THE FATE OF PHOSPHORUS IN WETLANDS
A REVIEW
Report No. 96/15
For the Queensland Department of Natural Resources
THE FATE OF PHOSPHORUS IN WETLANDS
A REVIEW
ACTFR Report No. 96/15
Prepared by J. W. Faithful of the
Australian Centre for Tropical Freshwater Research
James Cook University of North Queensland, Townsville Q 4811
The Fate of Phosphorus in Wetlands - A Review, ACTFR Report No 96/15
TABLE OF CONTENTS
1.
INTRODUCTION ...............................................................................................................................1
2.
SCOPE OF THE REVIEW............................................................................................................1
3.
WETLAND SYSTEMS.................................................................................................................2
4.
TYPES OF ARTIFICIAL WETLANDS .......................................................................................3
5.
MECHANISMS FOR PHOSPHORUS REMOVAL IN WETLANDS ........................................6
6.
REVIEW OF LITERATURE RELATED TO PHOSPHORUS RETENTION BY
NATURAL AND CONSTRUCTED WETLANDS ....................................................................13
6.1
Australian Research ...................................................................................................13
6.2
International Research ...............................................................................................20
6.3
Underlying Criticisms of the Reviewed Journal Articles and Proceedings Papers ...27
7.
CONSTRUCTED WETLAND CONSIDERATIONS FOR THE REMOVAL OF
PHOSPHORUS............................................................................................................................28
8.
MANAGEMENT RECOMMENDATIONS ...............................................................................34
9.
CONCLUSIONS..........................................................................................................................36
10.
REFERENCES ............................................................................................................................38
APPENDIX
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1.
INTRODUCTION
Natural and artificial (created or constructed) wetland systems are economical and aesthetic alternatives
to water treatment technology for the control of nutrients and pollutants in waters. Over the past twenty
years they have been used effectively to decrease the concentration of various pollutants from wastewater,
stormwater, mining discharge and agricultural runoff, particularly in Europe and North America.
There is a myriad of literature which reports the various functions of wetlands and the effectiveness of
certain wetland designs in removing and treating pollutants. Much of the information has been focused
on temperate wetlands, i.e. in North America and Europe, but research is growing on a wider international
basis. Australian research is increasing in this area and several key groups are gaining credibility within
this widening scientific field. The Cooperative Research Centre for Freshwater Ecology (Albury and Mt.
Waverley), CSIRO Division of Water Resources (Griffith) and the University of Western Sydney are
currently making effective inroads into understanding the use and design of wetlands in Australia.
Despite the numerous articles published on wetlands in recent years there is a notable gap in the literature
regarding research on natural and constructed tropical wetlands and their efficiency in reducing
pollutants. It is difficult to apply temperate wetland data to wetlands in the tropics because of large
biological and chemical differences due to warmer climates (such as increased plant productivity,
increased potential for plant diversity, soil matrix activity, etc.). It has been suggested by Barbier (1994)
that tropical wetlands have a high economic value and a crucial role to play in development, and the
current lack of scientific data on their ecological relationships and functions is a cause for concern,
particularly in developing countries.
The issue of wetland efficiency in pollutant removal is further confounded by the interpretation of some
data presented in books and journal articles, as well as the inconsistency of units used to express loading
rates, plant uptake rates, etc. Much of the data has been presented on an inflow/outflow basis without
consideration to the processes involved within the wetlands (chemical, physical and biological), with the
assumption that wetlands act as a ‘black box reactor system’ capable of removing all manner of pollutants
effectively. In recent years there has been a push to ensure that scientists and engineers, at all levels, take
these factors into consideration when deciding upon the use of wetlands for any wastewater treatment
application. A major source of the recent literature presented in this review has been from international
conferences which have brought scientists and engineers together in an attempt to determine the best
criteria for wetland design.
2.
SCOPE OF THE REVIEW
This review examines the fate of phosphorus (P) in natural and constructed (artificial) wetlands by
focussing on recent Australian and international research. There has been much debate regarding the long
term success of wetlands in removing P to any extent, and the processes by which the P is removed.
In general, the issues that are addressed in this review are :
• aspects of P mobility in both natural and constructed wetlands;
• the success of constructed wetlands in P removal;
• design criteria for the removal of P in artificial wetlands, such as plant species
composition, planting patterns, retention time, wetland depths, type of wetland (open
surface or gravel bed), effect of wetting and drying cycles on P removal, wetland shape,
substrate types, necessity for harvesting and other parameters;
• management recommendations for wetlands specific to P removal; and
• comments on the appropriateness of constructed wetlands for P removal especially in the
long term.
This review examines the importance of all mechanisms occurring in natural and constructed wetlands
that effect P removal, including plant uptake, assimilation by micro-organisms and algae, attachment to
sediments and soil particles, and precipitation. Special emphasis is given to long term removal systems.
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The success of artificial wetlands for P removal is also examined with primary attention given to raw
effluent quality, climatic conditions, age of the wetland and design criteria. The level of monitoring
carried out with respect to specific studies is also addressed.
3.
WETLAND SYSTEMS
The most productive ecosystems in the world are those dominated by aquatic macrophytes (Brix 1993),
and these may be divided into separate groups according to domination by free floating, submerged or
rooted emergent species. Their high productivity results from having high availability of light, nutrients,
and water, and from the plant’s morphological and physiological ability to take advantage of this
environment. High levels of activity also occur at the microbial level resulting in the decomposition of
organic matter and other substances. For these reasons, aquatic ecosystems (in particular wetlands) have
been considered as alternatives and/or supplements to a variety of water treatment and recycling processes
(Bavor et al. 1995; Kadlec 1995; Wood 1995; Brix 1994b; Cullen 1989).
The conditions generated by an environment which is constantly saturated or covered by water limit
gaseous exchange between the air and the sediment and this may result in anoxic sediment. As a
consequence decomposition and mineralisation rates are significantly reduced, prompting the
accumulation of organic matter on the sediment surface. This mass of organic litter, and the macrophytic
vegetation, provides a significant surface area for microbial growth and therefore provides the impetus for
large amounts of organic matter and nutrient transformation. The water/sediment interaction and
associated microbial activity is the driving force behind water purification processes and therefore a sink
for nutrients in both constructed and natural wetlands. Another significant factor which determines the
effectiveness of a wetland as a water treatment system is the amount of time that the water stays in contact
with the wetland, and this is related to the size of the wetland and the amount of water it receives.
Natural wetlands are capable of significantly improving the quality of water flowing through them but the
extent to which this occurs depends on the predominant water treatment process(es) acting within the
wetland, such as macrophyte assimilation and substrate adsorption. There is great difficulty predicting
the effectiveness of one wetland area compared with another. Natural wetlands are of high conservation
value as they are very susceptible to changes in structure (in terms of species composition) and
effectiveness with prolonged exposure to pollutants. There are, of course, exceptions to the rule, and
several studies have shown that natural wetlands can function effectively in reducing nutrient loading
from effluent discharge in various areas over a long period of time.
Constructed wetlands have been used in various parts of the world for the past twenty years with little
consistency in design criteria. Data obtained for these wetlands have generally been variable, but on the
whole indicate that gross pollutant indicators such as biochemical oxygen demand (BOD), suspended
solids and bacterial matter, can be effectively removed from the inflowing water. These constructed
wetlands have been used to ‘polish’ treated outflow from primary and secondary treatment plants. Their
design presents the water inflow with a defined flow pattern over a specific substrate and vegetation type
which, as opposed to natural wetlands, is site selected and sized specifically for a controlled hydraulic
pathway and retention time. Some advantages of constructed wetlands are their low cost of construction
and maintenance when compared to the costs of treatment plants, their low requirement for energy, their
flexibility and the fact that they are low-technology based systems. The disadvantages include the
requirement for large amounts of land, depending on their use, and seasonal variability in their
effectiveness.
The most common wetland system used in water treatment is the macrophyte-based system where a
wetland is constructed with one or several shallow ponds, and one or more species of aquatic macrophyte.
The water inflow is generally regulated surface flow or sub-surface flow. The pollutants are removed
from the inflowing water by a combination of processes (chemical, physical and biological) within the
wetland, such as sedimentation, precipitation, adsorption to soil particles, assimilation by plant tissue and
microbial transformations. Macrophytes can enhance pollutant removal within the system by either
assimilating them directly or by providing an environment for surface microbial attachment to transform
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and uptake pollutants. The rhizosphere of aquatic plants is also a primary site for pollutant uptake and
transformation as it is a zone of oxygen transfer between the plant and sediment which is a requisite for
sediment microbial activity and pollutant oxidation (Brix 1994a).
Over the past few years specific designs of constructed wetlands have resulted from data generated by
many experimental treatment facilities. Factors such as horizontal surface or subsurface flow, vertical
flow gradient and retention time have all been considered depending on the use of the wetland. These will
be discussed in further detail at a later stage in the review. Pollutants such as settleable and suspended
solids, BOD, metals and pathogens are generally removed by sedimentation and filtration processes, some
of which can be provided by primary mechanisms such as gross pollutant traps situated before the
wetland inflow. Soluble organic matter and nutrients (total and dissolved inorganic) are generally
removed by bacterial oxidation and/or anaerobic degradation. The major removal mechanism of
nitrogenous material is nitrification and denitrification, with some direct plant assimilation, and most
constructed wetlands seem to be fairly efficient in providing mechanisms for these treatment processes.
The mechanisms controlling the fate of the aforementioned pollutants are well understood, but P removal
in artificial wetlands is not. The major processes that govern the removal of P in wetland systems are
plant assimilation and substrate adsorption and complexation. Additionally, precipitation reactions can
occur under certain conditions when the inflowing water comes into contact with available aluminium,
iron, calcium and other clay minerals in the sediment (Howard-Williams 1985). Sediment fixation of P
can also occur, and is enhanced by alternating wet and dry conditions. Plant uptake may be another
important process and this is significant when specific loading and the soil P concentration is low. There
have also been reports that phosphate reduction to gaseous hydrogen phosphides may occur under
anaerobic conditions.
4.
TYPES OF CONSTRUCTED WETLANDS
Aquatic Macrophyte Wastewater Treatment Systems
The dominant form of macrophyte within the wetland classifies the wetland treatment system. There are
free-floating macrophyte systems, rooted emergent macrophytic systems, submerged macrophyte systems
and multi-stage systems, which are a combination of the preceding systems, and other kinds of low
technology systems (oxidation ponds, sand filters, etc.).
The free-floating macrophyte systems are generally limited to systems based on water hyacinth
(Eichhornia crassipes) or duckweed (Lemna, Spirodella or Wolffia spp.) in which the plants are allowed
to grow in shallow ponds or cells within treatment ponds. Much has been documented on the
productivity and effectiveness of these systems, particularly the water hyacinth (Karpiscak et al. 1994,
Aoyama and Nishizaki 1993, Reddy and Sutton 1984). Hyacinth is more effective in warmer climates
making it an ideal system for tropical environments but, as it is a declared noxious weed in Australia, it
cannot be used. Lemna spp. and other duckweeds are more manageable; however, they are a minute
plant, and therefore require barrier systems or compartments to minimise movement on the water by wind
to ensure water surface coverage. Duckweed has a wider geographical coverage and climatic tolerance
and, being a smaller plant without an extensive root system, is mainly used to assimilate nutrients and
provide anaerobic light reduced water columns for denitrification and precipitation. The design of a
Lemna based treatment system (The Lemna System) has been patented in the United States and has been
effective in the removal of BOD, suspended solids and total nutrients (Poole and Ngo 1993).
Emergent aquatic macrophytes are the most dominant form of aquatic plant in natural wetlands and marsh
systems. In general they grow within a water table ranging from 50cm below the soil surface to a water
depth of 150cm or more. They produce aerial stems and leaves and possess an extensive root and
rhizome system. The depth penetration and mass of the root system, and hence their exploitation of
sediment, differs amongst macrophytic species. In constructed wetlands, the most commonly used rooted
emergent macrophytes are the common reed (Phragmites australis), cattail or cumbungi (Typha latifolia,
T. orientalis or T. domengensis) and bulrush (Schoenoplectus validus or S. mucronatus). All species are
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adapted to growing in water-logged anoxic sediments as they have large internal air spaces for the
transportation of oxygen to the roots and rhizomes. The extensive lacunal systems possessed by these
plants occupy a large percentage of the overall plant. Oxygen is transported to the roots and surrounding
rhizomes by either direct diffusion and/or by convective air flow. Air diffusion by the roots and rhizomes
to the rhizosphere creates an envelope of aerated soil in the otherwise anoxic soil which stimulates
organic matter decomposition and the growth of nitrifying bacteria (Brix 1994a). There are three major
types of treatment systems based on emergent macrophytes as described by Brix (1993) :
1.
Surface flow systems
These systems represent the oldest type of artificial wetland design. They can vary from open
wetlands, such as the very large constructed wetlands sometimes utilised in the U.S. (thousands
of hectares), to short constructed channels. The channel system is typically no more than 100m
long, 3-5m wide and planted with a species of bulrush. The water treatment processes are
favoured by the presence of submerged portions of stems and litter which serve as a suitable
substrate for attached microbial growth. The bottom of the channels are generally sealed to
prevent wastewater leakage.
2.
Horizontal subsurface flow
This design was pioneered in Germany by Seidel in the 1950s and developed further in the 1970s
(Brix 1994b). The system typically consists of a bed planted with common reed and sealed by an
impermeable membrane to prevent flow loss. The medium in the channel is soil and/or gravel of
varying size fractions so that during the passage of flow through the system, water contact with
the plant rhizosphere is promoted and organic matter can be decomposed microbiologically,
nitrogenous compounds denitrified and P and metals fixed in the soil. Two important functions
occur through this system as a result of the reeds :
• oxygen is supplied to the heterotrophic organisms in the rhizosphere, and
• hydraulic flow through the medium is increased and stabilised.
As a result of the subsurface contact and adsorption activity, the amount of nutrient assimilated
by the plants represents a small portion of the overall content in the initial wastewater flow.
Recycling of the nutrients bound in the plant tissue occurs upon the senescence and decay of the
plant. BOD and suspended solids are removed effectively via these systems but the removal of
nitrogen and P varies greatly depending on the loading rate of the wastewater, type of substrate,
and the type and composition of the wastewater. The flow rate is an important factor as high
input resulting in surface flow has to be avoided as this prevents the wastewater coming into
contact with the sediment and the rhizosphere.
3.
Vertical subsurface flow
Vertical subsurface flow systems allow for improved hydraulic conditions and water/rhizome
contact. This design provides percolation flow with intermittent loading which improves soil
oxygenation when compared to horizontal flow systems. During the loading period, air is forced
out of the soil and during the percolation phase the surface soil dries out drawing air back into the
soil pore spaces. This process therefore provides alternating oxidising/reducing conditions in the
soil promoting alternating nitrification and denitrification reactions and P adsorption. Vertical
flow, and more significantly, vertical upflow systems are currently being developed by the CRC
for Freshwater Ecology in Albury and preliminary findings appear to indicate that these systems
are promising as single-use, low load systems such as household treatment systems, particularly
for P removal (Breen and Chick 1995; Chick and Mitchell 1995; Mitchell et al. 1995; Heritage et
al. 1995).
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The submerged macrophyte system uses plants which have their photosynthetic tissue entirely
submerged. The diversity of plants available for use is great and includes low-productivity oligotrophic
water species (Lobelia dortmanna), commonly occurring species in freshwater systems (Potamogeton
spp., Ceratophyllum spp. and Myriophyllum spp) and high productivity species that thrive in eutrophic
waters (Elodea, Elodea canadensis; Hydrilla, Hydrilla verticillata). These plants can assimilate nutrients
directly from the water but only grow well in oxygenated waters. Therefore, these systems are not
suitable for receiving wastewater with a high loading of organic matter. Their primary use could be to
polish treated water, whether derived from a secondary treatment system or a low pollutant effluent
source. High photosynthetic activity in these systems accompanied by good light penetration and warm
temperature depletes dissolved inorganic carbon in the water resulting in pH values greater than 7 and,
therefore, ideal conditions for P precipitation and the volatilisation of ammonia. The associated high
concentration of dissolved oxygen will also favour the mineralisation of organic matter. The nutrients
assimilated are generally thought to be translocated to, and retained within, the rooting tissues and
microflora attached to the plant. Senescence and decay of the plant means that the plant nutrient and
detrital matter rarely leaves the littoral detritus and macrophyte-epiphytic complexes. In some species of
macrophytes the translocation of nutrients to the below ground biomass is not efficient so that senescing
leaves or fronds have a substantial nutrient composition which may become available to the water column
upon decomposition.
Multi-stage systems are the most favourable designs to consider effective removal of pollutants. They
allow for a wider range of pollutant removal, as well as allowing for the effective removal of fractions of
the same contaminant, i.e. dissolved inorganic, particulate, and organic forms. Multi-stage systems permit
the user to work from the process point of view and so dictate the required removal mechanisms within
the wetland design. An example could consist of :
• a mechanical clarification step for primary treatment to remove coarse sediments, oil
and grease, etc., in the form of a gross pollutant trap;
• a floating or emergent macrophyte system for secondary treatment which incorporates
subsurface flow; and
• a floating, emergent or submergent macrophyte system for tertiary treatment which
can also incorporate subsurface flow.
The multi-stage system, therefore, attempts to mimic the conventional wastewater treatment systems by
separating the treatment processes to optimise the performance of the wetlands in relation to the needs of
the treatment system (Brix 1993). The system design will of course, be based on factors such as the
wastewater characteristics, the treatment requirements, the climate and the amount of available land.
Currently the design of artificial wetlands is based on the premise, or expectation, that the system will
have a simultaneous pollutant removal capacity, i.e. aerobic BOD degradation, microbial denitrification
and nitrification, and P fixation occurring within the same ‘reactor’. Understanding of the key processes
within wetlands has been qualitatively documented to date but the quantitative data on the rates of these
processes and the factors that influence them are poorly understood or ignored. Recent studies have
attempted to improve understanding of the wetland processes (Mitsch et al. 1995, Reddy et al. 1995,
Johengen and LaRock 1993, Mitsch and Reeder 1991, Howard-Williams 1985). One cannot assume that
an outcome of efficient pollutant removal can be successfully achieved in all cases. So, where wetland
systems are proven effective in removing certain pollutants they have to be examined and their processes
fully understood so that further wetland systems can be developed on the successful design criteria.
5.
MECHANISMS FOR PHOSPHORUS REMOVAL IN WETLANDS
There are several processes for the removal of P from wastewater in a wetland environment, but some
have only a limited capacity for removal, and once the P assimilating/adsorbing capacity is reached or
exceeded no further removal will be able to occur. In general, the ultimate pathway of P reduction is its
movement from the water column to the sediment, uptake by macrophytes (from the water column or the
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substrate), algae and epiphytes and finally incorporation by micro-organisms associated with the litter. It
does appear from work reported by Kadlec (1994) that accumulative P removal processes essentially
follow first order reactions, which means that the removal of P to new soils is proportional to the
concentration of P in the surface waters, and to the surface area of the wetland. Time should be allowed
for wetland equilibrium to be achieved and this will result in stable water quality performance. In
Kadlec’s case this was achieved after two years although the plant species composition and other biotic
components continued to adapt over a longer period of time.
Forms of Phosphorus within Wetlands Systems
When determining the role of P retention by wetlands, particularly wetland substrates, it is important to
understand the forms of P in the system. In wastewater entering a wetland, particularly secondarily
treated effluent from a sewage treatment plant (STP), the P component will be composed predominantly
of filterable reactive P (FRP) - which will consist of a mixture of orthophosphates and a proportion of
labile condensed phosphates (McKelvie et al. 1993). The total P content comprises both inorganic and
organic particulate and filterable non-reactive P forms, which will include a significant proportion of
colloidal-P complexes and other P containing compounds, such as polyphosphates, metaphosphates and
dissolved organic P (inositols, phospholipids, nucleic acids, phosphoamides, phosphoproteins, sugar
phosphates, aminophosphonic acids, organophosphorus pesticides and organic condensed phosphates).
The P composition of other sources of wastewater, such as urban stormwater, agricultural runoff and
mining leachate, will differ and will be dominated by a particular fraction(s).
The forms found within the substrate, particularly after water-substrate contact, are often delineated into
inorganic and organic pools of P. The major pools of inorganic P (Pi) are defined as loosely adsorbed P,
iron and aluminium P, and calcium and magnesium bound P. These forms are not generally discrete
entities as transformations between the forms occur continuously to maintain equilibrium conditions
(Sharpely 1995). The inorganic forms are dominated by hydrous sesquioxides, amorphous and crystalline
aluminium, and iron compounds in acidic, noncalcareous soils and by calcium compounds in alkaline
calcareous substrates.
The loosely adsorbed P is important for plant growth and controlling the P concentration of the overlying
water column (Reddy et al. 1995). This fraction responds to external P loadings and, as expected, is a
proportion of both constructed and natural wetlands.
The P associated with oxyhydroxides is readily desorbed under most conditions, but the P associated with
crystalline iron and aluminium is desorbed only under extended waterlogged conditions (i.e. anoxic
conditions). Redox conditions, however, have no effect on the sorption of P by aluminium, only on the
sorption by iron (moody pers. comm.). In wetlands that have high levels of iron and aluminium in the
catchment watershed, this fraction of P is generally the dominant pool of Pi.
The calcium and magnesium forms of P are generally unavailable to biological assimilation under natural
conditions and are not the predominant form of Pi under low pH conditions (i.e. those commonly
associated with marshlands or high organic soils). However, under anoxic conditions, the sediment pH
will most likely be neutral to alkaline and calcium and magnesium forms of P may be the dominant form.
The organic P fraction (Po) primarily consists of the forms of P associated with phospholipids, inositols
and fulvic acids, and forms of humic acids. This form of P is generally biologically reactive and can be
hydrolysed to bioavailable forms. Organic P can be mineralised by alternate wetting and drying cycles,
changes in substrate pH and increased microbial activity. Therefore, the bioavailability and mobility of P
in wetland substrates under aerobic conditions is greater than that found in dry aerobic soils.
The residual forms of P not able to be extracted by standard chemical extraction processes are considered
to be highly resistant and therefore biologically unavailable.
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Plant Assimilation
The capacity of aquatic macrophytes to assimilate P is dependent on their growth rates, the ionic
composition of the water, the water depth, the sediment characteristics, the oxygen transfer capability of
the plants into the root-zone, biochemical and physico-chemical processes functioning at the root-watersediment interface, plant density per unit area, plant harvesting and climate (Reddy et al. 1995). The
potential rate of P uptake is limited by the plants’ net productivity and the concentration of nutrients in
the plant tissue. Phosphorus storage is similarly dependent on plant tissue concentrations, and also on the
ultimate potential for biomass accumulation (i.e. maximum standing crop). In addition to the plant
assimilation of P, its removal in wetlands containing macrophytes is also affected by a number of
biological, physical and chemical processes functioning in the water, sediment and rhizosphere.
As aquatic plants grow, there is an uptake of P into the plant cells which will continue until after the plant
is fully grown. At the end of the growing season, aquatic plants such as reeds die back and the leaves and
stalks eventually fall to the bed where they break down. If the bulk of nutrients have not been
translocated to the roots and rhizomes as is the case with some macrophytes, P will eventually return back
to the system. New growth of these macrophytes will result in the uptake of P again so that an
equilibrium will eventually develop where the P take up by plant growth in a year will equal the P return
by dead plant breakdown. Therefore, if harvesting is not planned within wetlands vegetated by these
particular macrophytes or the removal of their senesced leaves, the plants will bring about no net P
removal.
Hocking (1989) reported that, in general, the take-up of P into the plant cells is relatively small. The P
content for plants such as reeds (P. australis) ranged from 0.9 to 1.35 mg g-1 (dry weight) for stems, 1.0 to
1.7 for leaves, and 0.9 to 1.63 for whole shoots. There is, however, no data on the assimilation rates into
the roots and rhizomes of the plants. Davies and Cottingham (1993) calculated, for example, that if a
yield of 95 tons dry weight per year per hectare of a very vigorous stand of P. australis up to 5 m tall was
established (using the highest P content of 1.7 mg g-1 provided by Hocking, which excludes root biomass
assimilation rates) a yearly removal of less than 6% P by harvesting the reeds would be expected from a
typical artificial wetland with a loading of primary settled sewage with a P content of approximately 8mg
P L-1 delivered at a rate of 96 L m-2 d-1. Hence the use of macrophytes to remove P from wastewaters,
even under optimal conditions, can be severely limited if they are expected to provide the main P
removing process. The main advantage of emergent macrophytes, however, is that they provide a greater
surface area for epiphytic colonisation, and an additional pathway for nutrient assimilation, and that they
produce litter stimulating microbial nutrient uptake.
Data are available for the herbaceous macrophytes that are generally used in constructed wetlands to aid
developers to select which plants are capable of storing high levels of nutrients. The general above
ground P concentration (in vegetation and leaves) in temperate environments is of the order 0.1 to 0.3%
(Johnston 1991). There is a wealth of literature available for temperate regions and the following plants
have been rated highly for P content for leaves and herbs : Alternantha philoxeroides, 2.7-5.3 g m-2 (P
content) and 0.39% P (P concentration at peak standing crop); Glyceria grandis, 5.2-6.8 g m-2 and 0.130.21% P; Phragmites communis, 2.0-5.3 g m-2 and 0.18% P; Sagittaria lancifolia, 3.58 g m-2 and 0.050.58% P; Typha glauca 3.17-3.74 g m-2 and 0.53% P, Typha latifolia, 0.7-3.2 and 0.1-0.4% P; Eichhornia
crassipes, 0.5-18.0 g m-2 and 0.14-0.8% P; and, Lemna minor, 0.1-3.3 g m-2 and 0.75% P. These values
underestimate the total P content within the plant because the root biomass and leaf litter are not taken
into consideration. It is important to note that some of the plants which are highly rated for phosphorus
removal are classified as noxious weeds in Australia (i.e. A. philoxeroides and E. crassipes) and cannot be
considered for use in constructed wetlands. Finlayson et al. (1984) provided some information on the
biomass and P assimilation potential of several macrophytes in a tropical lake in Mount Isa, north-west
Queensland. In the winter month of July, the biomass of Potamogeton crispus was estimated at ~6 kg dry
weight m-2, Hydrilla verticillata, ~3 kg dry weight m-2 and Salvinia molesta 0.81 kg dry weight m-2 and
although P was determined as a limiting nutrient, the annual average P content was 0.07 to 0.47% P as
dry weight for S. molesta, 0.09 to 0.41% P for H. verticillata and 0.11 to 0.38% P for P. crispus.
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A 12 month study by Shardendu and Ambasht (1991) in a tropical lentic wetland, found that the mean
annual biomass of four species of submergent macrophytes was highest for H. verticillata (54 g m-2)
which was present through the year, and lowest for Aponogeton natans with 12 g m-2 (the other species
were Potamogeton pectinatus ~42 g m-2 and P. crispus ~22 g m-2). Although the submergent species only
represented 6% of the total wetland biomass their involvement in net productivity around the submergent
zone of the wetland made the study of their nutrient assimilation an important investigation of the natural
wetland system. The low biomass percentage of these macrophytes was generally due to the absence of
the species in this zone for at least one of the seasons. The mean annual P concentration varied from 1.03
mg P g-1 for P. pectinatus and 1.08 mg P g-1 for H. verticillata which varied significantly seasonally
showing higher concentrations in the early growth phases and lesser concentrations when plants matured
in the rainy or late rainy months. Water column concentrations of P declined during the growing phases
of the macrophytes.
Various scientists have attempted to compare the efficiency of several commonly used aquatic
macrophytes in reducing nutrients and other organic matter under different flow rates of sewage
representing retention times of 1.5 to 3 days (100 to 200L d-1). Ansola et al. (1995) compared
Phragmites australis, Typha angustifolia, Iris pseudacorus and Scirpus lacustris cultures grown in small
glass fibre tanks which maintained a water depth of 20cm. They found that total P removal from
decanted village sewer wastewater (averaging 19mg P L-1) ranged from 47 to 61% depending on the
application rates and was similar for all macrophyte species, with the plants having a greater affinity for P
at higher retention times. The control tanks, however, removed a high proportion of P (40-47%) which
was most likely a function of substrate adsorption.
Rooted macrophytes obtain nearly all their nutrients from the sediment, whereas floating plants assimilate
nutrients directly from the water column. Eichhornia crassipes is often reported as an ideal free floating
plant to be used effectively to strip nutrients from wetland or pond systems due to its high productivity
and nutrient retention capacity (Reddy and Sutton 1984), but it is classified as a noxious weed in
Australia which prohibits its use as a wetland macrophyte in Australia. The plant is more effective in
warmer climates making it an ideal wetland plant for tropical environments but it possesses great
potential for blocking irrigation channels and rivers, spreading waterborne diseases and restricting
drainage. If the use of hyacinth is considered, management of hyacinth then becomes a major issue as
well as the constant consideration of harvesting to ensure optimum conditions for continual pollutant
removal. Data from Aoyama and Nishizaki (1993) support these findings stating that as long as there is
adequate area and harvesting for the plant to colonise, utilisation of water hyacinth would be a low cost
and easily managed effective water treatment system. Tripathi et al. (1991) also reported on the greater
potential for P removal by hyacinth than Pistia stratiotes (water lettuce), Lemna minor and Salvinia
rotundifolia during the summer wet season. Pistia and Salvinia are also classified as noxious weeds in
Australia and would not be considered for use in a constructed wetland system. Lemna dominated in the
dry cooler months (winter) but on an annual basis hyacinth had the highest nutrient removal capacity,
suggesting that the ideal free floating macrophyte strategy for nutrient removal in tropical ponds would be
to combine hyacinth and Lemna. Species of Chara have also been shown to absorb reactive phosphate
over a wide range of concentrations in ponded situations (Kufel and Ozimek 1994) and their capacity to
outcompete phytoplankton makes this plant a good choice as a potential emergent/submergent wetland
species. It is debatable, however, whether Chara would achieve a biomass sufficient to make it an
effective P assimilator (G. Lukacs pers. comm.). Johengen and LaRock (1993) showed that nutrient
uptake by Lemna sp. is significant and is advantageous as a free floating macrophyte for a wetland pond
system because it has the ability to restrict phytoplankton growth by competing for available nutrients and
limiting light availability. From mesocosm experiments, Lemna exhibited high removal rates (62%)
when compared to the sediment (49%) and the water column (21%), and was comparable in P
assimilation rates to the emergent macrophyte Pontedaria sp. The storage potential of P by floating
aquatic plants is short term, however, because of its rapid turnover, and if the plants are not harvested, the
stored P can be rapidly released into the water column during decomposition of detrital tissue. On the
other hand, emergent macrophytes have more supportive tissue and thus provide greater potential for P
storage.
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The role of plants in P retention has by and large been either not considered or grossly underestimated
until recently. Reeder (1994) reported that for a small 56ha marsh near Lake Erie (Ohio, U.S.A.)
plankton accounted for 67% of the water column biotic P removal with an uptake rate of 10g P m-2 y-1
compared to an estimated rate of 0.1g P m-2 y-1 by Nelumbo lutea (an emergent broad leafed macrophyte)
which covered 23% of the wetland. This estimate was much higher than that predicted by a model
previously developed from studies in the same area which estimated that 10-30% of any P flowing into
the wetland would be retained or transformed by algae, depending on flow conditions where it would
either be transported out of the system by viable plankton cells or deposited to the sediment of the
wetland (Mitsch and Reeder 1991). Similarly, a study by DeBusk et al. (1995) showed that a periphyton
filtration system removed at least 36.9g P m-2 y-1 from sugar cane runoff containing less than 100µg P/L.
If a wetland design includes phytoplankton or enhanced periphyton communities as a predominant P
removal process, management of the system must then ensure that overgrowth or blooms are controlled so
that the system can maintain efficiency. It is suggested that periphyton-based systems would be more
effective removing P from wastewaters providing the inflows are low load and low in P concentration.
Species composition of plankton and periphyton associated with macrophytes are strongly seasonally
dependent, particularly in warmer climates where a general shift from cyanobacteria in the summer to
diatoms in the winter is observed in most wetlands. Vymazal and Richardson (1995) list, for example,
the large range of species present in a study area in the Florida Everglades, and periphyton P
concentrations ranged from 117µg P g-1 (winter) to 454µg P g-1 (summer) for attached periphyton. The
periphyton P concentrations dramatically declined when the periphyton were detached. These biomass
estimates were considered low and reflective of non-enriched areas as periphyton P values exceeding
5000µg P g-1 have been reported for nutrient-enriched wetland sites. It was also shown that where
periphyton mass increased to such an extent as to form thick mats, the P assimilation capacity, and hence
biomass P was greatly reduced.
Table A.1 in the Appendix lists commonly used plants in international constructed wetlands with plants
considered suitable for use in constructed wetlands in Queensland, Australia, for the treatment of
wastewater. The table lists where possible, a comparison of mean annual production, mean annual
standing stock, phosphorus storage potential, phosphorus content and phosphorus uptake rates. The table
has many gaps which indicate that, particularly for plants considered suitable for Australian use, further
research is necessary to understand production potential and phosphorus storage ability of many of the
wetland plants. There are some cases of notably low estimates for mean annual standing stock (e.g. data
from Shardendu and Ambasht 1991) particularly when compared to other data, so it must be assumed that
some variability exists due to a number of factors that may not have been considered such as substrate
effect, the influence of periphyton, plant age, climatic conditions, analytical error, etc. It has also been
assumed from the data where it was not dictated, that most of the results have been expressed on a dry
weight basis.
Bacterial Assimilation
Bacterial assimilation is generally at the litter/detrital zone above the sediment. Detrital tissue is found in
wetlands dominated by macrophytes and generated as a result of factors such as the aging of plants,
macrophyte overcrowding, pest damage and disease. Bacterial decomposition of the litter is predominant
in the detrital zone and the principal form of nutrient transformation. Bacteria can assimilate P into their
cell structure; but again, as a steady state is reached within the bed, no net P removal may occur
thereafter. The biomass and productivity potential of bacteria in wetland systems, particularly their ability
to assimilate P, is largely unknown.
Physical Settlement and Accretion
Physical settlement and accretion is one of the most important mechanisms of nutrient removal which
could be exploited for P removal in natural and constructed wetlands. As a consequence wetlands are
assumed to act as a P sink since most of the P that is retained ends up in the sediment (Richardson 1985).
Providing that water retention is adequately addressed and the depth of the settlement area and flow rate
of the influent taken into consideration, the physical settling of P laden contaminants in a well oxygenated
system will result in a major proportion of the P being removed from the inflowing wastewater.
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Phosphorus accumulation in wetland substrates experiencing nutrient enhancement has resulted in peat
accretion rates that are proportional to the concentration of P within the water column above the
underlying substrate (Craft and Richardson 1993). The peat soils within eutrophic sectors of the
Everglades, Florida, U.S., have been shown by the previous authors to provide effective storage of
enriched P for extensive periods of time while at the same time providing a substrate stimulating net
primary production which serves to increase peat production. Kadlec (1994) supports this view stating
that the wetland biogeochemical cycle can operate to accrete new soils and sediments which contain P
and that the soil-building processes provide a more permanent storage of P.
Chemical Transformations
The manipulation of various forms of P by chemical means to a form that will become easier to remove
has been considered by several scientists. For example, the conversion of soluble P to insoluble
particulate forms by chemical means (i.e. primary wastewater dosing with flocculants or within-bed
means for adsorption or exchange) or by increasing pH in a photosynthetically active system, which
associated with high calcium levels could precipitate phosphorus as calcium-phosphate minerals (i.e. a
clear water lagoon with phytoplankton and/or rooted submergent macrophytes), would be an efficient
method of removal. Diaz et al. (in press) suggests that that calcium concentrations would need to be
greater than 100mg L-1 and the water column pH greater than 8.0 for effective P removal from surface
waters, but this process can often account for greater than 60% of FRP precipitation.
The ability of the beds to remove suspended solids ensures the retention of the contaminant if insoluble P
particles can be successfully produced. This idea has led to the concept that chemical dosing within the
bed, at a point where most of the initial suspended solids had already separated out, would be an effective
and economical means of P removal. This concept relies on the secondary step of substrate adsorption to
remove the formed insoluble particulate form and is discussed below. In any case, the ability of the
substrate to adsorb and retain P will be governed by the complex interactions of redox potential (and
therefore organic carbon), pH, iron, aluminium and calcium concentration and the amount of background
substrate P (Faulkner and Richardson 1989).
The direction of P flux across the substrate-water interface is regulated by the P concentration gradients
across the interface, pH of the water column, sorption/precipitation reactions at the substrate-water
interface, uptake by algae and macrophytes, the physico-chemical properties of the substrate and the
incidence of any bioturbation at the interface (Reddy et al. 1995). The substrate-water interface layer is
usually oxidised and its thickness is dependent on oxygen diffusion potential and oxygen demand within
the zone. This zone can therefore potentially function as a P sink by immobilising P into insoluble ferric
or calcium phosphate, as well as uptake and storage of P into the bacterial biomass. It is often assumed
that oxic or aerobic conditions completely prevent P release from the substrate but mass balance studies
by Ryding and Forsberg (1977) have shown that the release of P can be substantial from sediments to
well aerated waters. This is generally found when weakly buffered, low pH, low P concentration waters
(i.e. rainwaters) come into contact with sediments containing high concentrations of natural P.
Substrate Adsorption
The ability of wetland substrates to retain P depends on the physico-chemical characteristics of the
substrate (Reddy et al. 1995). If P is to be removed by means of the matrix material it must be contained
within the bed by ion exchange, adsorption or chemical reaction in an inert insoluble form. As these
factors have a finite capacity, P removal is expected to cease when the capacity is reached regardless of
the solution P concentration. The amount of P that a substrate will absorb will be determined by the P
concentration in the water column; the higher this concentration, the greater the absolute amount of P
which can be removed. Clay-type media with their abundance of aluminium, iron and calcium and large
surface area have the greatest potential to trap and hold P, but their very low hydraulic conductivity in a
bed designed for subsurface flow results in most of the water travelling across the surface and not making
contact with the bed proper, thereby precluding efficient P containment.
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Various laboratory experiments have shown that wetland soils can adsorb P. Richardson (1985)
demonstrated that organic soils were less suitable for P removal and that greater adsorption could be
achieved from substrates with a high amorphous iron and aluminium content, which is found in mineral
based soils. The presence of these ions under anoxic conditions resulted in more P being adsorbed by
substrates when the solution P was high (i.e. retained by oxides and hydroxyoxides of iron and
aluminium, and calcium carbonates) but more substrate P being solubilised and released when the
solution P concentration was low. This followed from a study by Hill and Sawhney (1981) which
concluded that reducing anoxic conditions caused by flooding cultivated soils actually increased P
mobility after prolonged P loading. They did show, however, that periods of resting water (i.e. standing
water due to no inflow) regenerated the sorption sites increasing the potential for additional P adsorption.
Gravel media, with their high conductivity, permit all of the water to flow within the bed but because of
their impermeable nature have only a limited surface area for adsorption, ion exchange, and/or chemical
reaction to take place. Once the active sites are utilised P removal ceases. Davies and Hart (1990) found
that unplanted and planted 30 x 5m channels of reed beds (Phragmites australis), unharvested over three
years, had virtually no P removal in a basaltic substrate. However, a channel with a low hydraulic
conductivity sandy soil substrate had an excess of 20% removal where most of the water flowed across
the bed surface. In the same study it was found that in the gravel bed channel most of the wastewater
solids were removed in the first 10 to 15m of the 30m reed bed channel so dosing at the beginning of the
final 10m section would be at a point where most of the void spaces in the gravel would be available for
the collection of insoluble P particles.
The presence of plant roots in the wetland substrate reduces the void spaces available for water-substrate
contact and hence the life of the substrate for P adsorption, but only to a limited degree as some of the P
content of the water will be transformed and assimilated by the plant root system. Most substrates within
the wetland beds will have a capacity to contain insoluble P and, if the wetland is designed with a
chemical dosing treatment system prior to the wetland or incorporated into the wetland system, the
wetland substrate would have a very long effective life. If, however, localised blocking of voids occurred
a small amount of surface flow would not be a serious problem as it would flow in the plant litter layer
and the litter would remove any phosphate particles until the flow re-entered the bed further down.
When alum dosing is used in wastewater to remove P (as insoluble aluminium phosphate) at low levels of
phosphate (10mg P L-1), there is competition from the formation of aluminium hydroxide and this
necessitates a higher Al:P ratio than 1:1. Formation of the hydroxide lowers the pH significantly and
reduces the formation of the aluminium phosphate because of its rapid increase in solubility as the pH
decreases below 7.0. The addition of lime (calcium hydroxide) with the alum prevents the formation of
acid conditions and thus improves the P removal. The issue of P reduction by dosing with lime and alum
led Davies and Cottingham (1993) to establish an effective dosing mixture of 150 mg L-1 lime and 50 mg
L-1 alum and found it to be economical and efficient for the removal of P within 30 x 5m gravel reed bed.
Dosing directly to the bed presented mixing problems resulting in the wastage of alum by the formation
of hydroxide instead of phosphate. The problem could be overcome by the use of a multi-stage system, a
planted bed for the removal of normal solids, BOD and pollutants; a small pond for the dosing, mixing
and sedimentation of the phosphate sludge; and a planted bed for final treatment and removal of any
insoluble phosphate carried over form the dosing pond.
Sediment or soil amendment within the channel bed by the addition or replacement of various substrates
could be considered to improve the P adsorption potential of soils by utilising material such as sandy
loam, industrial wastes or alkaline fly ash (Geohring et al. 1995; Cheung et al. 1994). Similarly, Mann
(1994) suggested that by selecting specific P adsorbing substratum, such as industrial by-products, instead
of relying on regional soils, P removal has the potential to be enhanced. Several potential problems exist
with the use of external sources of material for wetland channel beds. One is their potential to have
impacts on the groundwater environment by leaching and/or their potential toxicity to wetland plants.
Another is on a cost/benefit basis where the cost to the developer to purchase and provide the soil
supplement has to be measured against the requirement of the wetland for substrate amendment to
improve phosphorus removal.
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Jones and Amador (1992) suggest that many techniques used to determine estimates of soil and sediment
retention are mainly based on FRP, or orthophosphate, removal from water by soil or sediment columns
under saturated flow. They argued that differences observed between the apparent kinetics and
mechanisms of TP and FRP removal have ecological and environmental consequences as measurements
of FRP removal are likely to underestimate the extent and rate to which P retention occurs. The rate of
TP uptake was constant for a period of two days in peat soils from the Florida Everglades in saturated
flow, while the rate of FRP uptake slowed and followed saturation kinetics. The majority of TP removed
was within the 1-20µm range and it was shown that the primary mechanism for the removal appeared to
involve abiotic hydrophobic and ionic interaction mechanisms rather than a filtering process. The
removal of organic and inorganic P fractions from solution by physico-chemical processes is likely to
describe the P dynamics in Everglades soils more adequately than data from sorption isotherms using
only FRP. Much of the water within these marshes flowed vertically as well as horizontally through the
soils so much of the water was in constant contact with the soil (density 0.3 g cm-3).
Loss to the Atmosphere
Loss of P to the atmosphere as phosphine was considered by a number of early workers and it is still
debated to some extent (Devai et al. 1988, see also Brix 1993). Johnstone (1991), however, stated that
even if phosphine was produced it would be adsorbed by the soil and quickly transformed and would not
be liberated to the atmosphere.
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6.
REVIEW OF LITERATURE RELATED TO PHOSPHORUS RETENTION BY
NATURAL AND CONSTRUCTED WETLANDS
6.1
Australian Research
The predominant wetlands research in Australia is with constructed wetlands and their application to
improving water quality discharge for a number of industrial systems (i.e. sewerage treatment plants
(STP), processing plants, etc.). Much of this work commenced in the 1970’s (Mitchell et al. 1995) and
has been centred around south-eastern Australia. The use of constructed wetlands is, however, increasing
throughout Queensland with strong Queensland Department of Natural Resources initiatives to integrate
wetlands as a tertiary discharge treatment device for STPs and other applications in major regional areas.
Research into the effectiveness of wetlands to reduce P export has been limited to the last few years
although the data reported in many of the research publications and conference proceedings lists the final
concentrations of P export as almost a matter of interest. The literature reviewed here has been selected
on the basis of its reference to P reduction by natural and constructed wetlands.
Sewerage Treatment Effluent Water Quality Control
There are limited data and literature on P uptake and/or retention in natural wetlands. Finlayson et al.
(1986) found that the P load from a natural wetland receiving secondarily treated sewage effluent, from
the STP at Thredbo, N.S.W. in 1982, was reduced by 44% in the summer and 34% in the winter months.
Patruno and Russell (1992) reported on a natural wetland in Yamba, N.S.W. which successfully
‘polished’ effluent from an STP which serviced approximately 4500 persons over the last twenty years.
Phosphorus retention was on average 94%, despite the wetland not being in a pristine state. The wetland
drained into a marine lagoon (Wooloweyah Lagoon) and concern has always been expressed as to the
long term viability of the wetland as the loading is expected to increase due to a forecasted population of
8000 persons by the year 2000. Average P reduction was typically 5.5 mg TP L-1 to 0.34 mg TP L-1 at the
outlet to the lagoon. Little information was given on the description of the wetland, a common oversight
in many of the published articles on wetlands, so the predominant method of P removal cannot be
deduced.
Another natural wetland which has been successfully treating secondarily treated sewerage effluent is the
Racecourse Wetland in Wyong, N.S.W. (Soukoup 1995). The 90ha wetland has been receiving on
average 1.7ML effluent per day for twenty years from the STP which serves approximately 5000 people
for most of the year. The dominant macrophytes were mainly submerged (Myriophyllum sp. and
Potamogeton tricarinatus) and floating species (Ottelia ovalifolia, Azolla filiculoides, Lemna sp. and
Eichhornia crassipes) interspersed through a variety of trees and reeds. Notably the wetland is a
permanent lagoon which is covered by 0.3 to 1m water depth. Typical orders of reduction of P have been
10mg TP L-1 and 7.5mg FRP L-1 from the outlet of the oxidation ponds to around 50.0µg TP L-1 and 50µg
FRP L-1 at the outlet of the wetland to the Wyong river. Some slight seasonal differences were noted in
the retention/assimilation efficiency of P but were not considered significant.
Constructed wetlands in Australia have been more variable in efficiency when it comes to removing P
from effluent waters. Thomas et al. (1995) conducted a pilot study with a constructed wetland at the STP
in Wodonga, Victoria, utilising a sub-surface inflow/outflow design with and without macrophytes
(Schoenoplectus validus, Juncus ingens). Phosphorus removal was found to be approximately 13% mean
removal efficiency for each of the cases studied, i.e. unvegetated, vegetated (mixed and single species),
for several hydraulic loading rates and substrate sizings (gravel 10-20mm sizes) and through seasonal
cycles.
Studies in tropical regions of Australia are rare. Mockeridge (1995), however, reported at a National
Conference on Wetlands for Water Quality Control at James Cook University, Townsville, on the use of
wetlands to ‘polish’ secondarily treated effluent for re-use as a supply to the Town Common which is an
established water bird habitat. Six wetland channels (15:1 length to width ratio) were constructed and
planted with a variety of floating and emergent macrophytes (Typha orientalis, Lemna sp., Cyperus sp,
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Marsilea mutica (Nardoo), Phragmites sp. and Spikerush). Overall P reduction in the experimental
channels was negligible over the duration of the trials and it was found to be difficult to maintain the
planted vegetative communities without other macrophyte invasion or dominance by one or two species.
Some initial reduction in P was noted in the discharge from the channels early in the trial and when two
of the channels were put back on line but the capacity to assimilate P over any extended period by
whatever means was reached very early in the trial. Mockeridge (1995) did note, however, that the roots
and rhizomes of the macrophytes did have a higher P content than the macrophytes from the control
channels.
Vertical flow wetlands are an Australian development and the process has since been patented (Breen et
al. 1989). These small scale systems utilise upflow or downflow hydraulic patterns and rely on plant
uptake via the rootzone as one of the major nutrient removal mechanisms which supplements substrate
adsorption (Breen 1990; Rogers et al. 1990). In the case of an upflow system, influent is generally
introduced under positive pressure to the bottom of container housing the plant and substrate and the
effluent collected at the top of the container via peripheral drainage tube (the downflow system is the
reverse with the influent being applied to the top of the container). They are designed for use as small
scale treatment wetlands with the focus on single use domestic treatment where loading is low and
variable. There have been instances, however, where unplanted systems have been shown to be more
efficient in removing P than planted systems utilising Typha orientalis, Schoenoplectus validus, Cyperus
involucratus and Baumea articulatea (Heritage et al. 1995). The systems were loaded with 60L/day of
approximately 4 to 8mg P L-1 wastewater with a retention time of 5 days. Some seasonal differences were
noted over the thirteen month study whereby increases in outflow P concentrations where noted during
periods of old shoot senescence in the spring, particularly with the systems planted with Schoenoplectus.
The outflowing P was mainly found to be organic and particulate forms of P with little variation in FRP
which was always low in concentration, particularly in the first three to four months of the study when
absorption sites within the gravel substrate were more available.
Early work at the University of Western Sydney involved the determination of nutrient budgets for
wetland microcosms which again utilised the bucket approach and vertical upflow, i.e. a 10L bucket filled
with a gravel substrate and planted with macrophytes. Different macrophytes were used (Phragmites
australis, Schoenoplectus validus, Typha orientalis, Juncus ingens, Eleocharis sphacelata and Baumea
articulata) and compared to non-vegetated controls within the experiment whereby primary settled
sewage was applied for an approximate retention time of five days. The P load reduction was 99%
(Breen 1990) which contrasts with the performance of a planted horizontal trench system at the Wodonga
STP which received secondary treated effluent which only managed to reduce the P loading by 3%.
In recent years there has been a large body of wetlands research published by the University of Western
Sydney which have essentially reported the findings of applied research at several wetland treatment trials
at STPs in Byron Bay and Richmond, N.S.W. Bavor and Andel (1994) reported on preliminary work in
Byron Bay which involved discharging alum-treated effluent (dosed to reduce P concentrations in the
effluent discharge to 1mg P L-1) to a treatment system comprising a number of upstream experimental
scale units (rated at 120kL d-1), several small pilot scale systems and a 6ha broad acre wetland constructed
with a number of treatment cells receiving 1ML d-1. The upstream experimental scale units comprised of
four different units, each of which was monitored to determine efficiency in terms of P removal capacity.
The first unit was an all gravel substrate which was divided into six sectors - one control, two
Phragmites-planted, two Schoenoplectus-planted and one Typha-planted section. The preliminary results
showed that the annual discharge from this unit was 0.15mg P L-1, representing a reduction in P loading
by 85%. In preliminary results examining the capacity of individual species of macrophytes,
Schoenoplectus and Typha performed better than Phragmites. Unit 2 was another horizontal gravel
trench design planted with Phragmites followed by a grass planted open water meadow. The effluent
from this unit averaged a final concentration of P of 0.09mg P L-1. The shallow open meadow section
operated as a filtration system and enabled sediment adsorption of P. Unit 3 was an all-meadow,
intermittent feed system, consisting of two grassed meadows, alternatively flooded to 5-10cm over two
days and then drained over the next two days. P removal was not as good as the previous units for the
first five months and the average concentration of P in the outflow has been close to or less than 0.5mg P
L-1. Unit 4 was a shallow free water surface trench and results were not possible due to construction
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problems. The broad acre wetland has maintained low P levels (<0.5mg P L-1) with higher values
associated with breakdowns at the STP. With respect to the P removal capacity over the first two years of
operation an average of 80% reduction in P concentration, or 67% reduction in P loading has been
achieved.
Colleagues and students of John Bavor have investigated the P removal issue at the Richmond STP,
N.S.W. Sakadevan et al. (1995) used a free water flow multi-trench system to remove P (and nitrogen)
from secondary treated sewage under differing hydraulic loading conditions, P concentrations and
retention times. The trenches were identical in design and size (30m x 5m) with a depth of 1m and were
filled with clay to a depth of 30cm and topsoil to 30cm. Each trench was divided into seven segments
and planted with three macrophytes (Phragmites sp., Schoenoplectus sp. and Triglochin sp.). It was
found that the greatest P removal (62 and 63%) was observed in the trenches with the lowest hydraulic
loading (2000L d-1) and the highest retention time (15 days). On average the outflow P concentrations
were lower than their respective inflows but as is found when reviewing such work in many studies, some
decreases in the outflow concentrations are small in respect to the inflow concentration, and may not be
significant (e.g. one trench had an inflow of 8.3±0.5mg P L-1 and an outflow of 6.1±2.1mg P L-1). As
expected, it was found that the substrates from the trenches receiving the highest concentrations of P in
the inflow (~8mg P L-1) accumulated the most soil P.
Rob Mann, a former student of John Bavor, has presented several papers at conferences over the last few
years which report on the use of alternative and modified substrates in wetlands to improve P removal
efficiency (Mann in press; Mann 1995; Mann and Bavor 1993; Mann 1990). Most of his work was
conducted in field trials at Richmond with large trench systems (100m x 4m x 0.5m) along with
laboratory trials testing the P assimilation capacity of various substrates. In trials that compared an open
water gravel trench (T1) to a gravel trench planted with Typha sp. (T2), although 60 to 80% of P was
removed by both trenches, it was reported that more P was removed by T1 in the first 18 months only
because it contained more gravel, and that the capacity for P assimilation declined over the three years in
both trenches (Mann 1990). Additionally, release slugs of P were noted in the open water stages of T2
indicating that P removal will be highly variable, thus agreeing with the findings of other authors (Steiner
and Freeman 1990). At the end of the study, seasonal averages showed that the open water trench
removed 40% of the P compared to 25% for the vegetated trench and that the substrate capacity for P
adsorption declined after the first year of operation.
Mann and Bavor (1993) reported the findings of a similar study but included an additional gravel based
trench planted with Schoenoplectus validus. At best, only 40% of the inflow P could be removed by the
trench wetland systems with the control trench behaving similarly to the planted trenches. As the
substrate was again found to be the governing P removal factor, laboratory studies were used to
investigate the efficiency of material (such as industrial wastes) which could be used as alternative
substrates to enhance P removal by either substitution or modification of the existing substrate within
constructed wetlands. Mann (1990) concluded that selection of a local material for use as a substrate
within a constructed wetland should include an evaluation of the P assimilation capacity and its physical
and chemical characteristics, such as aluminium and iron oxides content, porosity, hydraulic conductivity,
particle size content and the calcium and magnesium content. Material such as blast furnace slag (160 to
420µg P g-1) and fly ash (260µg P g-1) were found to have much greater adsorption capacity when
compared to regional gravel soils (25.8 to 47.5µg P L-1), and taking into account the potential for
leachability and toxicity, are a strong candidate for substrate alternative and modifier for wetlands
substrate where P removal is a requirement of the wetland.
The other listed publications by Rob Mann (Mann in press, Mann 1995) are dedicated to management
issues with regard to the design of constructed wetlands for use as P removal tools. Design criteria for
sustainable operation were noted to be considered in relation to macrophyte harvesting requirements,
substrate selection, manipulation of the system to allow for additional stages and whether multiple units
are required for sustainability. The issue of monitoring was also considered which included the frequency
of effluent and substrate sampling, as well as appropriate testing parameters which can be based upon the
size of the treatment plants. Additionally, where wetlands are required for industrial effluent treatment, it
was suggested that further consideration needs to be given to characterise the influent and the monitoring
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based on these parameters. For stormwater treatment, monitoring was recommended for high flow and
flood events (during and following) where possible.
Mann (1995) also introduces chemical dosing as an addition to the wetland treatment system and makes
the point that where large P releases occur, an option of chemical dosing may need to be employed to
reduce P concentrations prior to effluent discharge to the waterways. The assumption is that when
treating secondary influent to a waterway where most of the P and nitrogen is in an oxidised form,
chemical dosing should be applied before the influent reaches the constructed wetland systems to reduce
the actual P load and thereby improve the sustainability of the wetland.
Adcock et al. (1995) have taken a step towards addressing the process of P removal by attempting to
determine the fate of P in a three year old constructed clay-based surface-flow wetland at Byron Bay
(West Byron Bay STP), N.S.W. They found that the plants contained higher P concentrations at the inlet
of the wetland and that the concentration declined with distance away from the inlet. Phosphorus
concentrations in the sediment were high (due to high background levels) but the plants were responsible
for the majority of P accumulation within this system (above ground biomass, roots and litter). The STP
system utilised an aerated activated sludge system dosing with aluminium sulphate reducing the
concentration of influent P to 1mg P L-1. The primary wetland (Unit 2a) was a subsurface flow wetland
(25m x 50m) planted with Phragmites australis which received wastewater directly from the STP catch
pond. It was then connected to Unit 2b by a series of V-notch weirs. Unit 2b (30m x 50m) had depths of
10-20cm and a retention time of two to three days. The macrophytic vegetation was dominated by
Leersia hexandra, Urochloa mutica, with some Typha orientalis, Persicaria lapithofolia, Aster subulatus
and Eichinochloa crus-galli which were all growing essentially hydroponically as there was no top soil
and penetration into the clay base was limited. Phosphorus loads decreased through the wetland over the
study period (inlet 400±130µg P L-1 and outlet 127±57µg P L-1) with the majority of assimilated P ending
up in the plant stems, particularly near the inlet site. The plants accounted for 46% of the total P retained
by the wetland while the standing water column P concentration accounted for less than 1% of the total P
retained. The sediment P concentration remained fairly consistent through the wetland and indicated that
water penetration did not exceed 2cm through the wetland. The high background P level made it difficult
to assess the amount of P adsorption by the soil during the study.
The CSIRO Division of Water Resources at Griffith has also provided an instrumental research team
working on experimental wetland systems. Rogers et al. (1990) understood that sediment adsorption
processes were instrumental for P uptake and became limiting when their capacity is reached. They
utilised a vertical hydraulic flow system (both upflow and downflow) and provided a high porosity
substrate which encouraged Schoenoplectus validus growth so that plant root contact was effectively
promoting plant P uptake. Over the forty week study several treatments were applied which varied the
application rates of P. During the first 29 weeks there was no significant difference in the performance of
the up and down flow systems (98% TP, 8.0 to 0.7mg P L-1; and 93% FRP, 4.0 to 0.1mg P L-1 removal).
When inflow rates where increased to 4L d-1 the upflow systems performance in removing FRP (85%)
and TP (82%) declined. The downflow systems, however, maintained high removal efficiency with FRP
from 10 to 0.2mg P L-1 and TP from 13 to 2mg P L-1. Unplanted control systems showed poor
performances and their assimilation capacity reduced from 60% reduction at week 1 to <30% at week 37.
It was found that the downflow systems were more efficient because of the high concentrated mass of
plant root mass at the surface of the systems which not only acted as a baffle to reduce hydraulic flow but
allowed for greater wastewater-plant contact. As a result 86% of the P load was found in the plant tissues
which was maintained throughout seasonal cycles. These systems have been ideally devised to treat oneperson loadings at an equivalence of primary settled sewage and the results indicated that the potential
effluent quality under these conditions would be acceptable.
Breen and Chick (1995) compared horizontal and vertical flow systems with respect to root zone
dynamics in nutrient retention with wastewater application. A comparison was made between 20L plastic
experimental buckets planted with Schoenoplectus validus for the vertical flow systems and a 50m
horizontal trench planted with Eleocharis sphacelata (pilot scale unit located at the Wodonga STP). Root
densities were found to be the critical factor determining efficiency of nutrient retention in the two
systems. The vertical flow systems operated more efficiently when the root densities were between 112
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and 251g m-2 which took between 221 and 412 days to achieve. The horizontal flow system was
disadvantaged by the less uniform vertical distribution of root mass which led to short circuiting of the
flow, reducing flow contact with the roots (see also Bowmer 1987) and nutrient retention by both root
assimilation and substrate absorption. The overall performance of the systems with regard to P removal
showed that the experimental buckets reduced approximately 90% of total P whereas the trench actually
released P (planted : winter; top 92.7%, bottom 84.6%; summer; top 93.0%, bottom 88.3%; unplanted :
winter; top 78.8%, bottom 80.5%; summer; top 29.1%, bottom 35.7%; trench : winter -1.2%; summer 14.8%). This compares to a vertical downflow process, reported in a poster by Rogers et al. (1990),
which assumed that 80% of P could be removed by the plants directly before the remaining P would be
able to be adsorbed to the substrate.
The efficiency of a series of pilot vertical flow wetlands located at Coffs Harbour, N.S.W. to treat primary
settled sewage was reported by Chick and Mitchell (1995). The study was based on monitoring the
inflow and outflow water quality on a weekly basis over a year (Jan 1991 to Feb 1992) and found that, in
general, the removal of nutrients was lower than expected despite excellent removal of BOD, suspended
solids and faecal coliforms. The overall treatment system was designed to treat 130KL d-1 whereby
several vertical flow wetlands (VFW systems 1-4) dealt with 10KL d-1, and VFWs 5 and 6 sustained
85KL d-1 (VFW 5 had a modified design). The VFW were 12m2 at the top reducing to 10m2 at the base,
filled with a gravel substrate and planted with emergent aquatic macrophytes with a theoretical retention
time of 5 days. The VFWs removed about 37% of the P concentration and load within the system but it
was shown that during extended retention times the removal capacity increased to 69% with 10 days
retention. The longer retention periods favoured plant uptake, adsorption and transformation, and microorganism assimilation processes. The concentration of inflow was highly variable (2 to 12 mg P L-1)
whereas the wetland outflow remained within 2 to 5mg P L-1, which is still higher than most Australian
regulatory bodies permit for discharge to receiving waters. Like other systems which remove some of the
P loads, the avenue is then open for other options to be implemented to apply additional forms of P
removal and disposal, i.e. chemical precipitation. A limitation of this system is that the flow was directed
upwards which has been shown in laboratory experiments by Rogers et al. (1990) to have a lower
capacity to reduce P than downflow vertical wetland systems.
The Environmental Sciences Faculty at Griffith University and the Queensland Department of Primary
Industries has had access to ten pilot wetlands in Queensland which have provided the researchers with
data based on different wetland configuration and application, and the effectiveness of different
macrophytes types and species for a variety of different climatic. This work is particularly important as it
involves the monitoring of wetlands constructed in different climatic conditions (the Queensland climate
extends from arid to sub-tropical to tropical conditions) and therefore provides information on warm
climate wetlands which has not been available in Australia to date. The research describes the wetlands
in use throughout Queensland with several case studies already been reported at national wetland
conferences (Bolton and Greenway 1995, see also Mockeridge 1995 - described earlier) presenting
preliminary results of various established wetlands (in some instances over 2 years old) used to ‘polish’
wastewater effluent (Townsville, Ingham and Blackall). Overall, the wetlands reported showed greater
than 40% reductions in BOD, TSS, and TN, with TP reduction at less than 10%. In all cases, as might be
expected in the warmer conditions, the wetlands were quickly colonised with additional species of
macrophytes. A potential problem in warmer climate constructed wetlands is increased density of plant
biomass which will most likely cause a decrease in influent detention times but additionally the increased
productivity of the plants might increase the potential for nutrient uptake. Phosphorus removal variability
reported for some of the aforementioned wetland systems may in fact be due to the increased productivity
because despite greater P uptake and accumulation in the plants (Greenway and Simpson 1996) increased
litter production and microbial release would be expected in these conditions. These highly productive
systems evident in tropical wetlands might suggest that macrophyte harvesting is an advantageous option
to be considered in tropical constructed wetland design.
In a follow-up study of a greater variety of wetlands, macrophytes in constructed wetlands were found to
maintain higher productivity indicating a tolerance of nutrient enriched waters and were found to have
higher nutrient bioaccumulation than similar species grown in natural control wetlands (Greenway in
press). The maximum concentrations of P accumulated (uptake rates were not established) (c.f. other P
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contents in Table A.1 of the Appendix) were very high and this suggested that potential for nutrient
removal utilising a variety of submerged and free and attached floating macrophyte species was very
good.
Pot trials using several species of Melaleuca trees under different conditions of effluent application were
conducted by Bolton and Greenway (1995) to determine their effectiveness in accumulating P. Melaleuca
commonly grows in seasonally inundated conditions and is often found fringing naturally occurring
wetlands in Queensland. Bolton and Greenway found that all species (M. quinquenervia, M. alternifolia
and M. leucadendra) accumulated high concentrations of P in the leaves with M. alternifolia
accumulating twice the levels of the other two species. It was also found that P was also excluded by the
Melaleuca in high P conditions as higher levels of P were found in senescent leaves, when alternatively in
lower P treatments P recycling from leaves was evident due to very low concentrations of P in senescent
leaves. These findings, and the generally tolerant nature of Melaleuca spp. in a variety of adverse
environmental conditions, suggested that they may have very effective potential for use in wetlands.
Water Quality Control of Rural Runoff
Work conducted by Max Finlayson and co-workers in the 1980s looked at the use of wetlands to improve
the quality of effluent discharge from rural industries such as piggeries, chicken farms, abattoirs, etc.
Finlayson et al. (1987) reported on a pilot scale linear trench system to treat effluent from a piggery
which was vegetated with Typha spp. but was relatively unsuccessful due to the concentration of the
effluent and the associated problems with very eutrophic standing water, such as odour, algal and
mosquito problems.
Finlayson and Chick (1983) showed that three trench systems with gravel substrates separately containing
Typha spp., Phragmites australis and Schoenoplectus validus slightly reduced phosphorus from effluent
originating from a chicken abattoir. Similar odour, algal and mosquito problems were encountered in the
trenches as the systems were initially designed to allow free standing water. Changes in design and
inflow rate allowed for water levels to be maintained below the surface of the substrate. Effluent from
these systems was clear and nutrient levels were markedly reduced.
The effluent from a rendering plant was extensively investigated by Bowmer (1987) utilising gravel based
trench systems vegetated with Typha spp., Schoenoplectus validus and Eleocharis sphacelata. She found
that by comparing the vegetated trench with an unvegetated control trench, the vegetated system was
prone to short circuiting, prompting preferential flow along the base and the sides of the trench.
Approximately 40% of phosphorus was retained in the summer months which was presumed to be a result
of substrate absorption.
Irrigation Drainage Water Quality Control
Natural wetlands have been used to control the quality of water that is discharged from agricultural
drainage. En-route wetlands, or in-line wetlands, are characterised as wetlands with a single input and
output and Cottingham (1995) has reviewed the effectiveness of four en-route wetlands in the Goulbourn
Irrigation Area, N.S.W., for removing nutrients from irrigation drainage. The study area concerned is
described in more detail in an establishment report (Milestone Report No. 1) to LWRRDC (Cottingham et
al. 1994). The study was in an early stage and the data presented suggested that nutrient levels,
particularly P, were higher in the leaving waters than the influent. It was suggested that as the wetlands
were not ‘constructed’ for nutrient reduction, unregulated discharge of irrigation discharge would result in
overloading and therefore reduced retention efficiency. As found in a similar study by Raisin and
Mitchell (1995) in the Kiewa river catchment, there appeared to be a seasonal influence in wetlands
nutrient removal efficiency of agricultural runoff where net retention was found in summer and net
release in the winter months. A conclusion drawn by both sets of researchers was that where wetlands
have to be used to efficiently remove nutrients from irrigation drainage, the processes involved in nutrient
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removal/retention have to be understood so that constructed wetlands can take into account the design
features necessary to provide these processes.
Wetlands in the highest area of P export from agricultural areas in the Peel-Harvey catchment, W.A. (the
Meredith drain sub-catchment) were examined by Chambers et al. (1993) to determine their potential as
nutrient sinks. It was found that in the three wetlands studied which were dominated by the sedge
Lepidosperma gitudinale and surrounded by fringing Melaleuca sp. (1, in a reserve and undisturbed; 2, in
farmland and undisturbed; and 3, in a farmland and disturbed), the surface sediments were the largest
store of P and that the vegetation mediated the P pathway and its transformations, and through the
formation of litter, its contribution to the sediment. The undisturbed wetlands had a higher organic
composition in the soil (peat), higher vegetation to water ratios and much lower P concentrations in the
water. Bucket experiments with Lepidosperma gitudinale and its peat substrate showed that after a three
month establishment period, the P reduction from a source containing 9.9mg P L-1, which was continually
applied via recirculation, was 67% in the first week and 98.8% after 40 days. Similar experiments were
conducted varying the substrate, P application concentration and flow rate which showed that these were
important considerations when attempting to design efficient systems if interested in P retention.
Lepidosperma after three months establishment in either sand or peat achieved P reduction of 92 and
91%, respectively, with the application of 19.7mg P L-1 at an application rate of 42.0mg m-2 d-1 at 2.4L m2 -1
d over 59 to 161 days, which was a greater reduction than achieved for the unvegetated sand or peat
(59% and 50%, respectively); and better than that achieved for higher application rates of 11.3mg P L-1 at
an application rate of 120.7mg m-2 d-1 at 12.0m-2 d-1 over 13 to 45 days (48 and 60% for Lepidosperma in
sand and peat, respectively) and 19.6mg P L-1 at an application rate of 248.3mg m-2 d-1 at 12.0L m-2 d-1
over 48 to 55 days (63 and 72% for Lepidosperma in sand and peat, respectively). When the substrate
volume was increased from buckets to larger scale containers it was shown that the substrates were not
overloaded and able to retain higher loads of P from 10 to 200 mg m-2 d-1 over a 68 day period. The
conclusions stated, however, that although wetlands, natural and constructed, had the capacity for nutrient
uptake from inflowing water, they may be of limited value in climates generating seasonal rainfall which
produce high volumes of runoff in short periods of time. The greater volume of runoff may overload the
wetland and disturb the vegetation which may in turn release nutrients to the receiving waters. The
design of constructed wetlands therefore needs to accommodate the higher flow rates in these climates,
including tropical regions where wet season rainfall can generate large volumes of runoff, and provide
mitigation to ensure that flow rate is kept constant.
Stormwater Water Quality Control
The Blacktown City Council, N.S.W., implemented a 75ha stormwater control channelled wetland at
Plumpton Park which was designed with a preceding gross pollutant trap (GPT) and 28 species of
indigenous rooted emergent macrophytes. On a mass loading basis P reduction was estimated at 27%
over an annual cycle which was retained by the wetland, with little influence in P reduction by the GPT
(Hunter and Claus 1995). The order of P decrease through the wetland was 0.01-1.97mg P L-1 at the
source to 0.01 to 0.18mg P L-1 at the outlet (averaged for wet and dry conditions). The GPT was found to
be an integral part of the wetland design, not only because of its role in removing coarse litter and
sediment, but that it reduced inflow velocities thereby reducing the potential for scouring and
sedimentation in the initial stages of the wetland. Extended studies are required to determine the long
term effectiveness of the system.
6.2
International Research
Much of the international research on constructed and natural wetlands with respect to nutrient retention
and/or removal has been carried out in Europe and North America, though some work is reported from
Asia and New Zealand. This section of the review will discuss aspects of this work that relate to P
removal, P dynamics and warmer climates.
Europe
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European countries have utilised constructed wetlands to a varied degree over the past ten to twenty
years. The more commonly used systems are essentially horizontal subsurface flow systems because they
are a cheap alternative to secondary treatment, particularly for sewage effluent. Stringent legal rules have
now led to the development of highly efficient vertical flow systems as well as combined systems (i.e.
constructed wetlands in association with conventional biological plants) (Haberl et al. 1995). Vymazal
(1995) presents an inventory of the constructed wetlands in the Czech republic which have a high
capacity for reducing BOD, COD and TSS for a variety of treatment areas and population but limited
potential for P removal. In general, the average P reduction capacity for European wetlands is 47.1%
(n=338, Borner 1992, in Haberl 1995).
A pilot scale system in Slovenia has reported a P reduction of 97% from greywater (25.7 to 0.7mg P L-1)
and 71% from secondary sewage (1.0 to 0.3mg P L-1) during the first months of use (Urbanc-Bercic and
Bulc 1995). This system relied on vertical flow through an initial bed system followed by a horizontal
subsurface flow trench to optimise nitrification and denitrification processes. The initial beds were
loaded with 30mm d-1 of secondarily treated sewage and greywater and were made up of two
compartments of differing substrate sizes (A, coarse sand and gravel; B, fine sand and gravel) which
drained into the horizontal trench at a rate of 1.5 mm d-1. The finer grained sediment in the vertical flow
system (VFS) was more efficient in removing P from the greywater (95%, compared to 28% for the
coarse grained substrate) but little difference was noted between the substrate grain size for the secondary
sewage (A, 36% and B, 43%).
The reduction of non-point pollution from agricultural runoff is not an easy task and often relies on the
knowledge of pollutant flux concentrations and the transfer mechanisms. Dorioz and Ferhi (1993)
examined the processes involved in a 14ha experimental watershed which flowed into a 3ha wetland with
a mean outflow of approximately 6L s-1 and a maximum flow during storm events of 1000L s-1 within the
Redon River watershed, France. The mass balance equation showed that the wetland received measured
P export coefficients of 0.6kg ha-1 y-1 of which half the P export was FRP. The P fluxes from the
watershed were very variable and essentially transferred during stormflows. These flows intercepted by
the wetland were shown to decrease P loads by 75% (via decantation and precipitation) which when
extrapolated to the whole Redon watershed (33km2), accounted for 50% of the P load. The buffer zones
generated by the wetlands therefore played an important role in the improvement of water quality from
agricultural areas particularly to receiving waters.
The United Kingdom Water Industry have been using Phragmites reed bed treatment systems for the past
ten years and research into their use and effectiveness has produced 200-300 operational wetlands which
have finally become appropriate and efficient systems for secondary and tertiary treatment and
nitrification for English villages (Cooper and Green 1995). Their focus is towards a mainstream
treatment process for the removal of BOD, TSS, ammonia and organic nitrogen with little consideration is
given for P removal potential. This treatment is achieved via horizontal flow, vertical flow wetland
systems or a combination of both. Green and Upton (1995) present more specific data on the performance
of secondary, tertiary and stormwater Phragmites reed bed wetland systems under the control of Severn
Trent Water, Birmingham, which is a privatised regional water authority serving a population of
approximately 8 million. The privatisation has brought much more emphasis on the legal conformity to
discharge requirements which has meant that the wetland design criteria has had to be successful for over
a hundred sites. The Yorkshire constructed wetlands have been operating since the mid 1980s, although
there were examples cited from 1903 (Earby) and 1964 (Highroyd) (Hiley 1995). These were mainly
surface flow, subsurface flow (soil and gravel) and raft lagoon (floating algal mats or dense macrophytes)
wetland systems with inflow volumes varying from 0.02 to 1.3m3 m-2 d-1. The P removal capacity was
minimal and the P concentrations, relative to the load, were highly variable and could not be correlated
with the type of wetland used. The surface flow wetlands achieved high effluent standards but generally
lower load removal than subsurface flow systems.
North America
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In North America there are a number of wetland studies running throughout the United States and Canada
which vary from large open natural wetland treatment systems for agricultural and stormwater runoff to
small horizontal flow systems for sewage treatment.
A study by Gehrels and Mulamoottil (1989) of the transformation and export of P from an 18ha natural
wetland (predominantly Typha sp. marsh) in Ontario, Canada, found that total P imports were nearly
double the exports and orthophosphate export was 22% greater than imports. Phosphorus budgets
indicated that surface P loadings to the wetland were 2.0kg ha-1 (TP) and 0.4kg ha-1 (FRP) which
accounted for 75% (TP) and 68% (FRP) entering the wetland from an 130ha agriculturally dominated
catchment. This led to the assumption that internal processes within the wetland transformed the
sediment bound P to available P which was then seasonally discharged in fall with higher water levels
promoting anaerobic conditions and at the end of the growing season when leaching occurred from
decaying vegetation. Very high retention rates were observed in the spring and 95% of this retained P
fraction was bound to the substrate. These findings showed that seasonal fluxes in environmental
conditions are limiting to the overall efficiency of wetlands and must be considered in the overall design
criteria if constructed wetlands are to be of use to water treatment managers.
A cold climate study was conducted on a small wetland treating fertiliser leachate from a 10km2 athletic
field prior to drainage into Lake Tahoe, U.S.A. (Reuter et al. 1992). The 660m2 (33 x 20m) wetland was
excavated to 1m depth and filled with sub-6mm gravel substrate and vegetated with Typha latifolia.
Inflow to the wetland was non-regulated and a function of climate (i.e. fall precipitation, winter snow,
snow melt and dry summers) with an estimated runoff volume of 873m3 for 1988 study period and
temperature range of -3 to 21°C. The wetland system had limited ability to remove FRP from the influent
with an annual removal capacity of -28% (concentration) and -41% (load) meaning that the system acted
as an exporter of FRP. During the summer months the system assimilated FRP, removing 14% but
dropped to -31% during the winter. The authors suggested that the assimilation in summer was mainly
due to plant uptake, but the poor annual retention was a result of the contractors using unwashed gravel to
fill the wetland bed during construction. Particulate P removal was more effective with an average annual
reduction of 47% (concentration) and 44% (load), dominated by the summer removal (68%). The
wetland was therefore a sink for the particulate P, unlike FRP which remained in solution and was easily
transported out of the wetland during periods of water flow. A compounding factor appeared to be that
the Typha had not established sufficiently for it to effectively reduce FRP, but then again higher
production rates would be expected in less mature growing plants. The study was preliminary and should
have been allowed to establish for a longer period of time for an accurate assessment of the P removal
capacity of this wetland system.
A practical design for a constructed wetland to improve the quality of agricultural runoff is shown in
Higgins et al. (1993) which treats runoff from a non-point source prior to discharge into Long Lake,
Maine, U.S.A. The design utilised an inlet flume which directed inflow to a sedimentation basin (5 x
46m) after which it was evenly dispersed by a level lip spreader (0.6 x 15m and 0.46m deep) over a grass
filter strip (38 x 46m) used to detect the presence of pesticides and herbicides (6% slope). The overflow,
and subsurface flow was then directed to the 980m2 wetland (2% slope), planted with Typha latifolia
which buffered the flow into a detention pond (15 x 46m and 2m deep). The wetland-pond system had a
significant role in decreasing the impact of desorbed FRP. Agricultural soils above the treatment system
had a low pH (therefore an increased ability to adsorb P) but when they were eroded during increased
surface flow from storm event runoff and entered surface waters (which may have higher pH), the soils
ability to hold P decreased. The pond maintained a pH greater than 7.0 which allowed for P assimilation
by the wetland and pond biological community. This may have been due to strong algal productivity and
a predominance of calcium and magnesium bicarbonate composition in the water as the pond sediment
was found to have 12% of the influent TP recorded over the study period. Outflow was then directed
through an outlet flume to a 46m long vegetated swale that drained to receiving waters before flowing
into the lake. Annual removal efficiencies for TP between 1990 and 1991 were 82 to 92%, although
seasonal removal variability was high (in spring the outflow exported more P than imported - mainly due
to outflow exceeding inflow due to a higher groundwater table). It was concluded that, for the amount of
land required and the relative construction costs, the system worked effectively because most of the P was
associated with the solids.
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An upland wetland wastewater system was used to test the effectiveness of wastewater treatment for a
single family home in North Carolina by House et al. (1994). It consisted of a mound which provided an
aerobic environment to effectively reduce P concentrations by precipitation as calcium and magnesium
phosphates which was then followed by a wetland cell. Two emergent macrophyte species were
compared, Typha augustifolia and Phragmites australis, with P. australis proving to be more effective.
The evaluation over 15 months showed that 86% of the P could be removed by the mound due to its marl
component, and that the remaining P reduction by P. australis was due to the higher above ground
biomass contributing to a greater accretion of P as organic matter. This indicated that substrate
adsorption and P precipitation within the mound component was most effective at the removal of the P
while the wetlands contributed little.
The Des Plaines River Wetlands Demonstration Project near Chicago, Illinois, has been used to assist in
evaluating the use of wetlands to improve the turbid river water quality of the Des Plaines River which
drains an urban and agriculturally impacted catchment, and assess the wetland function. Four wetlands
have been used, varying from 2 to 3.5 ha dominated by Typha spp., with a maximum depth of 1.5m (Hey
et al. 1994). The data showed that regardless of the relatively low concentrations of nutrients, the
constructed wetlands were effective in reducing the nutrient loads and clarifying the inflowing water.
Inflow total P concentrations were significantly reduced at the outfalls of each wetland (155µg P L-1 to 2 48µg P L-1, accounting for a 52-99% reduction of P) with the major removal mechanism assumed to be a
settling of sediment particles and biological uptake. Improved efficiency within the second year of
operation was attributed to increased plant uptake and carp control efforts. In a concurrent study by
Christensen (1994), the role of changing flow conditions in the wetlands on P retention was modelled and
it determined that increasing inflow rates resulted in increased P retention but a decreased retention
efficiency. A subsequent follow-up paper by Mitsch et al. (1995) showed that average P concentrations
decreased from a three year average of 176µg P L-1 by 64-92% (11 - 40µg P L-1) in low flow wetlands
and by 53-90% (12 - 57µg P L-1) in high flow wetlands and that intensive sampling showed that retention
in the low flow wetlands was better than high flow wetlands. Clear gradients of P removal were shown in
the high flow systems whereas the low flow wetlands exhibited near homogenous concentrations of P
within the system. The paper supported sedimentation as the predominant removal mechanism with some
capacity for macrophyte uptake and a lesser involvement by periphyton and microbial communities. The
constructed wetlands were shown to retain 0.5 to 3.0 g m-2 y-1 which agreed with the values suggested by
the model developed by Christensen (1994).
Two-stage lagoon systems are commonly used in the dairy industry and Chen et al. (1995) presented the
preliminary results of a project at Louisiana University using several types of aquatic plants as a tertiary
treatment to improve wastewater generated by the dairy feeding lot washings. The aquatic plant systems
consisted of a pond with black willow (Salix niger, selected because of its rapid growth in wet conditions)
and duckweed (Spirodela sp.), and a pond with water hyacinths (Eichhornia crassipes, selected because
of its rapid growth rates in summer months). Hydraulic loading to the ponds varied from 0 to 107 m3 ha-1
d-1 and they were found to be very effective for TSS, BOD and faecal coliforms. Phosphorus removal was
not significant (i.e. non-apparent) in either pond. It was suggested that harvesting the plants might have
improved the P removal, although given the high concentration of the lagoon effluent (20-125mg P L-1)
the system was probably superloaded. Alternatives such as dosing should have been considered to reduce
the P loading prior to the wastewater inundating the macrophyte systems to make them more effective.
A great deal of research is currently being conducted in the Florida region, particularly with emphasis on
the bioremediation of P from agricultural wastewater and runoff. Pioneering work in central Florida was
conducted by Dolan et al. (1981) on P dynamics within a freshwater marsh dominated by Sagittaria
lancifolia, Pontedaria cordata, Panicum spp. and Hibiscus sp. receiving secondarily treated effluent
(approximately 38g P m-2). The results showed that over the initial study year on 2km2 plots constructed
within 32ha of marshlands, 97% of the P was removed by the wetlands, little P was assimilated by the
groundwater and there was a much higher net production in the plant biomass. The effluent discharged
from the plots remained close to the background marsh and receiving water quality. The substrate stored
approximately 69.2% (assuming the validity of the soil uptake calculation), the below ground plant
biomass stored 23.2% and the dead above ground plant material accounted for 5.2% of the P mass. The
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substrate was typically acidic (pH 4.5) and consisted of approximately 1.5m of highly organic peat soil
containing sand and clay which was generally covered by water to a maximum depth of 1m during the
year. There was no reported flush or release of P during any seasonal stage of the study. The marsh
increased its P retention capacity by virtue of the fact that peat production continued because of the
higher plant net production brought about by the constant effluent application. Peatlands are noted for
their P retention capacity as a function of the assimilation capacity of both the organic soil and the litter
zone.
In another Florida based study, DeBusk et al. (1995) utilised two very different methods to ‘polish’ the
wastewater: i) a nutrient source for the culture of the wetland macrophyte Pontedaria cordata receiving
dairy wastewater with P in excess of 1.7mg P L-1, and ii) periphyton filtration of sugar cane runoff
containing less than 100µg P/L. The uptake of P in the P. cordata systems varied from 0.2 to 66.8 mg P
m-2 d-1 depending on the age of the plants and season (growth was very reduced in young seedlings and in
winter conditions). On the other hand the periphyton system was not limited by the cooler months
averaged 101 mg P m-2 d-1 during the winter months (December to May).
Gale et al. (1994) utilised an experimental wetland site in Orlando, Florida, to determine the P retention
capacity of wetland soils for the treatment of treated wastewater and the likely mechanisms of P removal.
During a 21 day retention period, the constructed wetlands (sandy, low organic matter soils) retained 5266% of the added P, compared to the 46-77% retained by the natural wetland (high organic matter soils)
from a 8.5mL d-1 inflow which was relatively low in nutrients, i.e. TP values around 0.1mg P L-1 (the
control was a natural wetland which only received rainfall and runoff inputs). The organic P pools and
the iron-aluminium-bound fractions controlled the P chemistry in both wetlands and anaerobic conditions
increased P solubility and availability. The removal followed first order kinetics so that the rate of
removal was proportional to its concentration in the inflowing water. The results did not take into
account the uptake by plants so the total P removal was expected to be underestimated.
Artificial wetlands are considered very valuable for lake protection and restoration in the U.S. (Cooke et
al. 1993). Their capacity to reduce inflow rates and improve water quality of the inflowing waters makes
them a necessary link within watersheds which incorporate either lakes or reservoirs within their
catchment. Unfortunately, the amount of wetland alteration and removal over the last fifty years has
resulted in many lakes and reservoirs located in developed regions to develop eutrophic conditions. For
example, Holland et al. (1995) stated that over a ten year period approximately 40% of the wetlands
identified by the National Wetlands Inventory (NWI) within the metropolitan area of Portland, Oregon,
had been lost to human activity or were missing due to drought. Lowe et al. (1992) proposed an
interesting idea which will hopefully culminate in the successful restoration of Lake Apopke (Florida)
which is hypereutrophic and suffers from algal problems all year round. As the problem is associated
with agricultural P runoff, of which the predominant form is particulate P, they proposed to recirculate
lake water through constructed wetlands. A combination of this proposed system which will effectively
filter out the P and the restriction of P application to the watershed area is predicted to allow the lake to
return to normal within sixty years. The use of wetlands to improve inflowing water, however, has not
always been successful as reported for Lake McCarrons, Minnesota (Oberts and Osgood 1991).
Vincent (1994) reported on the efficiency of a wetland that was built on the fringe of a lake at an artificial
beach to treat recreational water in the City of Montreal, Canada. The wetland system occupied 20km2
and consisted of four ponds of varying depth with over 100,000 macrophytes of different species and
treated water that was pumped from the swimming area to series of ponds before being pumped back to
the lake (Pond 1, 1m depth : Eichhornia crassipes, Hydrocharis morsus-ranae; Pond 2, 0.5m depth : Iris
versicolor, Phragmites communis, Scirpus acutus, Typha latifolia; Pond 3, 0.8m depth : Alisma triviale,
Elodea canadensis, Myriophyllum spicatum, Pontederia cordata; Pond 4, 1m depth : Elodea canadensis,
Potamogeton pectinatus, Vallisneria americana). The annual organic P loading reduced by 82% in 1990
over the course of the ponds, whereas only 38% of FRP and 15% of TP were reduced and the reduction
was more significant after Pond 2 where the highest plant density was noted. This reduction was not as
marked in 1991 (Organic P, 57%; FRP, 26% and TP 0%) particularly when TP loads in all of the ponds
were significantly higher than the inflow (assumed to be due to bank instability in Ponds 1 and 3 and the
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presence of large fish). The authors suggested that the greater reduction of organic P was probably
associated with epiphytic algal colonies growing on the submerged plants.
D’Angelo and Reddy (1994a) reported on the efficiency of a constructed marshland (from previously
drained agricultural land) in treating recycled nutrient rich water from the hypereutrophic Lake Apopke,
central Florida (see also Lowe et al. 1992 - mentioned previously). Lake water was recycled through the
wetlands with a 3-12 day retention period and this eventually resulted in a floc sediment layer forming on
the native peat soil surface. The soil-water flux over the first 13 months resulted in high concentrations
of water column FRP (31mg P L-1) which reduced to 0.8mg P L-1 after ten months. Microbial degradation
and nutrient regeneration from settled labile organic matter supported the nutrient flux to the water
column and after the thirteen month study period accounted for 65% of the FRP within the water column.
Floc sedimentation was due to the mineralisation of the settled organic matter under anaerobic conditions
between the floc sediment and peat soil layers. The microbial process which mediated the decomposition
process was supported by the supply of electron acceptors (D’Angelo and Reddy 1994b). The recently
deposited, organic rich soils maintained a strong demand for electron acceptors which is presumably due
to the microbial catabolic process. As aerobic and anaerobic decomposition by fermentation, SO42reduction and methanogenesis accounted for P and nitrogen resuspension as soluble products (rates of 3.3
to 14mg N L-1 d-1, and 0.5 to 0.6mg P L-1 d-1), these processes therefore depended on the availability of
electron acceptors within the wetlands. The nutrients that were made available by this process were then
translocated to the water column by diffusive and advective transport.
Work in New York state examined specialised substrates to add to wetland sediments to increase the
removal capacity of P (Geohring et al. 1995). Various wetland substrates were investigated on the basis
of their potential P assimilation capacity, some of which were the waste products of industrial processes,
i.e. wollastonite tailings (calcium metasilicate mixed with ferrous metasilicate), iron ore tailings and paper
mill waste coal fly ash. The other substrates examined were garnet (ferrous metasilicate), sand with
oxidised iron and sand oxidised with aluminium. Wollastonite was shown to be the best adsorbent of P
adsorbing 5mg P per gram of substrate in a 5 to 1 solution to substrate mixture and was more effective
than sand with iron oxide over a greater pH range. Ninety percent of the influent P was retained in the
wollastonite substrate as it was passed through 1m long columns over a retention time of 12 hours. The
coal fly ash was a disappointment as it only retained 20% of the P and this was interpreted as being a
result of not being stabilised with lime before use. The sand treated with aluminium showed poor
retention showing only 21% P removal at pH 4.2, and 0% at pH>6.0. This was found to be due to the
aluminium oxide source not being an activated form and hence not very soluble. Sand amended with iron
oxide exhibited the highest removal of P (97%) below pH<7.0 buts its effectiveness dramatically
decreased at higher pH (in fact P was again released at the high pH values). The examination of grain
size and P retention showed that finer grain sizes, particularly the fine wollastonite, increased the capacity
for P adsorption particularly at pH 7.0.
Asia
Recent Asian studies follow the trends of the north American and European regions as to the steady
increase in utilising wetland processes for domestic sewage treatment systems. Hosomi et al. (1994)
reported on a four year study of a 1224m2 natural wetland on non-cropping farm land (former rice paddy)
vegetated with emergent macrophytes (Phragmites and Typha spp.) which treated domestic sewage from
a small residential area in Japan. The results were expressed on a seasonal basis as the authors were
concerned that microbial activity would decrease in the winter months thereby reducing the effectiveness
of the wetland. The average influent flow rate was 38.9m3 d-1 over the four year study period with TP
reducing from 0.97mg P L-1 to 0.24mg P L-1 through the wetland system. Despite seasonal variation
noted for many study parameters (e.g. BOD, COD, TOC and TN) no clear variation was noted with TP.
The wetland removed on average 77% of P from the influent over the study period. Adsorption to the
substrate constituted the predominant removal process (24 mg m-2 d-1) of which 8.2 mg m-2 d-1 was
assimilated by the plants. The wetland purification rate for P, even in winter, was cited as 0.023 g m-2 d-1.
China has practised sewage irrigation since the late 1960’s over millions of hectares and as a result crops,
soils and groundwater have been polluted to varying degrees (Xianfa and Chuncai 1995). Since the mid-
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1980s many resources have been utilised in upgrading sewage land treatment systems with agricultural
application sewage. The predominant wetland system employed in northern China appears to be the reed
bed system but some of the examples cited in this paper refer to a combination of rapid infiltration, slow
infiltration, overland flow and surface flow wetlands depending on the resources available. The three
examples of sewage treatment wetlands and one of a papermill effluent treatment wetland are of reed bed
designed to treat approximately 120 to 1800m3 d-1 of municipal sewage (with preliminary settling, a
hydraulic load of 1.0 to 6.0 cm d-1 and a retention time of between 1.5 to 10 days) and showed a P
removal capacity of 55 to 86%. The findings of studies of these wetland systems have allowed Chinese
researchers to develop an empirical model for the design of wetlands. The parameters stipulate a
retention time of 7 to 10 days, a hydraulic load of 2 to 20cm d-1, a water depth of <10cm in summer and
>30cm in winter, an organic load of 15 to 120kg BOD ha-1 d-1 (less for non-preliminary treated sewage), a
rectangular shape with a length to breadth ratio >10 with various water distribution facilities and
vegetation dominated by Typha and Phragmites spp. A point worth noting is that most of these wetlands
are followed by an aquatic plant pond and fish ponds which have an enormous economic value in that
they provide water in water-poor regions as well as fish, waterfowl and other food sources. The effluent
quality from these wetlands was found to meet the national standards of fish culture water quality
guidelines. The authors’ concern is the problem of availability of area required to implement these
systems, although they argue that such integrated systems would prove to be economical.
Yang et al. (1995) reported on a four year research program analysing the nutrient and waste removal
efficiency of a large (8400m2) wetland in Bainikeng, Shenzen, a subtropical region of China, receiving
3100m3 d-1 of municipal wastewater. The constructed wetland was a subsurface flow system with a series
of Phragmites communis beds (loading rate 95.4cm d-1, gravel 80cm deep and 1512m2), attached to two
beds separately containing Phragmites communis and Cyperus malaccensis (gravel 100cm deep and
1739m2) which were then attached to three parallel oxygenation ponds (water depth 150cm and 1710m2)
containing Nelumbo nutifera and Eichhornia crassipes (which drain finally into two gravel Cyperus
malaccensis and one gravel Lepironia articata bed (loading rate 100.7cm d-1, gravel 100cm deep and
2850m2). The removal capacity for P within the subsurface beds was 30.6% for the study period and
relatively poor when compared with BOD, COD and TSS removal rates, but this increased to 41% after
the oxygenation ponds although there was a marked improvement in wetland removal capacity over the
study period (1991, 20% to 1993, 39%). It was concluded that the vegetated beds were therefore quite
important in removing the gross pollutants, while the ponds were equally important in nutrient removal.
Artificial wastewater with a concentration of up to 40mg P/L was passed through a 10 x 70cm column of
mangrove soil from Hong Kong containing 0.028% P (and 10.3mg L-1 FRP) and found to produce an
effluent P concentration of less than 0.2mg L-1 consistently over a 55 day period (Tam and Wong 1994)
for a variety of wastewater concentrations (10 to 40mg P L-1). Wastewater (100mL) was applied daily
(equivalent to 1.3cm d-1) at a rate of 40 to 50mL h-1 and the column conditions mimicked tidal inundation,
i.e. an eight hour wet/dry cycle with alternate seawater/wastewater applications. FRP was found to be
adsorbed quickly within the first 4cm of the column for each treatment.
In a comparative study of domestic raw wastewater treatment in Thailand between a facultative pond and
a water spinach pond (Ipomea aquatica) it was found that hydraulic retention times of greater than eight
days were required to achieve any significant P removal (Ipomea; 8 days 32.9%, 16 days 90%: facultative
pond; 20.1 and 21.4% respectively) with an influent P concentration averaging 1.9mg L-1
(Karnchanawong and Sanjitt 1995). This contrasted with the facultative ponds ability to remove nitrogen
which was much better than that by the spinach ponds. The Ipomea tended to die off quickly after high
TSS inflow due to a clogging of the root biomass and it was suggested that harvesting would be necessary
to ensure continued growth, P uptake and to reduce the potential for accumulating P in the bottom of the
ponds.
There is an increasing environmental problem in India which generates an estimated 8642 x 106m3 of
wastewater per annum of which only 23% is treated at primary level (Juwarkar et al. 1995). The
remaining 77% is apparently discharged directly into surface waters or onto the land. Results of a simple
study of a constructed wetland in India at Bhubaneshwar consisting of the macrophytes Phragmites carca
and Typha latifolia (cement pipes having 0.1256m2 and 0.8m deep filled with 30% soil and 70% sand)
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found that between 23 and 48% P was removed from an application rate of 5cm d-1 with a mean inflow P
concentration of 14.9mg P L-1. The Phragmites was the preferred plant due to its ability to establish
quickly and grow more profusely.
New Zealand
Wetland research has been conducted in New Zealand over the past few years with an emphasis on
treating rural runoff such as dairy farm wastewaters. Tanner (1994) utilised wetlands at the Ruakura
Research Farm in Hamilton, to assess the performance of four horizontal channels (9.5 x 2m filled to
0.4m with pre-washed alluvial rhyolitic gravel 10-30mm diameter, 36% effective pore space) and two
upflow pilot-scale (1.5m diameter filled to 1m with the same substrate) wetlands planted with
Schoenoplectus validus over a twenty month period. The removal of TP was correlated to retention time
with the horizontal flow beds showing better results than those produced by the upflow systems,
particularly when the loading rates were increased due to the longer holding times and wastewater
production during the spring/summer periods. The TP retention varied markedly through both flow
systems (-50 to 90% reduction), except in the longer retention period of the horizontal flow (7 days) but
showed a gradual decline in effectiveness from 90 to 40% reduction capacity during the study period. It
is important to note that a comparison between the two systems was difficult to make due to the
difference in substrate contact volume and relative retention periods, but when these factors were
corrected for, the results from the upflow system were similar to some of the tested parameters under
horizontal flow conditions, i.e. TSS and BOD.
Tanner et al. (1995) reported on the removal capacity of alluvial rhyolitic gravel wetlands (19m2) planted
with Schoenoplectus validus over varying retention times for the same wetland facility at Hamilton
mentioned above. The average TP inflow concentrations were 11.2g m-3 during the study period and
resulted in mean loading rates between 0.17 to 0.8g m-2 d-1 of which the FRP comprised 80% of the
influent. The rate of P mass removal over the twenty month study period increased from 36 to 74% in the
planted wetland as the retention time increased from 2 to 7 days which exceeded the retention capacity for
the unplanted wetlands (12 to 37% respectively). The mean annual removal rates of the planted wetlands
were 0.13 to 0.32g m-2 d-1 which increased gradually with the mass loading rates. The net storage by the
plants accounted for 3 to 60% of the P removal.
A natural wetland system in Waitangi forest, Northland, New Zealand, receives sewage effluent after
primary treatment in an oxidation pond in the resort town of Paihia which had an approximate population
2000 in winter rising to 8000 in mid-summer. Studies of the character and effectiveness of the wetlands
are found in Cooke 1992, Cooke et al. 1992, Cooke 1994, and Nguyen et al. 1995. The wetland area
subjected to the sewage inflow was a small proportion of the total area of the natural wetland and as such
was subject to different hydrologic conditions and dominating macrophytes (permanently flooded with
Typha orientalis and Eleocharis sphacelata compared to the natural wetland areas which have a summer
drawdown and were dominated by Baumea-Isachne). The chemistry of the two wetland waters was
distinctly different as the natural wetland waters were more pristine and oligotrophic, and had a lower pH.
Mass transport studies reported that 30 to 70% of the influent P was removed from inflow waters which
were highly loaded with P (approximately 34g P m-2 y-1). Phosphorus was shown to decrease markedly
after the sewage discharge waters encountered the natural wetland flow primarily because of the iron
content in the flow (present as a soluble complex or colloidal state with humic material). The iron
adsorbed and/or flocced out the P as a Fe-P-humic complex due to the increase in pH, and because the
sewage-impacted wetland waters were mixed with those originating from a much larger area of natural
wetland. As the iron input was relatively unlimited it implied that the P removal was sustainable. The
authors suggested that if P removal was the main objective of a constructed or natural wetland treatment
system, then a high loading to a small area which also receives drainage from mineralised natural
wetlands may be a more cost-effective and sustainable design criterion than the more general approach of
lightly loading a large, hydrologically-isolated wetland system with a finite capacity to retain P. The
potential for release of P is a constant problem to be considered regardless of wetland design and
ultimately relies on factors such as the redox potential of the iron compounds, the concentration of P in
the surface sediment and upon the sustainability of iron in the natural wetland waters.
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6.3
Critical Evaluation of the Reviewed Journal Articles and Proceedings Papers
During the course of the review process it was sometimes difficult to interpret data presented in the
articles, particularly proceedings papers. There are often a large number of variables that are not defined
which makes it very difficult to draw conclusions. Some examples of these oversights are as follows :
• Wetland substrate composition, both chemical and physical, were often not fully described;
• Macrophyte densities were often not reported and turnover rates were seldom measured.
Standing stock differences between climatic regimes can be difficult to interpret and can be
enormous, e.g. in some cases established wetlands containing Phragmites can vary from 600g
m-2 to 15000g m-2 (see Table A.1 in the Appendix). These differences may at least be partly
due to climate but are more likely the result of other undefined variables. Also in some
instances it was not clear whether estimates referred to above ground or whole plant biomass;
• Productivity rates also seemed to vary by an order of magnitude for given species of
macrophytes between various studies (see Table A.1 in the Appendix) so it is not surprising
that there seems to be so much conflict regarding on the role of macrophytes. There was only
a few occasions where periphyton, plankton and microbial assimilation were taken into
account when determining the total P assimilation;
• Seasonal growth varies enormously in temperate climates and reported growth rates in these
conditions seemed to adequately address the effect of season upon growth rate. However,
there is little data relating to the role season plays in tropical conditions, i.e. what effect does
extended dry periods or extended wet periods actually have on macrophytes in tropical
climatic conditions?;
• It was not always clearly defined whether P values were referring to concentration or mass;
and
• The P composition of effluent was often not sufficiently described to discriminate potential
differences between, for example, agricultural runoff from different soils and cropping
methods, stormwater, and primary and secondary treated sewerage effluent produced by
different types of aerobic and anaerobic treatment processes. This refers to differences in the
relative proportions of various chemical forms of P such as organic particulate, dissolved
organic, dissolved inorganic, particulate (sorbed) inorganic, etc. The composition of P is
expected to be quite variable and therefore the requirements for final treatment and removal
by wetland processes are most likely to vary accordingly.
As a consequence, it is very difficult to produce a simple optimum design for wetlands which will
effectively remove P and does not waste resources by incorporating unnecessary steps or excessive land
area. Given the available data and the increasing research into wetlands (especially in the tropics) it
should become easier for managers to implement designs based on a knowledge of general biological and
chemical principles.
7.
CONSTRUCTED
PHOSPHORUS
WETLAND
CONSIDERATIONS
FOR
THE
REMOVAL
OF
The P removal capabilities of free surface water emergent macrophyte bed wetlands appears to be
somewhat unreliable and is probably influenced by subtle variables that have not yet been fully taken into
consideration. It is not possible to discount the possibility that this situation will be improved in the near
future but at this time it appears that this type of wetland alone is not the best design for P removal. High
diversity natural wetlands appear to be capable of achieving net P reductions and it is possible that with
improved understanding of the precise mechanisms involved, constructed wetlands that function similarly
may soon be developed. However, at this time the most consistent removal success is being
accomplished using alternative or combined approaches and some of these are described below. The
following considerations are made on the assumption that discharge data has been fully established such
as the hydraulic potential, P loading, available land area, target effluent quality, etc.
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Taking into account the variability of results described in the literature, success in removing or
assimilating P can be accomplished by utilising several different processes. The fundamental
requirements for achieving success in specific wetland construction for this purpose is a complete
understanding of the wetland function, allowing for a suitable establishment time and allowing for the
‘self-designing’ capacity of nature (Mitsch and Wilson 1996). Phosphorus can be removed from both the
water column and sediments, particularly during the growing season, and on a short term basis, when it
can be taken up and stored by wetland vegetation. Interactions within the sediment occur slowly in
wetland systems, and P can be readily immobilised by calcium, aluminium and iron by adsorption and
precipitation reactions. Fine mineral soils usually have higher concentrations of these ions and therefore
have a greater capacity to retain P than organic soils.
Options for Wetlands
When considering the design of the wetland whose primary function is to retain and transform nutrients,
the wetland must be capable of physically detaining the nutrients (Marble 1992). This is best achieved by
regulating the inflowing water velocity so that sediments and their adsorbed nutrients settle out.
Various types of wetlands and pre-wetland systems may be proposed for the effective removal of P and
each are considered below. In all cases, the purpose of the wetland needs to be taken into account as well
as the characteristics of the wastewater influent (type, P loading potential and volume). Several wetland
systems that can be considered are following :
Gross Pollutant Traps
This is a very simple component that can be applied immediately before a wetland system or incorporated
into the preliminary stage of a wetland. They can be particularly useful if the influent is high in
particulate matter (inorganic or high in BOD). The traps reduce flow rates and this subsequently allows
time for particulate matter to settle out. Suspended matter or particles are often associated with
contaminants, such as trace metals and phosphorus, so by taking steps to reduce the concentration of
suspended solids from the inflow, a proportion of the contaminant load should be readily extracted. The
only draw back is that the traps may require regular maintenance to clean out the settled particulate
matter, and any resulting algal growth, so that they maintain effectiveness and do not promote
contaminant pulses, particularly if the traps are deep enough to promote stratification during hot
conditions.
The design has to be deep enough to reduce flow and promote settlement, and the margins protected or
baffled to prevent wind induced turbulence which may re-suspend some of the finer material. Inflow
should be regulated to minimise the effects of occasional flood or high volume discharges. As mentioned
earlier, stratification could be potentially problematic, especially in tropical conditions, so shading the
traps may reduce surface temperature increases as well as reduce the potential for algal growth. Shading
will, however, increase the likelihood of anaerobic conditions which would potentially produce
malodorous conditions (unless floating macrophytes were used) so a suitable alternative would be to
introduce the influent at various points along the bottom of the trap enhancing circulation. Sedimentation
would then be restricted to various regions of downwelling.
Clear Water Lagoons
Clear water lagoon incorporating rooted submerged macrophytes should eliminate a major proportion of
FRP from inflowing waters. Apart from the focus of direct plant assimilation of FRP, the additional
chemical precipitation of P is expected due to photosynthetic activity of the submerged macrophytes and
other associated vegetation (i.e. phytoplankton, algae and periphyton). This process will reduce the level
of dissolved carbon dioxide in the water raising the pH levels. Phytoplankton, periphyton and some
submerged macrophytes, can utilise bicarbonate as an alternate source of inorganic carbon at these higher
pH conditions thereby maintaining the effectiveness of P reduction in these lagoons. The high pH levels
associated with high concentrations of calcium (which may be naturally high) will also precipitate P as
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calcium-phosphate minerals. Dosing application of calcium and aluminium based compounds can be
considered as an additive option in waters naturally low in these minerals to precipitate residual FRP.
Investigations of the potential for utilising clear water lagoons with submerged macrophytes in any
climatic condition have been limited to date. Gumbricht (1993), however, has suggested that the use of
these systems should be more applicable for use as a tertiary treatment step, ‘polishing’ effluent or
treating eutrophied natural waters, as pre-treated waters will be less likely to impact upon a more sensitive
system (i.e. influent with low turbidity and nutrient loads). Submersed macrophytes species such as
Cladophera spp. Enteromorpha spp., Potamogeton spp., Ceratophyllum spp. Myriophyllum spp., Elodea
spp. and Egeria sp. are commonly used species and have good potential for P uptake (see Table A.1 in the
Appendix). Values for P uptake by these species range from 5.9mg P m-2 d-1 (Ceratophyllum) to 410mg P
m-2 d-1 (Egeria densa) and all have the potential to generate large biomass. Productivity rates are likely to
increase markedly in warmer tropical conditions so P removal potential is likely to be very significant.
It is envisaged that maintenance of these lagoon systems should be minimal once they are fully
established and providing influent rates and P loading (as well as the loading of other nutrients) are
constant. The ideal use of this system would be to locate it downstream of more rigorous treatment
systems so that the predominant form of P is FRP which is present in concentrations between 0.1 and 1mg
P L-1. Locating the lagoon downstream will also reduce the potential for problematic inflows if any of the
previous systems fail causing loading rates to dramatically increase for any period of time. In the case of
treating secondary effluent from a sewage treatment facility, dosing with a flocculation agent to
precipitate the majority of FRP (FRP should be the predominant form of P), especially during the
establishment stages of the lagoon should be effective in stripping P from the inflow. In warmer climates
it is possible that invading species will be a problem (both submersed and free floating species). Free
floating species can be very productive and have the capacity to shade submersed species, and any other
associated periphyton and phytoplankton, which will reduce the effectiveness of the lagoon, particularly
if the lagoon is well established. Depending on the design of the lagoon, harvesting of the free floating
species should alleviate this problem. Invading submersed species will be more difficult to eliminate but
monitoring the system should allow managers to determine if invading species are supplementing P
reduction.
The occurrence of algal blooms is a potential problem particularly if the treatment system is relying on
regulated inflow. Spikes of high nutrient inflow, due to a system malfunction above the lagoon, is a
possibility and this is most likely to lead to blooms of cyanobacteria. Regular monitoring the lagoon
system and inflow quality will assist in the prediction of high concentration pulses which should allow
managers to shut down the inflow for a period giving the lagoon time to assimilate any increase in load or
to flush out the lagoon with an alternate water supply diluting the algal concentrations and nutrient spike.
It is recommended that prior to use the lagoon is well established to allow the manager to make these
decisions without any detrimental effect on the lagoon system.
Banked Grassland Overflow
Controlled flow over a banked grassland is a possibility particularly in constant influent treatment
systems to further reduce flow and utilise grassland/sediment uptake of dissolved nutrients. This
approach was termed a grass filter strip by Higgins et al. (1993) and proved very effective. Flow can be
regulated to allow for substrate drying so that particulate and organic P fractions present in the
wastewater can be mineralised and therefore become available to the grasses. This can be achieved by
having a valve switch at a constant inflow point to split flow to two grassland systems so that as flow to
one is shut down for the drying cycle, flow to the other begins for the wetting cycle. Alternatively inflow
can be cyclical so that inflow ceases and resumes after set periods of time.
The inflow is directed across a ‘level lip spreader’ which is essentially a narrow trench filled with crushed
rock the purpose of which is to distribute the wastewater evenly over the width of the grassland bank
which is set on a 6° down angle. This form of application essentially reduces channelisation and the
erosion of the bank causing a short circuiting of the overland flow. Higgins et al. (1993) utilised drainage
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ports underneath the filter strips to promote infiltration of the overflow thereby improving the efficiency
of nutrient uptake.
The effectiveness of this system will be based on the size of the filter strip, the rate of inflow and loading
concentrations, the type of grass selected and the character of the substrate selected. Maintenance should
be minimal providing invading grass species are monitored. Periodic harvesting by ‘mowing or bailing’
the strip is recommended (or if the strip is large enough periodic grazing by cattle, etc.) and should be
effective at maintaining a monoculture. The selection of grasses would need to take into account a
tolerance of wetting and drying cycles and ideally a disposition to ponded pasture conditions.
Channelisation is a potential problem with this system particularly in large run off events. The placement
of the inlet relative to the location of the lip spreader and grassed bank and inflow control will be
essential to minimise the potential for channelisation and ensure sheet flow which is critical for adequate
performance of a grass filter (Higgins et al. 1993).
Horizontal Subsurface Flow Systems
A wetland system allowing for a sub-surface flow input to the wetland dominated by Typha or
Phragmites species is recommended mainly due to the potential for mean annual production and standing
stock (see Table A.1 in the Appendix). Whenever possible, the wetland substrate can be amended to
contain a grain size of predominantly 10 to 30mm with well mixed additives to the gravel base, such as
sandy loam or coal fly ash, to ensure adequate and constant sub-surface flow ensuring stable hydraulic
conductivity and the supplementary adsorption of P. An adequate depth of substrate should be provided
(nominally 0.5m) to promote root and rhizome growth and sufficient substrate/water contact.
It is suggested that the width and length of the wetland is designed to accommodate the intended inflow
volume and loading of the wastewater, and if necessary allow for harvesting. In the warmer climates high
production rates of Typha and Phragmites should prevent any colonisation by invading species but a
drawback to the higher production, and depending on the level of nutrient loading, is that these species
could become very dense over a short period of time potentially causing flow barriers throughout the
wetland. The selection of this form of channelled wetland over surface flow systems, and these particular
emergent macrophyte species, for effective phosphorus removal is due to their potential for large above
and below ground biomass which ensures adequate phosphorus storage and sufficient below ground
surface area for bacterial growth (also a filter for solids) and oxidised micro-environments to promote
organic matter decomposition. These plants, particularly Typha spp., are also hardy and the nutrient
content in their senescing leaves are readily translocated meaning that the litter will not be a significant
source of nutrients (Greenway pers. comm.).
Cyperus involucratus may also be chosen as a suitable emergent macrophyte because it grows in dryland
and flooded conditions (Hocking 1985) and therefore may be useful where inflow is erratic. Its root
system is not as aggressive as Typha or Phragmites and this may make it suitable for planting in
channelled systems which do not have an extensive substrate depth.
Maintenance of this particular horizontal flow system is not expected to be high once the wetland is
established. High productivity rates may require that harvesting is considered and this may improve P
removal efficiency. However, in these systems where there is a significant soil/sediment component it is
envisaged that harvesting the macrophytes may reduce aerenchymatic oxygen transport and microbial
activity and decrease soil conductivity by inducing sediment compaction (Gumbricht 1993). It is
therefore suggested that for these wetland systems harvesting is not to be considered a significant
requirement for maximum P removal.
Plant Composition
The plant composition of a wetland designed to assimilate and retain as much P as possible will be a plant
or combination of plants that are able to grow quickly and store P efficiently in the region where the
wetland is required. Wetland plants function in several ways to reduce P within the water column or
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sediment. The stems and leaves of emergent and multi-stemmed plants offer resistance to inflow
promoting further settling. Seasonal assimilation from the sediment and water column by biological
growth processes also occur which is served best by emergent and aquatic bed vegetation.
In Queensland, plants that are capable of best achieving the objective over the majority of climatic
conditions are Typha, Schoenoplectus, Triglochin or Phragmites spp. and are used fairly commonly. In
the wetland systems described above, the submergent macrophytes for the clear water lagoon could be
Myriophyllum spp., Ceratophyllum demersum, Najas tenuifolia, Hydrilla verticillata, Vallisneria spp.,
Potamogeton spp., Egeria densa or Elodea spp. The introduction of Melalueca quinquinervia around the
outer portions of the wetland may also serve to assimilate nutrients and maintain subsurface water uptake
thereby minimising the potential for impacts on the localised groundwater.
Overall, the selection of macrophyte species should be done on a local basis only, and following a local
survey. Some species that may be considered ideal may easily become weeds and difficult to control (see
Mockeridge 1995). This includes Typha and Cyperus involucratus as they have the propensity to become
nuisance species, as noted in irrigation channels and drains, and within natural wetlands (G. Lukacs pers.
comm.). The submerged macrophyte species will require special selection, and careful management, as
some are considered noxious weeds (e.g. Elodea spp.).
An important consideration when selecting emergent macrophytes is to determine whether the plant or
plants have the capability to translocate nutrients from senescing fronds or aboveground biomass. Bald
(pers. comm.) stated that Triglochin procerum retains nutrients in the senescing fronds which are then
released back into the water column after decomposition. Harvesting the senescing material, or
harvesting the plant on a seasonal basis to prevent wide scale die-off, is then an important consideration
in macrophyte selection to maintain year round low water column nutrient concentrations but is difficult
to carry this out without disturbing the bottom sediment.
Planting Patterns
Whole submersed plants should be translocated to the clear water lagoon system and it is expected that
the establishment of the macrophytes will take a period of time depending on the season, and the quality
of water they are planted in. The rooted submerged macrophytes could be planted initially with 0.5m2 and
monitored throughout their development.
The planting pattern of emergent macrophytes within the wetland systems described above will be such
that water contact with the substrate and hence root biomass will be constant. Therefore an establishment
period of at least three to four months is necessary for the plants which will need to be planted initially
within every 0.5 m2 along extensive stands completely spanning the wetland to achieve the maximum
growth potential. The planting pattern will vary depending on the climatic regime but generally, the more
dense the vegetation, the greater the ability to remove and assimilate phosphorus. As a rule of thumb,
Marble (1992) stated that for emergent macrophytes the water depth should not exceed 50% of the plant
height.
There is potential to increase the vegetation diversity by planting different forms of macrophytes within
the lagoon and the wetland to ensure the greatest potential for nutrient cycling processes.
Retention Time
The longer the retention time, the better a wetland will be in reducing the P load, providing P uptake is
working efficiently.
The retention time for the clear water lagoon could be as low as one or two days especially for lagoons
that will be receiving influent that has been pre-treated with a flocculating agent to reduce the P load (to
less than 1 mg P L-1). This will be entirely dependant on high standing stocks of submersed macrophytes
and clear water, the chemistry of the lagoon water and influent loading rates.
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A retention time of two weeks is recommended for the sub-surface flow wetland as the minimum
requirement for total P removal from the wastewater, the actual timing varying with wetland size. Crites
(1994) recommended that significant P removal will require a long detention time (15 to 25 days) and will
be enhanced with lower P loading rates (<0.7g P m2 d-1).
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Wetland Depths
It is suggested that the depth of the gross pollutant trap would be 1m to allow for immediate settling of
coarse organic and inorganic matter.
The water flow over the grasslands would not be expected to exceed 1cm depth at any one time and
regulation of storm flow would be expected to maintain the sheet flow conditions. The level lip spreader
would act as a settlement trap to some degree and could effectively operate at approximately 0.5m deep.
The depth of the submerged macrophyte pond should be 2m providing an extensive retention time and a
large medium for macrophyte growth, including plankton and periphyton communities. Fish and other
aquatic wildlife would be introduced to the pond but depending on their density may in fact become a
significant source of nutrients.
The horizontal sub-surface flow wetland water depth would be expected to be a maximum of 0.5m. This
shallow water depth would increase frictional resistance and reduce water flow velocity. The dense
annual macrophytic vegetation would be optimal to assist flow resistance.
The lagoon and wetland area are systems that would remain constantly saturated. As mentioned earlier
the grassed filter strip is designed to be split into two systems so that inflow to each grassland/substrate
can be alternated by a flow switch control. This would allow cycling between wet and dry hydroperiods
which are important to mineralise organic forms of P.
Wetting and Drying Cycles
Wetting and drying cycles would be utilised for the grassland filter strip treatment (although it could be a
possible design option for the horizontal flow wetland) to enhance mineralisation of P fractions
immobilised by the substrate for plant uptake, and to allow time for harvesting or grazing, if necessary.
Willett (1982) has shown that flooding and drying has a strong effect on increasing the soil sorption
capacity by immobilising added P, especially when soil was rich in organic carbon and reducible iron.
The cycle period is expected to be on a two day rotational basis which is within the minimum flooding
period suggested by Willet (1982), but is dependant on grass and soil type, and climate. A potential
drawback for this type of treatment is the gradual reduction in the availability of soil P for uptake by the
grasses, particularly if there is a finite amount of reducible iron in the substrate.
Wetland Shape
The shape of the horizontal flow wetland has been traditionally rectangular with a length to breadth ratio
of approximately 3 or 4 to 1 and a slope of around 2°. Variations to this rectangular shape are sometimes
necessary due to land availabity and particularly to avoid channelled shortcircuiting flow ensuring water
contact with all of the wetland. It is also recommended that the horizontal flow wetland has a varied
depth profile with shallow and deep portions throughout the course of the wetland allowing the water to
spread over a wide area and mimic as close as possible natural wetland/marsh structure. The size of the
wetland should reflect the nutrient loading of the water entering the system. The multi-depth profile will
be provided with vegetative barriers and will assist in reducing wind effects and resuspension problems.
The gradient of the inflow prior to wetland entry should be gradual to allow for slow inundation of
inflowing waters.
The outlet, or series of outlets, at the end of the wetland would be constricted to promote wetland
saturation and therefore increased potential for sedimentation, adsorption, biological transformation and
nutrient retention.
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Substrate Types
The substrate best suited for P removal is primarily alluvial, alfisols, ferric, clay or other fine soils. If the
natural substrate is to be used, it should be comprehensively analysed for soil type, pH and P retention
capacity prior to the wetland construction. It is important that the P retention capacity is determined
under both oxic and anoxic conditions because of the effects of redox potential on P sorption. Where the
P adsorbing capacity is found to fall below 80% (remembering that the grassland and wetland substrate
will be mixed with a gravel base to improve hydraulic conductivity), substrate amendment with a material
capable of adequately adsorbing P should be considered. Marble (1992) suggested that where possible
fine mineral sediments or soils containing high levels of aluminium and iron should be utilised. The long
term P removal capacity of acidic wetland soils would be directly related to the extractable content of
aluminium .
The depth of the substrate should be at least 0.5m.
The substrate should be placed within the wetland area through general excavation of the terrain to the
shape required. A sealed surface should then be put in place to prevent groundwater seepage into and
wetland water out of the wetland.
Flooding
Regulated flow will reduce the potential for flooding through the wetland but in the event of increased
inflow (i.e. during the instance of plant mechanical breakdown or heavy storm rainfall) it is envisaged
that the design of the wetland should allow for periodic inundations of the system.
The gentle slope grasslands may be affected by storm inflow with the development of channels within the
slopes. If the grass is kept densely vegetated, flow short-circuiting should be avoided.
Necessity for Harvesting
Harvesting should not be considered a necessity unless emergent or floating macrophyte invasion of the
pond is apparent or the macrophyte community within the wetland is so dense that wetland flow is
impeded.
The grassland slope will need to be harvested (by mowing, bailing or grazing) and this should be
conducted prior to, and throughout, the growing season when necessary.
8.
MANAGEMENT RECOMMENDATIONS
The main management issue is the monitoring of inflow volume to the wetland treatment system.
Constant flow would be required to the pond or clear water lagoon which has to be maintained throughout
the system. Diurnal and hydraulic loads will, however, vary cyclically so a balancing storage would be
needed to achieve this. A balancing storage may not be feasible for a variety of reasons (e.g. limited
space, etc.) so the clear water lagoon may have to be designed so that the lagoon could accommodate
expected cyclic variations. Discharge from the source would also be expected to vary in contaminant
concentration during the year for each STP facility but it would be anticipated that the system should be
able to deal with gradual variations of up to 100% for nitrogen and phosphorus loadings.
The monitoring of water quality of wastewater inflow and the outflows of each of the treatment stages
would be necessary and an adaptation of the monitoring schedule described by Tchobanoglous (1993) is
recommended. The schedule is detailed in the Table 1.
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Table 1 Summary of suggested monitoring parameters for constructed wetlands
Parameter
Water Quality
Dissolved Oxygen
Diurnal Dissolved Oxygen
Temperature
Conductivity
pH
BOD
TSS
Particle Size Distribution
Nutrients
Chlorophyll a
Metals
Bacteria (total and faecal
coliforms)
Trace Organics
Biotoxicity
Sediments
Redox Potential
Salinity
pH
Nutrients
Wetland Phase
(Pre- or During
Construction, or
Ongoing)
Location
(In, Out, Along the
profile or Selected
locations)
Sampling Frequency
(Weekly, Monthly,
Quarterly, Semi-annually
or Annually)
O
O
P,D,O
P,D,O
P,D,O
P,D,O
P,D,O
P,D,O
P,D,O
O
P,D,O
P,D,O
I,O,A
S
I,O,A
I,O
I,O
I,O,A
I,O,A
I,O,A
I,O,A
S
I,O,A
I,O
W
Q1
W
W
W
W
W
W
W
W
Q
M
P,D,O
P,D,O
I,O,A
I,O
S
S
P,D,O
P,D
P,D,O
P,D,O
I,O,A
I,O,A
I,O,A
I,O,A
Q
Q
Q
Q
O
O
O
P,D,O
P,D,O
S
S
S
S
S
Q
Q
Q
Q
Q
O
O
O
I,O,A
S
S
C
A
S
Biota
Plankton
Invertebrates
Fish
Birds
Mosquitoes
Wetland Development
Flow Rate
Flow Rate Distribution
Water Surface Elevations
Note :
1
- The frequency of these measurements will vary according to climatic regime.
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9.
CONCLUSIONS
Wetlands are sensitive to nutrient inputs and, although nitrogen is normally lost from the system,
phosphorus inputs usually accumulate within the system. Phosphorus enriched waters interact with the
vegetation, aquatic biota and wetland substrate. The assimilation of phosphorus into aquatic vegetation is
usually short term and decomposition of the detrital matter is usually rapid which may release the P back
to the water column. The undecomposed organic P within the litter zone is an integral part of the
substrate P pool. Depending on the physico-chemical conditions within the wetland, it may act as a sink
or a source which always questions the potential of a wetland to retain P. The pre-requisite of producing
a wetland which is effective at removing P from inflowing water is an understanding of the substratewater P interaction and the factors regulating P retention and or release. There is clear evidence from the
review that high diversity natural wetlands are capable of effecting long term removal of water column P
(albeit that loading rates are relatively low and residence times are high) but limited diversity constructed
wetlands, particularly emergent macrophyte monocultures, can prove to be generally unpredictable and/or
unreliable. It is therefore envisaged that the system that mimics the natural wetlands systems (with some
modification) to the best degree will find the most efficient P removing capability.
Phosphorus adsorption is governed by the availability of sorption sites within the substrate and by the
presence of amorphous and crystalline forms of iron, aluminium and organic matter. Aluminium is not
involved in oxidation-reduction reactions and is therefore not affected by oxic or anoxic conditions but
rather is regulated by pH and the presence of organic matter. However, there is a often a strong
correlation between redox potental and pH in sediments it is likely that P reactions with aluminium may
occur indirectly. Phosphorus interactions with iron can be significant at the substrate-water interface
where ferric ions in the oxidised zone can function as a sink for phosphorus diffusing from the water
column and from the underlying anaerobic substrate layers. During water flow, suspended particles can
sequester available phosphorus from the water column and, because of their large surface area and
concentrations of iron and aluminium oxides, will have a high capacity for sorption of FRP.
Wetland substrates can be modified if not already capable of retaining phosphorus. Their capacity for
retention can also be enhanced by the process of substrate accretion due to the mineral particle deposition
during stream flow and the incorporation into the organic pool.
The design of a wetland system which will efficiently remove P from inflowing wastewater is dependent
on a number of factors such as the consistency of wastewater inflow and the effectiveness of the wetlands
themselves, which rely on careful management. The lagoon system is the process that will require the
greatest management time but, once established, should be self sustaining.
The recommended approach to wetland design for effective P removal is to ensure an adequate
development time, and to gradually introduce contaminant loading. The design considerations presented
in this review are based on general biological principals that would be best suited to remove P from
wastewaters but will require some trialing and review before final design criteria can be developed for
each individual case. A combination of these systems will be more expensive than the simple single
channel design system that has been incorporated into treatment plants to date, but the overall principles
should be applicable for a variety of different climatic conditions. An advantage of selecting
macrophytes which are native to a region are that they should be naturally productive within the
constructed wetland, capable of storing P, and therefore more beneficial for nutrient uptake.
It is very difficult to design a simple system that will optimally assimilate P from any given source of
wastewater for a variety of different climatic conditions which exist throughout Queensland. Not enough
information is currently available to accurately predict the effectiveness of some of the options considered
in this review but they at least provide a starting point for researchers to develop design criteria for each
region and each wastewater type. Research is beginning to move in this direction, particularly with
respect to Queensland and its tropical areas (by the Queensland Department of Natural Resources, James
Cook University of North Queensland, Griffith University and the University of Queensland) so
gradually site specific practices should be developed ensuring that where required P is effectively
removed form wastewaters prior to their reuse or discharge.
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On a cost return basis it is hard to quantify whether P removal by wetland design will be a more realistic
option as opposed to chemical precipitation, or pollutant filtration or trapping of source wastewaters,
particularly in the long term. However, the literature supports the concept of successful P removal, and
the potential for increased effectiveness of P assimilation by macrophytes and substrate in warmer
climates, suggest that wetland processes are a valid alternative for P treatment. The best option is a
configuration which takes into account several systems in line each having a different process of P
removal or assimilation, much like the multi-stage approach alluded to by Brix (1993). More
importantly, however, and repeating what has been mentioned before, is that the approach required for
any wetland design is going to be based on the source of wastewater, the location of the wetland and the
rate and loading which is then applied to the system.
Buchberger and Shaw (1995) support the application of model based design approaches to determine the
ideal construction criteria for wetlands for wastewater treatment. This scheme should be of some
assistance as it can be used to synthesise sub-models which can adequately describe variable loadings,
atmospheric moisture and energy fluxes, contaminant fate and transport, and effluent release and recycle
rules. As an understanding of fundamental wetland processes, particularly nutrient transformation and
transport mechanisms improves, the model can be updated. The flexibility that a wetland model can offer
is that the wetland does not have to be built to compare its effectiveness due to different release and
recycle loads, and to determine how a wetland’s performance is affected by different treatment
mechanisms, wetland operating depths and cell aspect ratios. It is doubtful that P transformation
processes are sufficieently well understood for accurate models to be developed at this time but the
collection of the emperical data that would be needed is a worthwhile goal. As our knowledge of the fate
of P improves, Australian data can be compiled for each wetland system relevant to climatic, substrate,
macrophyte type and diversity, etc., which should assist in the design of the most efficient wetland system
for P removal.
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Australian Centre for Tropical Freshwater Research
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APPENDIX
Table A.1
A list showing commonly used plants in international constructed wetlands and plants considered suitable for the use in Constructed Wetlands in Queensland used for the
treatment of waste water (plant list adapted from updated Appendix A of the Interim Guidelines on Planning, Design and Management of Artificial Wetlands in Queensland, DPI
(1995)).
SCIENTIFIC NAME
COMMON NAME,
HABIT
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
0.08 (Shardendu and Ambasht 1991)
6 - 27, mean 12 (Shardendu and Ambasht 1991)
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
PHOSPHORUS CONTENT
mg P /g dry wt
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
SUBMERGED PLANTS
Aponogeton natans*
Blyxa spp.
Ceratophyllum spp.
C. demersum
Chara spp.
Egeria densa
Coontail
Hornwort
Stonewort
Dense waterweed
Najas tenuifolia
Nitella spp.
Potamogeton perfioliatus
P. pectinatus
P. crispus
P. scweinfurthii*
Triglochin procera
Utricularia gibba subsp.
exoleta
Vallisneria gigantea
Waterweed
Eriocaulon
Hydrilla
Water nymph
Stonewort
Clasped pondweed
Sago pondweed
Curly pondweed
5.9 (in Brix 1994)
6
13.3, max 14 (Greenway in press)
2.5 (in DeBusk and Ryther 1987)
850 (Denny 1985)
38 - 48 (Reddy and DeBusk 1987)
202 - 410 (Reddy and DeBusk 1987)
20 (in Gumbricht 1993)
1
3.6 - 12.9 (in Gumbricht 1993)
Elodea spp.
E. densa*
E. nuttallii*
Eriocaulon setaceum
Hydrilla verticillata
700 - 1000 (Lakshman 1987)
800 (Denny 1985)
1300 - 4700 (Lakshman 1987)
4.6 - 23 (in Gumbricht 1993)
6
2.8 - 12.9 (in DeBusk and Ryther 1987)
0.0 - 4.5 (Eighmy et al. 1987)
11.0 - 23.0 (Eighmy et al. 1987) - dry wt
4.2 - 10.4 (in DeBusk and Ryther 1987)
0.13 (Shardendu and Ambasht 1991)
40 - 2970 (Finlayson et al. 1980)
25 - 71, mean 54 (Shardendu and Ambasht 1991)
0.9 - 4.1 (Finlayson et al. 1980)
1.1 (Shardendu and Ambasht 1991)
0.19 (Shardendu and Ambasht 1991)
6.8 (Denny 1985)
25 - 69, mean 42 (Shardendu and Ambasht 1991)
1200 (Denny 1985)
0.10 (Shardendu and Ambasht 1991)
0.14 (Denny 1985)
60 - 6410 (Finlayson et al. 1980)
5 - 35, mean 22 (Shardendu and Ambasht 1991)
40 -70 (in Denny 1995)
60 (Denny 1985)
1.0 (Shardendu and Ambasht 1991)
1.0 (in Denny 1985)
4.0 (Greenway in press)
1.8 - 3.9 (Chambers et al. 1989)
1.1 - 3.8 (Finlayson et al. 1980)
3.3 - 11.0 (Denny 1985)
2
23.4 (in Brix 1994)
Water ribbons
Yellow bladderwort
Ribbon weed
FLOATING PLANTS
Azolla spp.
A. pinnata
Eichhornia crassipes*
Azolla
Water fern
Water hyacinth*
2.9 - 7.9 (in DeBusk and Ryther 1987)
10 - 33 (Reddy and DeBusk 1987)
128 - 135 (Reddy and DeBusk 1987)
1
25 - 2200 (Lakshman 1987)
2000 - 2400 (Reddy and DeBusk 1987)
460 - 21200 (Lakshman 1987)
2
6
10.0 (DeBusk and Reddy 1987)
24.2 - 64.4 (DeBusk and Ryther 1987)
6
6 - 18
(Reddy and DeBusk 1987)
1.4 - 12 (Reddy and DeBusk 1987)
1.4 - 12 (in Gumbricht 1993)
50 - 240 (in DeBusk and Ryther 1987)
49 - 243 (Reddy and DeBusk 1987)
1
SCIENTIFIC NAME
COMMON NAME,
HABIT
Hydrocotyle spp.*
H. umbellata*
Lemna spp.
Pistia spp.
P. stratiotes*
Penny wort
Duckweed
Water lettuce
Ricciocarpus natans
Salvinia spp.
Liverwort
Water fern
S. molesta*
S. rotundifolia*
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
16.4 -30.1(Reddy and DeBusk 1987)
16.4 - 30.1 (in Gumbricht 1993)
4.1 - 12.1 (Denny 1985)
8.2 - 16.4 (in Gumbricht 1993)
2000 - 2400 (in Gumbricht 1993)
5.7 (Aoyama and Nishizaki 1993)
1.4 - 8.0 (Johnston 1991)
700 - 1100 (in Gumbricht 1993)
10.3 - 29.7 (in DeBusk and Ryther 1987)
8.2 - 16.4 (Reddy and DeBusk 1987)
700 - 1100 (Reddy and DeBusk 1987)
2.0 - 13 (in Gumbricht 1993)
4.3 (Reddy et al. 1995)
2.0 - 12.5 (Reddy and DeBusk 1987)
2.3 - 7.5 (Reddy and DeBusk 1987)
PHOSPHORUS CONTENT
mg P /g dry wt
252 - 371 (Reddy and DeBusk 1987)
96 - 308 (Reddy and DeBusk 1987)
93.7 (in Brix 1994)
2
80 - 90 (DeBusk and Reddy 1987)
81 - 86 (Reddy and DeBusk 1987)
240 - 265 (Reddy and DeBusk 1987)
36 - 211 (Reddy and DeBusk 1987)
18 - 87 (Reddy and DeBusk 1987)
205 - 234 (Reddy and DeBusk 1987)
32 - 110 (Reddy and DeBusk 1987)
1
3.8 - 12.0 (in DeBusk and Ryther 1987)
1.6 - 7.1 (Reddy and DeBusk 1987)
1.6 - 7.1 (in Gumbricht 1993)
130 (Reddy and DeBusk 1987)
30 - 9000 (Lakshman 1987)
130 - 350 (in Gumbricht 1993)
13.7 - 21.9 (in Gumbricht 1993)
14.2 - 40.0 (in DeBusk and Ryther 1987)
13.7 - 21.9 (Reddy and DeBusk 1987)
600 - 1100 (in Gumbricht 1993)
600 - 1050 (Reddy and DeBusk 1987)
460 - 14600 (Lakshman 1987)
0.1 - 1.6 (Reddy and DeBusk 1987)
6
2 - 5.7 (Reddy and DeBusk 1987)
6
4000 - 6700 (Lakshman 1987)
200 - 300 (in Gumbricht 1993)
Water fern
2.5 - 12.3 (Reddy and DeBusk 1987)
2.5 - 12.3 (in Gumbricht 1993)
9.4 (Finlayson et al. 1982)
Water fern
6.4 - 13.9 (in DeBusk and Ryther 1987)
240 - 320 (Reddy and DeBusk 1987)
4.0 - 15 (Reddy and DeBusk 1987)
2 - 20 (in Gumbricht 1993)
10.9, max 18.4 (Greenway in press)
7.5 (Johnston 1991)
1.5 - 12 (in Gumbricht 1993)
1.5 - 11.5 (Reddy and DeBusk 1987)
7.3 (Greenway in press)
2
1
70 - 220 (DeBusk and Reddy 1987)
72 - 218 (Reddy and DeBusk 1987)
202 - 297 (Reddy and DeBusk 1987)
82 - 301 (Reddy and DeBusk 1987)
23.4 (in Brix 1994)
2
1
2
2 - 9 (in Gumbricht 1993)
6
0.4 - 2.4 (Reddy and DeBusk 1987)
1.4 - 4.7 (Finlayson et al. 1980)
5.7 (Greenway in press)
1.8 - 9.0 (Reddy and DeBusk 1987)
2.1 - 7.4 (Johnston 1991)
3.4 - 5.9 (in DeBusk and Ryther 1987)
Spirodela polyrhiza*
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
32 - 105 (Reddy and DeBusk 1987)
203 - 217 (Reddy and DeBusk 1987)
25.2 - 123 (Reddy and DeBusk 1987)
34 (Reddy and DeBusk 1987)
139 - 248 (Reddy and DeBusk 1987)
1
2
1
2
Wolffia spp.
Duckweed
EMERGENT PLANTS
Alternantha
philoxeroides*
Baumea teretifolia
B. articulata
B. rubiginosa
Blechnum camfieldii
B. indicum
Bolboschoenus fluviatilis
Brachiaria mutica
Caldesia oligococca
3.9 (Johnston 1991)
Sedge
Sedge
Sedge
Fern
Swamp water fern,
Creeping fern
Sedge
Para grass
Caldesia
2.4 - 3.7, max 8.7 (Greenway in press)
SCIENTIFIC NAME
Carex spp.
C. fascicularis
Colocasia esculenta*
Cyperus spp.
C. platystylis
C. difformis
C. involucratus*
C. papyrus
COMMON NAME,
HABIT
Sedge
Tassel sedge, Sedge
Wetland taro
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
370 - 3400 (Lakshman 1987)
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
PHOSPHORUS CONTENT
mg P /g dry wt
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
1.0 - 4.5 (Johnston 1991)
5
5.2 (in DeBusk and Ryther 1987)
2.5 - 12 (Reddy and DeBusk 1987)
0.2 - 6.3 (in Reddy and DeBusk 1987)
3
Sedge
Rice sedge, Sedge
2200
Papyrus, Sedge
(Hocking 1985) - above ground
8.8 (Hocking 1985) - leaves and
shoots
7.0 (Greenway in press)
5.2 - 26.5 (Hocking 1985)
22.9 (in DeBusk and Ryther 1987)
13.2 - 39.2 (Howard-Williams and Gaudet
15.2 (in Brix 1994)
1985)
34.2 (Denny 1985)
C. unioloides
Damasonium minus
Eleocharis spp.
Sedge
Starfruit
Spike rush
7.0 (Reddy and DeBusk 1987)
1.0 - 3.0 (Reddy and DeBusk 1987)
1.3 (Johnston 1991)
E. phillipinensis
E. ochrostachys
Fuirena umbellata
Hibbertia salisifolia
Ipomea aquatica*
I. diamantinensis*
Isolepis inundata
Emergent, Herbaceous
Emergent, Shrub
Water spinach
8.3 (in DeBusk and Ryther 1987)
max 9.5 (Greenway in press)
max 9.9 (Greenway in press)
14.6 (in DeBusk and Ryther 1987)
0.5 - 3.0 (Johnston 1991)
Juncus effusus*
Lepidosperma
longitudinale
Lepironia articulata
Ludwigia peploides subsp.
montevidensis
Marsilea mutica
M. drummondii
Monochoria cyanea
Myriophyllum spp.
M. variifolium
M. verrucosum
Napar advena*
Nelumbo lutea*
N. nucifera
Nymphaea spp.
Nymphoides crenata
Swamp club rush,
Sedge
Pithy swordsedge,
Sedge
Sedge
Water primrose
Nardoo
Nardoo
Monochoria
Watermilfoil
4.1 - 5.4, max 9.9 (Greenway in press)
10
7.0, max 9.8 (Greenway in press)
1.4 - 24.7 (in Gumbricht 1993)
120 - 1150 (Lakshman 1987)
50 - 900 (in Gumbricht 1993)
40 - 50 (Denny 1985)
6
7.7, max 13 (Greenway in press)
13 - 27 (in Gumbricht 1993)
120 - 330 (in Gumbricht 1993)
Red water milfoil
5.7 (in DeBusk and Ryther 1987)
10 - 160 (Reeder 1994)
Lotus
Water lily
Wavy marshwort
50 - 128
(Denny 1985)
2.2 - 6.1, mean 3.4 (Reeder 1994)
0.6 - 0.9 actual,
27 - 41 theoretical (Reeder 1994)
SCIENTIFIC NAME
N. indica
Ottelia ovalifolia
Panicum sp.*
Persicaria decipiens
Philydrum lanuginosum
Polygonum spp.*
Pontedaria spp.*
P. cordata*
Potamogeton spp.
P. tricarinatus
P. javanicus
Rorippa nasturtiumaquaticum
Sagittaria spp.
Sagittaria graminea
Schoenoplectus litoralis
S. mucronatus
S. validus
Schoenus brevifolius
Sphaerolobium vimineum
Triglohin striata
Typha spp.
COMMON NAME,
HABIT
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
Water snowflake
Swamp lily
PHOSPHORUS CONTENT
mg P /g dry wt
8.4, max 16.6 (Greenway in press)
0.9 - 2.9 (Reddy et al. 1995)
Slender knotweed
Frogsmouth
0.1 - 24.1
0.8 - 2.3 (Reddy et al. 1995)
0.5 - 0.9 (Reddy et al. 1995)
2.3 - 7.4 (DeBusk et al. 1995)
(DeBusk et al. 1995)
Pondweed
Floating pondweed
Pondweed
Water cress
200 - 930 (Lakshman 1987)
0.2 - 66.8 (DeBusk et al. 1995)
17 - 23 (Kirby and Albers in press)
6
0.5 - 4.5 (Johnston 1991)
Sagittaria
Sedge
Sedge
Sedge
streaked arrow grass
Cattail
T. orientalis
Cumbungi
T. domingensis
T. latifolia*
Cumbungi
0.8 - 4.0 (Greenway in press)
1800 - 3600 (Tanner et al. 1995)
2.2 - 16.7 (Reddy and DeBusk 1987)
2.2 - 16.7 (in Gumbricht 1993)
4.4 - 8.2 (Howard-Williams and Gaudet 1985)
12.3 - 19.2 (Denny 1985)
970 - 10140 (Lakshman 1987)
430 - 2250 (in Gumbricht 1993)
700 - 3600 (Hosomi et al. 1994)
5
4.5 - 37.5 (Reddy and DeBusk 1987)
690 - 1690 (Cooke 1992)
52.6 (in DeBusk and Ryther 1987)
40.0, mean 17.7 (in DeBusk and Ryther 1987)
11.8 (in Johnston 1991)
6.0 - 10.7 (Tanner et al. 1995)
10
0.5 - 4.2 (Reddy and DeBusk 1987)
1 - 4 (in Gumbricht 1993)
1.6 - 2.4 (Hosomi et al. 1994)
3.2 (Greenway in press)
6.5 - 9 (Breen 1990)
2.0 - 4.4 (Cooke 1992)
1.5 - 3.2, max 7.2 (Greenway in press)
0.5 - 5.3 (Johnston 1991)
21 - 110 (Reddy and DeBusk 1987)
8.2 (Hosomi et al. 1994)
1.7 - 17 (Mitsch et al. 1995)
10
3.6 (in Johnston 1991)
48.0 (in Brix 1994)
Villarsia exaltata
Herbaceous Plants
Cyperus procerus
C. trnervis
C. pilosus
C. exaltatus
C. haspan
C. lucidus
Eleocharis acuta*
E. sphacelata
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
Sedge
Sedge
Sedge
Tall flat sedge, Sedge
Sedge
Sedge
1.2 - 4.5 (Greenway in press)
1.3 - 3.9, max 9.4 (Greenway in press)
10
Tall spikerush
10
SCIENTIFIC NAME
E. spiralis
E. cylindrostachys
E. dulcis
E. equisetina
Fuirena ciliaris
Juncus spp.
J. prismatocarpus
J. polyanthemos
J. continuus
Leptocarpus tenax
Rumex brownii
Schoenus apogon
Smithia sensitiva
COMMON NAME,
HABIT
Emergent
Rush
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
PHOSPHORUS CONTENT
14.6 (Reddy and DeBusk 1987)
14.5 (in Gumbricht 1993)
9700 - 14000 (Lakshman 1987)
220 (in Gumbricht 1993)
4.0 (Reddy and DeBusk 1987)
2 (Reddy and DeBusk 1987)
2 (in Gumbricht 1993)
0.1 - 0.7 (Reddy et al. 1995)
5
mg P /g dry wt
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
30.1 (Reddy and DeBusk 1987)
Branching rush, Sedge
Sedge
Sedge
Sedge
Swamp dock
Fluke bogrush
Grasses
Diplachne fusca
Echinochloa inundata
E. colona
E. crus-galli
Hymenachne acutigluma*
Oryza meridionalis*
Paspalum distichum
Pennisetum alopecuroides
Phragmites spp.*
P. australis
P. communis*
Pseudoraphis spinescens
Xanthorrhoea fulva
Brown beetle grass
Marsh millet
Awnless barnyard
grass
Barnyard grass
Water couch
Swamp foxtail
Reed
Common reed
Common reed
Spiny mudgrass
Swamp grass tree
Shrubs
Acrostichum speciosum
Baeckea stenophylla
B. diosmifolia
Banksia robur
3.0 (Finlayson 1991)
1.8 (Finlayson 1991)
3.4 (Greenway in press)
5.7 (Finlayson 1991)
1.4 (Finlayson 1991)
Mangrove fern, Fern
Weeping baeckea
Fringed baekea
Broad leaved banksia
5
5
2.7 - 16.4 (Reddy and DeBusk 1987)
2.7 - 16.4 (in Gumbricht 1993)
12.3 - 19.2 (Denny 1985)
600 - 3700 (in Gumbricht 1993)
1500 - 3000 (Hosomi et al. 1994)
25.9 (in DeBusk and Ryther 1987)
26.1 (Hocking 1989)
9960 (14945) (Hocking 1989)
2.0 - 3.0 (Reddy and DeBusk 1987)
2 - 3 (in Gumbricht 1993)
1.7 - leaves (Hocking 1989b)
1.7 - 3.8 (Hosomi et al. 1994)
6.1 - 9.4 (Hocking 1989a)
1.4 - 2.0 (Greenway in press)
1.8 (Johnston 1991)
1.3 (Finlayson 1991)
4
1
820 - 12290 (Lakshman 1987)
5.2 (Finlayson 1991)
1.4 - 5.3 (Reddy and DeBusk 1987)
5
5
9.6 (Reddy and DeBusk 1987)
8.2 (Hosomi et al. 1994)
12.2 - 45 (Hocking 1989a)
45.7 (in Brix 1994)
SCIENTIFIC NAME
Boronia falcifolia
B. parviflora
Callistemon pachyphyllus
Comesperma defoliatum
Gahnia sieberiana
Hibbertia salicifolia
Leptospermum liversidgei
L. semibaccatum
L. juniperinum
Melastoma affine
Pultenaea paleacea
Todea barbara
COMMON NAME,
HABIT
MEAN ANNUAL
PRODUCTION
g dry wt/m2/d
MEAN ANNUAL STANDING
STOCK
g dry wt/m2
PHOSPHORUS STORAGE
POTENTIAL
mg P/m2
PHOSPHORUS CONTENT
mg P /g dry wt
PHOSPHORUS UPTAKE
RATES
mg P/m2/d
Swamp boronia
Wallum bottlebrush
Red fruited sawsedge
Emergent shrub
Blue tongue
Southern king fern,
Fern
Viminaria juncea
Trees
Callistemon viminalis
Eucalyptus robusta
Glochidion ferdinandi
Hibiscus diversifolius
Lophostemon suaveolens
Melaleuca alternifolia*
M. quinquenervia
Swamp messmate
Cheese tree
Swamp hibiscus
Swamp mahogany
11 - leaves (Bolton and Greenway 1995)
1.5 - litter (Greenway, in press)
Paperbarked tea tree
34 - 1061 Above ground
1460 - 4580 Below ground
Combination of :
Pontedaria, Saggitaria,
Panicum and Hibiscus
(in Dolan et al. 1981)
7
5 - 4.5 Above ground
1.3 - 4.6 Below ground
(in Dolan et al. 1981)
0.6 - 29.5 (in Dolan et al. 1981)
7
PERIPHYTON
Periphyton and
Phytoplankton
Phytoplankton
10 - 35, mean 21.2
(DeBusk et al. 1995)
3.7
(DeBusk et al. 1995)
0.6 - 0.9 (in Mitsch et al. 1995)
79 - 101 (DeBusk et al. 1995)
27.4 (Reeder 1994)
8
7
Notes :
* - indicates a plant not selected by the QDPI as an appropriate wetland plant species.
1 - calculated using the growth rate in the linear phase of the plant’s growth and the average tissue phosphorus content. This represents phosphorus removal due to the plant alone.
2 - calculated using the phosphorus concentrations in the water and represents phosphorus removal due to the plant uptake and nutrient transformations.
3 - values represent predominantly above ground storages.
4 - Concentration includes above ground and below ground biomass (roots, young and mature rhizomes, and live, young and dead shoots)
5 - Represents above and below ground biomass production.
6 - Represents above ground biomass production only.
7 - Range of values between a control plot and a plot receiving high effluent waters.
8 - 79 mg P/m2/d was based on algal biomass only, the 101 mg P/m2/d included phosphorus adsorption to water borne particles.
9 - Whole plant production by the minimum-maximum method with the corrected estimate value in parentheses.
10 - Range includes nutrient accumulation in selected native Australian wetland species from constructed and natural wetlands (lowest phosphorus values in the constructed wetland
species).