibit steady decreases in concentration over 3 to 5 years due to continued flushing
of the refuse bed. Among these are iron, zinc, phosphate, chloride, sodium, copper,
organic nitrogen, total solids and suspended solids. In some cases however, the
concentration of heavy metals fluctuated because of precipitation, dissolution,
adsorption and complexation mechanisms that may retain or mobilize the metals
within the landfill microenvironments.
Early research indicated that pathogenic bacteria are present in fresh leachate.
Viruses are occasionally detected. There is however, a significant deactivation in
bacterial and viral populations with solid waste age. This is attributed to adverse
environmental conditions such as initial general associated with solids decomposition
Table 2.1 below shows the comparison between new landfill which is
operating less than 2 years with a mature landfill which operating age is greater than
10 years. Large differences can be observed from the constituent shown in Table
Table 2.1 Typical Data on the Composition of Leachate from New and Mature
New landfill Mature landfill
(less than 2 years) (greater than 10
Constituent Range Typical Range
BOD5(5-day biochemical oxygen 2,000-30,000 10,000 100-200
TOC (total organic carbon) 1,500-20,000 6,000 80-160
COD (chemical oxygen demand) 3,000-60,000 18,000 100-500
Total suspended solids 200-2,000 500 100-400
Organic nitrogen 10-800 200 80-120
Ammonia nitrogen 10-800 200 20-40
Nitrate 5-40 25 5-10
Total phosphorus 5-100 30 5-10
Ortho phosphorus 4-80 20 4-8
Alkalinity as CaCO3 1,000-10,000 3,000 200-1,000
pH (no units) 4.5-7.5 6 6.6-7.5
Total hardness as CaCO3 300-10,000 3,500 200-500
Calcium 200-3,000 1,000 100-400
Magnesium 50-1,500 250 50-200
Potassium 200-1,000 300 50-400
Sodium 200-2,500 500 100-200
Chloride 200-3,000 500 100-400
Sulphate 50-1,000 300 20-50
Total Iron 50-1,200 60 20-200
Source: G. Tchobanoglous, H. Theisen, and S. Vigil, Integrated Solid Waste
management, New York: McGrawHill. 1993.
2.2 Constructed Wetlands
A wetland is a complex assemblage of water, substrate, plants, litter,
invertebrates and an array of micro organisms. Wastewater is purified during contact
with media and the roots of reeds in constructed wetlands. Physical, chemical and
biological processes are involved in wastewater treatment. Aerobic and anaerobic
bacteria growing on media and rhizomes are responsible for removal of organic
carbon and nitrogen in wastewater (Badjoubi et. al., 1998). The mechanisms that are
available to improve water quality are therefore numerous and often interrelated.
These mechanisms include:
i. Settling of suspended particulate matter
ii. Filtration and chemical precipitation through contact of the water with the
substrate and litter
iii. Chemical transformation
iv. Adsorption and ion exchange on the surfaces of plants, substrate, sediment
v. Breakdown, and transformation and uptake, of pollutants and nutrients by
micro organisms and plants
vi. Predation and natural die off of pathogens.
2.2.2 Types of Constructed Wetlands
Generally, constructed wetlands can be divided into two major types which
are free water surface constructed wetlands and sub surface flow constructed
126.96.36.199 Free Water Surface Treatment Wetlands
The FWS wetland technology started with the ecology engineering of nature
wetlands for wastewater treatment. Constructed FWS treatment wetlands mimic the
hydrological regime of natural wetlands. In surface flow (SF) wetlands, water flows
over the soil surface from an inlet point to an outlet point or, in a few cases, is totally
lost to evapotranspiration and infiltration within the wetland.
FWS treatment wetlands have some properties in common with facultative
lagoons and also have some important structural and functional differences. Water
column processes in deeper zones within treatment wetlands are nearly identical to
ponds with surface autotrophic zones dominated by planktonic or filamentous algae,
or by floating or submerged aquatic macrophytes. Deeper zones tend to be
dominated by anaerobic microbial processes in the absence of light. However,
shallow emergent macrophyte zones in treatment wetlands and aerobic lagoons can
be quite dissimilar. Emergent wetland plants tend to cool and shade the water. Net
carbon production in vegetated wetlands tends to be higher than that in facultative
ponds because of high gross primary production in the form if structural carbon,
accompanied by resistance to degradation and low rates of decomposition of organic
carbon in the oxygen-deficient water column. This high availability of carbon and
the short diffusional gradients in shallow vegetated wetlands result in differences in
biogeochemical cycling compared with ponds and lagoons. The net effect of the
complex processes is a general decrease in pollutant concentrations between the inlet
and outlet of treatment wetlands. However, because of the internal autotrophic
processes of the wetland, outflow pollutant concentrations are seldom zero, and in
some cases for some parameters they can exceed inflow concentrations. Figure 2.1
shows diagram of FWS wetland containing rooted, floating leaf plants.
Figure 2.1: FWS wetland containing rooted, floating leaf plants
188.8.131.52 Subsurface Flow Treatment Wetland
Many of the earliest treatment wetlands in Europe were SSF system
constructed to treat mechanically pre treated municipal wastewater. Soil and gravel
based SSF wetlands are still the most prevalent application of this technology in
Europe (Cooper et al. 1996; Brix 1994; Vymazal et al. 1998). SSF wetlands that use
gravel substrates have also been used extensively in United States. This technology
is generally limited to systems with low flow rates and can be used with less than
i) Horizontal-Flow system
Figure 2.2 shows a typical arrangement for the constructed wetland with a
horizontal flow (HF). In this system, wastewater is fed in at the inlet and flows
slowly through the porous medium under the surface of the outlet zone, where it is
collected and discharged at the outlet. During this passage, the wastewater will into
contact with a network of aerobic, anoxic and anaerobic zones. The aerobic zones
occur around roots and rhizomes that leak oxygen into the substrate. When the
wastewater passes through the rhizosphere, the wastewater is cleaned by
microbiological degradation and by physical and chemical processes (Cooper et al.,
Figure 2.2: Typical arrangement of horizontal system SSFCW
ii) Vertical Flow system
Vertical flow (VF) treatment wetlands are frequently planted with common
reed. Other emergent wetlands plants such as cattails or bulrush can also be used.
VF reed beds typically look like the system shown in Figure 2.3. They are composed
of a flat bed of gravel topped with sand, with reeds growing at the same sort of
densities as HF systems. They are fed intermittently. The liquid is dosed on the bed
in a large batch, flooding the surface. The Liquid then gradually drains vertically
down through the bed and is collected by a drainage network at the base. The bed
drains completely free, allowing air to refill the bed. The next dose of liquid traps
this air and this together with aeration caused by the rapid dosing on the bed leads to
good oxygen transfer and hence the ability to decompose BOD and to nitrify
ammonia nitrogen (Cooper et al. 1996).
Figure 2.3: Typical arrangement of vertical SFCW system
2.2.3 Application of Constructed Wetland in Leachate Treatment
As landfills become larger, the enormous quantities of putrescible wastes will
increase the potential to generate highly polluting leachates as they decompose
anaerobically over many years. Landfill leachates contain various quantities of
undesirable, and even toxic, organic and inorganic substances. Treatment of this
highly polluting wastewater is becoming mandatory worldwide. On site high tech
leachate treatment systems are also avoided because of high costs of construction and
Historically, aerated lagoons have been popular for the treatment of landfills
leachate and have proved to be successful in the removal of COD and NH4-N (Lu et
al., 1984). Wetland treatment of landfill leachates has been successfully tested at
several locations. A facility at Ithaca, New York, USA, that has been operating since
1989, has used SSF wetlands, SF wetlands have been operating successfully in
Escambia Country, FL, USA, since 1990 (Martin et al.1994). Cool climate systems
are functioning properly in Norway, as well as at several locations in Canada.
Constructed wetlands have the advantage of long-term, sustainable treatment
with very low costs of operation and maintenance. This is especially important for
leachate control, which is also often important to build projects with guaranteed
long-term stewardship. Passive constructed wetlands offer very long lifetimes, with
little or no equipment replacement.
In contrast with chemical and physical processing alternatives, wetlands
provide insurance against unanticipated new pollutions. For instance, conventional
air stripping can be used effectively to decrease the concentrations of NH4-N and
other volatiles. However, that technology has no capacity to deal with metals.
Wetlands have the capacity to deal with both. If in the lifetime of the leachate source
it becomes a source of metals, the wetland will have some capacity to treat this new
2.3 Heavy Metals
Heavy metals are common environmental pollutants that are produced as the
result of industrial, commercial and domestic activities. New pre-treatment
standards require some industrial discharges, such as electroplating and metal
finishing operations, to limit heavy metal levels to very low residual concentrations.
Conventional primary and secondary unit processes at municipal wastewater
treatment plants are inadequate for efficient removal of heavy metals. Advanced
processes including chemical precipitation, electrolysis, reverse osmosis, and ion
exchange are used for pre treatment of known sources of heavy metals of heavy
metals in industrial wastewater. Use of these processes to remove low
concentrations of heavy metal in wastewater has the disadvantage of high capital
cost, high operation and maintenance costs. Nevertheless, wetlands can accomplish
the same level of removal at lower labour and energy costs. The goal of treatment
for heavy metals is to remove the metals from the larger environment and from the
food chain, especially the food chain in river and ocean waters.
Chromium is a naturally occurring element found in rocks, animals, plants,
soil, and in volcanic dust and gases. Chromium is present in the environment in
several different forms. The most common forms are chromium (0), chromium (III),
and chromium (VI). No taste or odour is associated with chromium compounds.
Chromium (III) occurs naturally in the environment and is an essential nutrient while
chromium (VI) and chromium (0) are generally produced by industrial processes.
Chromium (0) is used for making steel while chromium (III) and chromium (VI) are
used for chrome plating, dyes and pigments, leather tanning, and wood preserving.
184.108.40.206 Effect of Chromium to Environment
Chromium enters the air, water, and soil mostly in the chromium (III) and
chromium (VI) forms. In air, chromium compounds are present mostly as fine dust
particles which eventually settle over land and water. Chromium can strongly attach
to soil and only small amount can dissolve in water and move deeper in the soil to
ground water. Fish do not accumulate much chromium in their bodies from water.
220.127.116.11 Effect of Chromium to Human Health
Chromium (III) is an essential nutrient that helps the body use sugar, protein,
and fat. However, breathing high levels of chromium (VI) can cause irritation to the
nose, such as runny nose, nosebleeds, and ulcers and holes in the nasal septum.
Ingesting large amounts of chromium (VI) can cause stomach upsets and ulcers,
convulsions, kidney and liver damage, and even death. Skin contact with certain
chromium (VI) compounds can cause skin ulcers. Some people are extremely
sensitive to chromium (VI) or chromium (III). Allergic reactions consisting of
severe redness and swelling of the skin have been noted.
Cadmium is a natural element in the earth's crust. It is usually found as a
mineral combined with other elements such as oxygen (cadmium oxide), chlorine
(cadmium chloride), or sulfur (cadmium sulfate, cadmium sulfide). All soils and
rocks, including coal and mineral fertilizers, contain some cadmium.
Cadmium is a component of some of the lowest melting alloys; it is used in
bearing alloys with low coefficients of friction and great resistance to fatigue; it is
used extensively in electroplating, which accounts for about 60% of its use. It is also
used in many types of solder, for standard E.M.F. cells, for Ni-Cd batteries, and as a
barrier to control nuclear fission. Cadmium compounds are used in black and white
television phosphors and in blue and green phosphors for color TV tubes. It forms a
number of salts, of which the sulfate is most common; the sulfide is used as a yellow
pigment. Cadmium and solutions of its compounds are toxic.
18.104.22.168 Effect of Chromium to Environment
Cadmium enters air from mining, industry, and burning coal and household
wastes. Cadmium particles in air can travel long distances before falling to the
ground or water. It enters water and soil from waste disposal and spills or leaks at
hazardous waste sites. It binds strongly to soil particles. Some cadmium dissolves in
water. It doesn't break down in the environment, but can change forms. Fish, plants,
and animals take up cadmium from the environment. Cadmium stays in the body a
very long time and can build up from many years of exposure to low levels.
22.214.171.124 Effect of Chromium to Human Health
Breathing high levels of cadmium severely damages the lungs and can cause
death. Eating food or drinking water with very high levels severely irritates the
stomach, leading to vomiting and diarrhoea. Long-term exposure to lower levels of
cadmium in air, food, or water leads to a buildup of cadmium in the kidneys and
possible kidney disease. Other long-term effects are lung damage and fragile bones.
Animals given cadmium in food or water had high blood pressure, iron-poor blood,
liver disease, and nerve or brain damage. We don't know if humans get any of these
diseases from eating or drinking cadmium. Skin contact with cadmium is not known
to cause health effects in humans or animals.
2.3.3 Metal Removal
Trace metals have a high affinity for adsorption and complexation with
organic material and are accumulated in a wetland ecosystem. The processes of
metal removal in wetlands are shown in Figure 2.4 (IWA Specialist Group, 2000).
Although some metals are required for plant and animal growth in trace quantities
(such as barium, beryllium, boron, chromium, cobalt, copper, iron, magnesium,
manganese, nickel, selenium, sulphur, molybdenum and zinc), these same metals can
be toxic at higher concentrations. Other metals such as arsenic, cadmium, lead,
mercury and silver have no known biological role, and can be toxic at even lower
Figure 2.4: Processes of metal removal in constructed wetlands
126.96.36.199 Adsorption and Cation Exchange
Adsorption involves the binding of particles or dissolved substances in
solution to sites on the plant or matrix surface. In a cation exchange reaction,
positively charged metal ions in solution bind negatively charged sites on the surface
of the adsorption material. The attractive force for cation exchange is electrostatic;
the size of this force depends on a wide range of factors. A cation in solution will
displace a cation bound to a site on the surface of a material if the electrostatic
attraction of the site for the dissolved cation exceeds that of the bound cation. The
cation exchange capacity (CEC) of a material is a measure of the number of binding
sites per mass or volume.
The CEC value has been shown to be the same whether the plant is alive or
dead. Wetlands sediments and soils also have large CEC values. The adsorption of
metals in the surface of soils is therefore a significant process in treatment wetlands.
The CEC of wetlands depends on the material of construction selected. Most gravel
or soil material tend to become saturated with metals in time (Howard et al. 1988).
188.8.131.52 Microbial Mediated Processes
The wetlands can be differentiated into two zones: aerobic and anaerobic.
The presence of metal oxidizing bacteria in the aerobic zones and sulphate reducing
bacteria in the anaerobic zones, which cause the precipitation of metal oxides and
sulphates respectively, has been established by Batal et al. (1987). For instance,
microbial mediated iron oxidation by Thiobacillus ferrooxidans, followed by the
subsequent precipitation of iron oxyhydroxide, is considered the most important iron
removal mechanism in wetlands treating metal rich mine wastewater. In unbalanced
Fe2+ + O2 + H2O → Fe(OH)3 + H+ (2.1)
Similar chemistries and limited investigation suggest similar oxidations for
many other metals including nickel, copper, lead, zinc, silver and gold. Wetlands
plants can potentially stimulate the growth of metal-oxidizing bacteria by oxygen
transfer into the rhizosphere.
Microbial mediated sulphate reduction consumes sulphate ions and produces
hydrogen sulphide and alkalinity in the form of bicarbonate ion. In balanced
equation form, where ‘CH2O’ represents a simple organic molecule:
SO4 + CH 2 O → H 2 S + HCO2 (2.2)
The H 2 S dissolves and ionizes to give sulphide ions, which react with a
range of metal ions to produce metal sulphide precipitates. Precipitation of metals as
sulphides rather than oxides has the following advantages: alkalinity produced by
sulphate reduction helps to neutralize acidity; sulphate precipitates are denser than
oxide precipitates; sulphides are precipitated within the origanic sediments andso are
less vulnerable to disruption by sudden surges in flow.
Vegetation can assist in metal removal by aiding the direct filtration of
particulate matter. Macrophyte species with high plant surface areas have been
shown to be very effective at retaining metal hydroxide particles that have
precipitated out of solution.
184.108.40.206 Plant Uptake
Some wetlands species have a well established ability for direct uptake of
heacy metals. Unfortunately, accumulation can become sufficient to kill the plant
within just one growing season. Fortunately, some species such as Typha latifolia
have a species-wide constitutional tolerance for heavy metals and do not accumulate
metals to toxic levels. The presence of an iron plaque in plant root system decreases
the uptake of metals by the root hairs (IWA Specialist Group, 2000).
It should be noted that direct uptake is an active process, requiring the plant
to be alive. Plant matter liberates its metal content on decomposing. Harvesting of
the foliage would only minimally assist metal in the above ground parts of the plants.
It is preferable to allow litter to form, as this can provide new sites for metal removal
and thermal insulation.
Nitrogen is a key element in wetlands biogeochemical cycles. Nitrogen
occurs in a number of different oxidation states in wastewaters and in treatment
wetlands. Numerous biological and physicochemical processes can transform
nitrogen between these different forms.
The IWA Specialist Group (2000) stated that major removal mechanism of
organic nitrogen in constructed wetlands is the sequential processes of
ammonification, nitrification and denitrification. Ammonia is oxidized to nitrate by
nitrifying bacteria in aerobic zones. Organic N is mineralized to ammonia by
hydrolysis and bacteria degradation. Nitrates are converted to nitrogen gas (N2) and
nitrous oxide (N2O) by denitrifying bacteria in anoxic and anaerobic zones. The
oxygen required for nitrification is suppled by diffusion from the atmosphere and
leakage from macrophyte roots. Nitrogen is also taken up by plants, incorporated
into the biomass and released back as organic nitrogen after decomposition. Other
removal mechanisms include volatilization and adsorption. On average, these
mechanisms are generally of less importance than nitrification – denitrification, but
they can be seasonally important.
220.127.116.11 Ammonia Volatilization
Ammonia volatilization is a physicochemical process in which NH4-N is
known to be in equilibrium between gaseous and hydroxyl forms as indicated below:
NH4 (aq) + H2O → NH 4 + OH − (2.3)
Losses of NH3 through volatilization from flooded soils and sediments are
insignificant if the pH is below 7.5 and very often losses are not serious if the pH is
below 8.0. At pH of 9.3 the ratio of ammonia to ammonium ion is 1:1, and the losses
via volatilization are significant . Algal photosynthesis in constructed wetlands as
well as photosynthesis by free-floating and submerged macrophytes often creates
high pH values (IWA Specialist Group,2000).
In a broad literature review, Vymazal (1995) summarized that volatilization
rate is controlled by the NH 4 concentration in water, temperature, wind velocity,
solar radiation, the number of aquatic plants, and capacity of the system to change
the pH in diurnal cycles.
18.104.22.168 Ammonification (Mineralization)
Ammonification (mineralization) is the process in which Org-N is converted
into inorganic N, especially NH4-N. Mineralization rates are fastest in the
oxygenated zone and decrease as mineralization switches from aerobic to facultative
anaerobic and obligate anaerobic microflora. The rate of ammonification in wetlands
is dependent on temperature, pH, the C:N ratio of the residue, available nutrients in
the system, and soil conditions (Reddy and Patrick, 1984).
22.214.171.124 Nitrification / Denitrification
Nitrification is usually defined as the biological oxidation of ammonium to
nitrate with nitrite as an intermediate in the reaction sequence. Nitrifiaction is a
chemoautotrophic process. The nitrifying bacteria derive energy from the oxidation
of ammonia and nitrate, and carbon dioxide is used as a carbon source for the
synthesis of new cells. These organisms require O2 during NH4-N and nitrite-N
oxidation to nitrate-N. Nitrification process occurs in two steps, which is illustrated
by the following equation (Hauck, 1984)
NH 4 + 1.5O2 → NO2 + 2 H + + H 2 O
NO2 + 0.5O2 → NO3
NH 4 + 2O2 → NO3− + 2 H + + H 2 O (2.4)
The first step, the oxidation of ammonium to nitrite, is executed by strictly
chemolithotrophic (aerobic) bacteria, which are entirely dependent on the oxidation
of ammonia for the generation of energy for growth. While during the second step in
the process of nitrification, the oxidation of nitrite to nitrate, is performed by
facultative chemolithotrophic bacteria, which can also use organic compounds in
addition to nitrite for the generation of energy for growth. Vymazal (1995)
summarized that nitrification is influenced by temperature, pH, alkalinity, inorganic
C source, the microbial population and concentrations of NH4-N and dissolved
oxygen. The optimum temperature for nitrification is pure cultures ranges from 25 to
35 °C and in soils from 30 to 40 °C.
The first anoxic oxidation process to occur after oxygen depletion is the
reduction of nitrate to molecular nitrogen or nitrogen gases, which is called
denitrification. Denitrification is a bacteria process in which nitrogen oxide serve as
terminal electron acceptors for respiratory electron donating substrate through
several carrier system to a more oxidized N form. The resultant free energy is
conserved in ATP, after phosphorylation, and is used by denitrifying organisms to
support respiration. Denitrification is illustrated by the following equation (Hauck,
6(CH 2 O) + 4 NO3− → 6CO2 + 2 N 2 + 6 H 2 O (2.5)
This reaction is irreversible and occurs in the presence of available organic
substrate only under anaerobic or anoxic conditions, in which nitrogen is used as an
electron acceptor in place of oxygen. Most denitrifying bacteria are
chemoheterotrophs, obtaining energy solely through chemical reactions and use
organic compounds as electron donors and as a source of cellular carbon (Hauck,
1984). Vymazal (1995) summarizes the environment factors known to influence
denitrification rates, including the absence of O2, redox potential, soil moisture,
temperature, pH, the present of nitrifiers, soil type, organic matter and present of
Nitification and denitrification are known to occur simultaneously in flooded
soils in which both aerobic and anaerobic zones exist, such as would occur in a
flooded soil or water bottom containing an aerobic surface layer over an anaerobic
layer, or in the aerobic rhizosphere microsites in otherwise anaerobic soi.
126.96.36.199 Plant Uptake
Nitrogen removal by plant uptake can only be accomplished if the plants are
harvested. Nitrate, produced through nitrification, is removed by denitrification and
plant uptake. In the past, there has been some question as to whether nitrification or
plant uptake is the principal ammonia-nitrogen conversion mechanism that ultimately
leads to nitrogen removal. Based on data collected for a review of existing water
hyacinth treatment systems, it has been concluded that nitrification followed by
denitrification was the principal nitrogen removal mechanism. Only when water
hyacinth systems received low nitrogen loadings and significant harvesting was
conducted did plant uptake become the principal nitrogen removal mechanism (IWA
Specialist Group, 2000).
Constructed and natural wetlands are capable of absorbing new phosphorus
(P) loadings and in appropriate circumstances can provide a low-cost alternative to
chemical and biological treatment. Phosphorus interacts strongly with wetlands soils
and biota, which provide both short term and sustainable long term storage of this
Phosphorus removal from aquatic macrophyte system is due to plant uptake,
microbial immobilization into detritus plant tissue, retention by the underlying
sediments, and precipitation in the water column. Since P is retained by the system,
the ultimate removal from the system is achieved by harvesting the plants and
dredging the sediments.
In SSF wetlands, the sorption capacity of the media can be designed to
provide significant phosphorus removal (Maehlum et al 1995). This storage
eventually becomes saturated, necessitating the replacement of the medium and the
establishment of the wetlands.
Settle able organics are rapidly removed in wetlands system under quiescent
conditions by deposition and filtration. Attached and suspended microbial growth is
responsible for the removal of soluble organic compounds, which are degraded
aerobically as well as anaerobically. The oxygen required for aerobic degradation is
supplied directly from the atmosphere by diffusion or oxygen leakage from the
macrophyte roots into the rhizosphere. The uptake of organic matter by the
macrophytes is negligible compared with biological degradation (Watson et al. 1989;
Cooper et al. 1996).
Basic to understanding of any biological treatment mechanism is an
understanding of the micro organisms undertaking the treatment. To continue to
reproduce and function properly, an organism must have a source of energy, carbon
for the synthesis of new cellular material, and inorganic elements (nutrients) such as
nitrogen, phosphorus, sulphur, potassium, calcium and magnesium. Some organic
nutrient can also be required. Often industrial effluents require the addition of
nutrients such as phosphorus or nitrogen for effective biological treatment. The two
main sources of cell carbon are organic chemicals and carbon dioxide. Organisms
that use organic carbon for the formation of cell tissue are called heterotrophs.
Organisms that derive cell carbon from carbon dioxide are called autotrophs. Both
groups use light or a chemical oxidation-reduction reaction as an energy source for
If the major objective of treatment is a decrease in organic content
(carbonaceous BOD), the heterotrophic organisms are of primary importance
because of their requirement for organic material as a carbon source and their higher
2.6 Role of Wetlands Vegetation in Wastewater Treatment
The macrophytes growing in constructed wetlands have several properties in
relation to the treatment processes that make them an essential component of the
design. The most important effects of macrophytes in relation to the wastewater
treatment processes are the physical effects that the plant tissues give rise to (such as
erosion control, filtration effect and the provision of surface area for attached micro
organisms). The metabolism of the macrophytes ( such as plant uptake and oxygen
release) affects the treatment processes to different extents depending on design. The
macrophytes have other site – specific valuable function, such as providing a suitable
habitat for wildlife and giving systems an aesthetic appearance. The major roles of
macrophytes in constructed wetlands are summarized in Table 2.2
Table 2.2 Roles of macrophytes in constructed wetlands
Macrophyte property Role in treatment process
Aerial plant tissue Light attenuation: reduce growth of phytoplankton
Reduce wind velocity: reduces risk of resuspension
Aesthetically pleasing appearance of system
Storage of nutrients
Plant tissue in water Filtering effect: filter out large debris
Reduce current velocity: increases rate of sedimentation,
reduces risk of resuspension
Provide surface area for attached biofilms
Excretion of photosynthetic oxygen: increases aerobic
Uptake of nutrients
Roots and rhizomes Provide surface for attached bacteria and other micro
in the sediment organisms
Stabilizing the sediment surface: less erosion
Prevents the medium from clogging
Release of oxygen increase degradation and nitrification
Uptake of nutrients
Release of antibiotics
2.7 Rainfall in Malaysia
Most of Malaysia (Peninsular & Borneo) lies between the equator and the
northern 7° latitude. So for the most part, we experience a predominantly equatorial
climate (i.e., it's hot and humid for most of the year).
Over the east coast districts (Kelantan, Terengganu, Johor, Pahang),
November, December and January are the months with maximum rainfall, while
June and July are, generally, the driest months. Most island resorts on the east coast
during the wet, monsoonal, periods would have been closed tourist.
Over the rest of the Peninsula, with the exception of the Southwest coastal
area (Johor), the monthly rainfall pattern shows two periods of maximum rainfall
separated by two periods of minimum rainfall (i.e., between the two monsoons, we
experience low periods of rain, sometimes even periods of drought). The first wet
season occurs in October-November while April-May are also wet months. Over the
north-western region (Perlis, Kedah), the driest months occur in January-February,
furthermore, June-July are also dry (inter-monsoon) months. Elsewhere in Peninsular
Malaysia the driest months usually occur in June-July with the inter-monsoon
months seeing drier weather in February.
The rainfall pattern over the Southwest coastal area (Johor or Malacca) does
not conform to the above patterns as these states are affected by early morning
"Sumatra-rains" between May to August. In these states, October and November are
the wettest months and February the driest.
The coastal areas of Sarawak and Northeast Sabah experience similar rainfall
patterns, but the periods of drought between these two regions differ somewhat.
While the wettest month occurs during January in both areas, the drier months vary
between these regions. In the coastal areas of Sarawak, the driest months are between
June and July while in the Northeast coastal areas of Sabah, this is in April. The
wettest months occur during the Northeast monsoon usually between December and
Inland areas of Sarawak generally experience quite evenly distributed annual
rainfall. Nevertheless, slightly less rainfall is received during the period June to
August, which corresponds to the prevailing south-westerly winds. It must be pointed
out that the highest annual rainfall area in Malaysia may well be found on the hill
slopes of inland Sarawak. Long Akah, by virtue of its location, receives a mean
annual rainfall of more than 5000 mm!
The Northwest coast of Sabah experiences a rainfall regime of which two
maxima and two minima can be distinctly identified. The wettest months occur
around October and a somewhat wet month also occurs in June. The dry months
occur in February followed by August. The inter-monsoonal "drought" van be severe
in this part of Sabah especially during el nine years.
In the central parts of Sabah where the land is hilly and sheltered by mountain
ranges, the rainfall received is relatively lower than other regions and is evenly
distributed. In general, the driest months occur in February and August while the
wettest months are in May and October. Southern Sabah has evenly distributed
rainfall. The annual rainfall total received is comparable to the central part of Sabah.